UNIVERSITY OF COPENHAGEN FACULTY OF SCIENCE

Master Thesis Lasse Gottlieb

Woodland grazing Effects of horse grazing on ground vegetation and forest structures

Supervisor: Rita Merete Buttenschøn

Co-supervisor: Niels Worm, Danish Nature Agency Submission: 29/07/2015

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Name of department: Department of Geosciences and Natural Resource Management

Forfatter: Lasse Gottlieb

Title / Subtitle: Woodland grazing – effects of horse grazing on ground vegetation and forest structures

Vejleder: Rita Merete Buttenschøn, Institut for Geovidenskab og Naturforvaltning, Københavns Universitet

Medvejleder: Niels Worm, Naturstyrelsen

Afleveret den: 29. juli 2015

Antal studieenheder: 60 ECTS

3 Content 1. Abstract ...... 6 2. Introduction ...... 6 2.1. Effects of horses in relation to primary production ...... 8 3. Study area ...... 11 3.1. Gribskov in a historical context ...... 12 3.2. Study sites ...... 13 3.2.1. Kollerup Enghave ...... 13 3.2.2. Store and Lille Hessemose ...... 14 3.2.3. Sandskredssøen ...... 14 4. Methods ...... 16 4.1. Ground vegetation...... 16 4.2. Forest structures...... 16 4.3. Data analysis ...... 18 5. Results ...... 20 5.1. Ground vegetation...... 21 5.1.1. Species richness and density ...... 21 5.1.2. Implications of abiotic factors ...... 24 5.1.3. Spatial heterogeneity ...... 28 5.1.4. Diversity ...... 29 5.1.5. Species composition and abundance ...... 30 5.2. Forest structures...... 35 5.2.1. Species richness and regeneration of woody species ...... 36 5.2.2. Browsing impact ...... 40 5.2.3. Insect pollinated woody species ...... 42 5.2.4. Deadwood...... 43 5.2.5. Other forest structures; epiphytic bryophytes and lichens, hollowness, decay and woodpecker holes ...... 44 5.3. Forest condition ...... 45 6. Discussion ...... 46 6.1. Ground vegetation...... 47 6.1.1. Species richness and density ...... 48 6.1.2. Species specific abundance alterations ...... 57

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6.2. Forest structures...... 61 6.2.1. Tree regeneration and browsing impact ...... 62 6.2.2. Bark stripping ...... 68 6.2.3. Browsing and grazing induced succession alterations ...... 69 6.2.4. Other forest structures and the biodiversity ...... 72 6.3. Management ...... 76 7. Conclusion ...... 78 References ...... 80 Appendix 1 – Ground vegetation ...... 86 Appendix 2 – Seedlings and saplings of woody species ...... 97 Appendix 3 – Trees >2 m ...... 108 Appendix 4 – Deadwood ...... 112

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1. Abstract The lack of natural dynamics and the intensive forestry management over the past centuries have had devastating consequences for the biodiversity found in woodlands in Denmark and the rest of northwest Europe. Woodlands are the ecosystem in Denmark that contains most species, but a high proportion of all the red-listed species are also associated with woodlands, particular either in the form of veteran trees and deadwood or with woodland glades and woodlands with enough light to develop species rich ground vegetation. Grazing management with livestock may affect these structures beneficial for the woodland biodiversity. However, examples of grazing management with domesticated herbivores in woodlands are few and hence the knowledge is mostly theoretical. Thus, there is a lack of the understanding of how large herbivores influence the woodland dynamics. In this study the short-term effects of woodland grazing with Icelandic horses on the ground vegetation and forest structures were investigated in three different woodland habitats (alder swamps, oak forest and beech forest) in Gribskov, Denmark. After just a few year of management the richness of vascular plants species were found increased in the alder swamps and the oak stands while the effects appeared negative in the beech stands. The species colonizing the grazed areas were along with those species which increased abundance found to be more tolerant toward disturbances. No effects were found on the richness of woody species. However, the horses had high browsing impacts on Fagus sylvatica, Alnus incana, Euonymus europaeus and Carpinus betulus but apparently without a higher impact in areas grazed throughout the year. Though the browsing by the horses are impeding the saplings of the browsed species, and in a single case reversing the succession in a dense beech stand, it is yet not possible to conclude whether and to which degree the horses are able the delay the succession.

2. Introduction With the change of The Danish Forest Act in 2004 up to 10 per cent of an individual area designated as forest-preserved land may be used for grazing with domesticated large herbivores (Danish Forest Act, 2004). Prior to this, woodland grazing was forbidden in Denmark for about 200 years, resulting in a division of the landscape into agricultural and forestry land with distinct boundaries between land uses. The consequence of the 200-years absence of domesticated herbivores was an increased tree density. This happened through enhanced growing conditions, if necessary by drainage, but also by plantation of all open parts of the woodland (Ejrnæs et al., 2011; Johannsen et al., 2013). The last 200 years of increasingly dense, dark woodlands, which are optimized for forestry production of straight trunks, have had a major negative impact on the woodland biodiversity (Ejrnæs et al., 2011). Today the conservation status for all 10 woodland habitat types represented in Denmark (Council Directive 92/43/EC), are rated as highly unfavourable and without any prospect of improvement within the next 12 years (Fredshavn et al., 2014b).

Looking at biodiversity, woodlands are the ecosystem in Denmark that contains most species (Ejrnæs et al., 2011), and of all the red-listed species in Denmark, about 65% are associated with woodlands and nearly 40% are exclusively found here (Bruun & Heilmann-Clausen, 2012). A third of the total woodland biodiversity, including several of the species protected by the Habitat Directive (Council Directive 92/43/EC) which live in woodlands, are typically associated with either old hollow trees and deadwood or woodland glades and woodlands with enough light to develop a species rich understory of shrubs and herbs (Ejrnæs et al., 2011; Fredshavn et al., 2014b). The rare and vulnerable species associated with these habitats are often so fragmented and dispersed that their genetic exchange and dispersal opportunities are limited

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(Johannsen et al., 2013; Fredshavn et al., 2014b). This increases the risk for local extinction as a result of inbreeding, interspecific competition with other species or stochasticity. Many populations of the red-listed species are of these reasons so isolated, that they risk extinction even when their current habitats are protected (Johannsen et al., 2013).

The most important factor of influence for the highly unfavourable conservation status for the woodland habitats in Denmark is the intensive forestry management with plantation, draining, thinning, plantation of woodland glades, felling of biological young trees and the removal of deadwood (Fredshavn et al., 2014b). Johannsen et al. (2013) compared different mechanisms to promote the biodiversity in Danish woodlands, and found that old-growth forest with a natural high grazing intensity was the best.

Within the last decades increasing focus on biodiversity in woodlands has led to a change in the management regime towards a more sustainable management, especially in the state-owned woodlands (Ejrnæs et al., 2011). Today there is a renewed interest in woodland grazing with domesticated ungulates, and the Danish government has a wish for more woodland grazing (Miljøministeriet, 2014). In Gribskov, the northern part of Zealand, the management regime has also changed within the last couple of years from a focus on timber production to an increased focus on biological values and recreation (Overballe-Petersen et al., 2014). Here several enclosures have been established within the last years with grazing Icelandic horses.

However, the impact of large herbivores is complex and depends among other things on grazing/browsing behaviour and selection of the herbivore species, the grazing intensity as well as the individual plant species responses and the existing plant community (Rosenthal & Kotanen, 1994; Olff & Ritchie, 1998; Hester et al., 2000; Kuiters et al., 2006; Vulliamy et al., 2006; Buttenschøn, 2007). Furthermore, the effects vary greatly both spatially and temporally (Huntly, 1991). Hence, the effect on the vascular plant species composition, as well as the overall biodiversity is highly a matter of these factors, and the precise effect on the biodiversity may therefore be unpredictable.

In northwest Europe, examples of long-term grazing management with domestic herbivores in woodlands are scarce and mainly restricted to wood-pasture remnants (Mitchell & Kirby, 1990; Spencer, 2001; Gill, 2006; Van Uytvanck & Hoffmann, 2009). Hence, the knowledge of the effects is mostly theoretically, testimonials from the past or studies of the regeneration and establishment of woody species in grazed open-land pastures. The scientific studies of the effects of domesticated herbivores in woodland systems in northwest Europe are few. Even fewer are the studies of the effects of horse grazing in these habitats. Thus, there is a lack of predictive understanding of the relative importance of larger herbivores in woodland dynamics, e.g. the processes determining browsing patterns, plant responses at the population and landscape scale, and herbivore-plant-soil interactions driving vegetation change (Gill, 2006; Skarpe & Hester, 2008). Attempting to achieve the management objectives besides gaining scientific knowledge, it is therefore necessary to monitor and evaluate the effects of the management. This is indeed the purpose for this thesis; to evaluate the effects of horse grazing in Gribskov on the ground vegetation and forest structures.

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2.1. Effects of horses in relation to primary production Large herbivores, wild as well as domestic, can play an major role in shaping the woodland environment with a significant impact on woodland structure, plant species composition and primary production by creating spatial, structural and temporal heterogeneity and modulate the succession (Putman et al., 1987; Mitchell & Kirby, 1990; Huntly, 1991; Hobbs, 1996; Hester et al., 2000; Vera, 2000; van Wieren & Bakker, 2006). Large herbivores may have a direct and indirect impact on tree seedling establishment and sapling recruitment, as well as on ground vegetation through the removal of plant material, selection of forage and hence, alteration of vegetation structure and light conditions, trampling, deposition of urine and faeces, and zoochorous seed dispersal (Hester et al., 2000). Grazing in woodland systems can re-establish more open and diverse woodland types, with smaller glades and inner fringes. It often allows a much more dynamic development with natural gradients between woodlands and meadows or pastures than the often distinct boundaries between woodland and arable land which dominates today. The gradients are reflected in varying degrees of light and humidity in many possible combinations and in resulting heterogeneous habitats, of which some are extremely rare in the present Danish landscape (Hobbs, 1996; Buttenschøn, 2007; Bruun & Heilmann-Clausen, 2012). Grazing by large herbivores in woodlands may provide a greater diversity in vegetation structure and species composition than the absence of grazing. This may create conditions for the highest diversity of both flora and fauna (Mitchell & Kirby, 1990).

Like in many other terrestrial ecosystems, the species composition of vascular plants is determined by a vast number of conditions, e.g. the edaphic conditions, humidity, light, nutrients, continuity, anthropogenic influences in past and present and the interspecific competition between the species. The special feature for the woodland is however, that a substantial part of these conditions is controlled or influenced by the trees (Skarpe & Hester, 2008), especially the amount of light reaching the forest floor (Petersen & Vestergaard, 2006; Friis Møller, 2010). Hence, the impacts on regeneration and establishment of woody species are, in the long-term, of fundamental importance for the richness and community of the ground vegetation. But the impacts are, however, highly dependent on the herbivore species, their grazing behaviour and selectivity, grazing intensity and the present vegetation composition (Rosenthal & Kotanen, 1994; Olff & Ritchie, 1998; Hester et al., 2000; Kuiters et al., 2006). With different preferences large herbivore species will vary in their impacts on woodland environments, with different tree species, ground flora, growth structure, and, ultimately, different forest types being favoured by each (Mitchell & Kirby, 1990; Latham, 1999).

Ungulates are traditionally ordered along a continuum of functional types from “grazers” to “browsers”, according to food plant selection, diet composition and digestive anatomy (Hofmann, 1989). Here horses fall within the category “browser”. Although browsers by definition consume woody plants, grazers also consume some woody plant material. The main differences are that browsers actively select woody material whereas grazers may consume it as a part of the sward, and grazers seems to be less selective of particular woody species than browsers (Hester et al., 2000). In a comparative study of the annual diet composition of large herbivores, Van Dyne et al. (1980) found horses, along with cattle to be mainly grass eaters (grasses and grass-like plants, particular sedges) (Figure 1). Studies from New Forest, England, found an overlap in the habitat use between cattle, deer and horses, but even though there was an overlap in habitat use, their diet differed significantly throughout the year (Putman, 1986; Putman et al., 1987).

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100 Shrubs/ Trees n o i t

i 80

s Forbs o p m

o 60 c

t e i d

l

a 40

n Grasses u n n a 20 %

0 e e p t e n r r r r s tl e a s o e e e e r t e o o s e e e e o a h G o i d d d d H C S B e a d M o k w e i llo R S a R F Figure 1: Mean annual diet composition (%) of some herbivore species (after Van Dyne et al., 1980).

Size of teeth and the morphology of mouthparts are important determinants of the degree of selectivity exhibited (Crawley, 1997). Horses compared to cattle have both upper and lower incisors and, hence, can grasp the grass between their teeth to tear it off (Van Dyne et al., 1980). Furthermore, the two sets of opposing facing incisors also make them effective at cutting even quite fibrous vegetation (Mayle, 1999). In contrast, cattle use their tongues as prehensile organs. The herbage is pinched between the tongue and the lower teeth and torn off (Van Dyne et al., 1980). It means that horses are able to graze much closer to the ground (Mayle, 1999). Hence, horses are more selective than cattle in regard to plant species, while the selective behaviour of cattle is more directed toward plant communities.

Differences in the digestive system (ruminant and hindgut digester) will also dictate large herbivores’ preferred food and where they will feed. Both leaves and to a higher degree stems contain considerable amounts of cellulose. No multicellular animal is known to produce an enzyme that is successfully able to break down cellulose, thus, any herbivore subsisting on a fibrous diet must therefore enter into some form of symbiotic association with cellulase producing microorganisms, and must provide a fermentation chamber within the digestive tract in which the breakdown of cellulose can take place (Janis, 1976). In ruminants, the fermentation chamber, the rumen, lies before the small intestine. Cud chewing and further fermentation divides the forage to particles with a size suitable to enter the reticulum and omasum. As the fibre content of the forage increases it remains in the rumen for a longer time, since it takes longer to reduce the particles to a suitable size, and thus the ruminant’s intake is depressed (Figure 2) (Janis, 1976; Van Dyne et al., 1980). Intake in ruminants is largely dependent on the volume of the rumen, as they are not able to increase their mass flow rate. Hence, there is a definite cut-off point in the percentage of fibre in the diet that they can tolerate, beyond which they will be unable to support itself (Janis, 1976). In horses, on the other hand, which are monogastic herbivores, the fermentation takes place in the caecum (between the small and large intestines), while the absorption of the fermentation products occur in the enlarged colon (Janis, 1976). Thus, the digestive anatomy of horses makes them able to compensate for a fibre-rich

9 diet by increasing the mass flow rate and thus the forage intake (Figure 2). This strategy may be essential for their use of herbage above a given fibre/protein ratio (Janis, 1976).

Figure 2: Comparison of ruminant and horse feeding strategies depending on the type of herbage. I) Both ruminant (cattle) and horses do equally well, though both select different layers of herbage. II) Horse continues to do well, but ruminant cannot maintain itself on this type of herbage (Janis, 1976).

Horses are not in the same way as ruminants able to break down toxic secondary plant metabolites before the actual nutritional uptake. The stomachs of ruminants are not acidic as in horses but alkaline with a diverse bacterial and protozoan flora which are capable of degrading a wide variety of secondary plant compounds (Freeland & Janzen, 1974). Thus, horses are generally more sensitive to toxic plants than for example sheep and cattle (Buttenschøn, 2007). The degree of sensitivity in combination with the concentration of plant secondary metabolites will influence the choice of diet (Gill, 2006) and can affect the impacts difference larger herbivores have. As an example was the browsing effect of cattle and horses compare in a study from Oostervaarderplassen (Cornelissen & Vulink, 2001). During the study Sambucus nigra invaded both the cattle and horse grazed area, but contrary to the area grazed by cattle, Sambucus nigra increased its abundance in the horse grazed area much more. This was probably caused by the fact that horses, as hind-gut fermenters, cannot overcome the effects of the secondary compounds in

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Sambucus nigra, whereas cattle, as ruminants, to a certain extent can detoxify the cyanogenic glycocides in Sambucus nigra (Freeland & Janzen, 1974). Nevertheless, there are also examples of ruminants tolerating certain plant species worse than horses because plant chemical substances are decomposed into toxins during rumination (Buttenschøn, 2007). However, the role of different plant secondary compounds in different herbivore species is still not well known (Gill, 2006).

However, the preference for plant species is not fixed. The horse’s forage selection pattern is evidently modified by the availability, palatability and digestibility of the forage, which again depends on a bundle of factors (Van Dyne et al., 1980; Holtmeier, 2015). The same forb or shrub species may contribute to one third of the diet in one habitat and may be neglected in the other (Kolter et al., 1999). Hence, horses exhibit a general adaptability in relation to seasonal and habitat related variations in the diet composition (Van Dyne et al., 1980; Putman et al., 1987), where especially the smaller hardier breeds like Icelandic, Konik, Dartmoor and Shetland are more adaptable to a range of different environmental conditions (Tolhurst & Oates, 2001) and tend to be less selective, taking poorer quality forage than other breeds (Mayle, 1999). Within extensive systems, browse material may be important dietary component, particularly during winter months. As an example from New Forest, horses diet showed a pronounced seasonal variation where more than 45% of the diet in February was constituted by woody species - especially considerable quantities of Ulex europaeus and Ilex aquifolium browse being consumed (Putman et al., 1987). However, data of the mean annual diet of several large herbivore species (Figure 1) can serve as a useful summary for the comparison of herbivore diets. But ideally, a selected diet should always be expressed in terms of the available diet.

3. Study area Gribskov is placed in the northern part of Zealand, Denmark, and is the only place in the country where such a huge area has been covered with woodland for 10,000 years. This is due to a combination of the local geological conditions and the culture-historical development (Rune, 2009; Overballe-Petersen et al., 2013a; Overballe-Petersen et al., 2013b). Gribskov is a 5,600 ha cultural woodland, which naturally would consist of deciduous species (Overballe-Petersen et al., 2014). Today Gribskov consists of a mosaic of managed, monoculture stands with 36% conifers (mainly Picea abies), 54% broadleaved trees (mainly Fagus sylvatica, Quercus sp. with some Betula sp. Fraxinus excelsior and other minor types), 3% mires and 7% non-forested areas (Overballe-Petersen et al., 2014). Gribskov is today a part of the European Natura 2000 network.

Four species of large herbivores do naturally occur within Gribskov. The woodland is the core-area for the largest population of fallow deer (Dama dama) in Denmark with just over 1000 individuals. Especially in the wintertime they live in the northern and central part of the woodland where their rutting places are found and where winterfeeding of the population takes place. In the summertime they spread across most of the woodland (Rune, 2009). Gribskov houses a roe deer (Capreolus careolus) population of approximately the same size as the fallow deer population. In contrast, they are scattered throughout the woodland and like to stay near the woodland fringes, giving them opportunities to forage at the open land (Rune, 2009). A population of around 50 individuals of sika deer (Cervus nippon) also live in Gribskov, though mainly in the

11 central part where they as the fallow deer have their rutting places (Rune, 2009). A population of around 370 individuals of reed deer (Cervus elaphus) lives in the northern part of Zealand as well (Kanstrup et al., 2014), though they are only occasionally found within Gribskov, primarily in northern part (Kanstrup, 2013).

3.1. Gribskov in a historical context A number of factors have influenced the biodiversity found in Gribskov today. Even though the impact has been at a minimum, smaller areas in Gribskov were already with the beginning of the early Neolithic farming culture (2800-3900 BC) deforested for cultivation of wheat and barley and the introduction of livestock grazing. From medieval time deforestation was speeded up. Even though timber was an important resource, the main value at that time was the areas “between the trees” where livestock grazing, mowing and grain production were possible (Rune, 2009).

Gribskov has a long history with horse grazing which makes the (re-)introduction of the horses today even more interesting. The period started in the middle of the 16th century, where the first proper grazing pastures within Gribskov were designated in the northern part of the woodland. The king showed personal interest for the horse breeding and the grazing pastures within Gribskov, and around 50 years later, the pastures were expanded and a one kilometre wide border along the west coast of lake Esrum Sø (Figure 3) was laid out as stud pastures for the royal horses (Rune, 2009). Often it was only the stallions that were stabled during wintertime while mares and foals grazed all the year round in the pastures with minor fodder supply during the winter. No descriptions of the pastures in Gribskov from the first 200 years exist today, but in the 18th century, the greatest period for the horse studs in Gribskov, the pastures were described as remaining naturally without draining ditching or cultivation but with widespread ponds, bogs, stone and thickets. The vegetation was some places with Calluna vulgaris, thus it has not been a rich soil (Rune, 2009). Towards the end of the 1770s the Royal Danish horse breeding changed drastically. The stud purebreds for colours during many generations had reduced the mares’ fertility to less than half over large parts of Gribskov, and the resistance to diseases was decreasing. Although the number of horses was nearly 700, only 50 foals were produced yearly (Rune, 2009). The horse stud was expensive to keep running without any particular income, and in 1799 the pastures were formally handed over to the forestry authorities for silviculture. However the last horses first left Gribskov several years later. At the handover, around 55% of the pastures were regarded as forest, 25% as agriculture and 20% as meadow (Rune, 2009). The areas were thus predominantly forested, but obviously not with a canopy as dense as seen today.

In the 17th century the trees were cut to the ground faster than the natural regeneration. For centuries the ground vegetation had never experienced as good growing conditions as these and hardly had animals previously had such good grazing opportunities. By the end of the 17th century, a period often described as the darkest in Danish woodland history, Gribskov was a paradise for flora and fauna. The volume of timber was at that time less than a quarter of what is found today (Rune, 2009). In the 18th century woodland glades and meadows and several grazing pastures were turned over to afforestation, flower covered woodland floors in the deciduous woodland were turned into vegetation-less dark coniferous forest, mainly of the introduced Picea abies. During the 19th and 20th centuries this management planning and conifer cultivation continued and increased (Rune, 2009; Overballe-Petersen et al., 2014). A seriously decreased amount of grazing, woodland mires and decaying wood were the results (Rune, 2009). Hence, the past 200 year woodland management has been the determining factor for the present diversity and survival of organisms including the ground vegetation. Although, the individual woodland cultivation interventions’

12 impacts on diversity are often quite poorly studied, the determining factors includes: choice of tree species and the continuity, stand size and mosaic, degree of thinning, soil tillage and genetic origin of the trees, the amount of deadwood and the number of standing trees in natural decay, drainage of wetlands and the number and size of open habitats, nutrient cycles and the impact on this from air and water, and last but not least invasive species (Rune, 2009).

Within the last years the focus timber value is to some extent replaced by an increased focus on biological values and the management is in transformation from a clear-felling regime to near-natural forestry with focus on continuous tree cover, natural regeneration and a change in dominant tree species from conifers to deciduous trees (Overballe-Petersen et al., 2014). The history of the present plant species however extends centuries back in time and their presence in an area are in many cases a matter of the culture- historical influence. The huge variation in Gribskov, from the nutrient-poor sandy soils in the central woodland to the fertile soils at the edges, provides an opportunity for a great diversity. In the period 1991- 96 nearly 700 different vascular plant species were registered in Gribskov – almost half of all wild vascular plant species found in Denmark. Of the species found in Gribskov, approximately 12% are rare – not necessarily at a national level, but at least locally uncommon, and most of these species have that in common that they require attention in the future management for survival at their present locations (Rune, 2009).

3.2. Study sites Gribskov has several areas grazed by Icelandic horses of which four areas were chosen as study areas (Figure 3). The grazing management is relatively new for all areas and has been carried out for a continuous period varying between two and four seasons. All grazed woodland areas are in the vicinity of, and within the same enclosure as open grassland habitat – mostly meadows. Forest dominated by Quercus robur (henceforth oak) (Code 9160 and 9190 in the Habitat Directive: Council Directive 92/43/EC), by Fagus sylvatica (henceforth beech) (Code 9110 in the Habitat Directive: Council Directive 92/43/EC) and alluvial forest with Alnus glutinosa (Code 91E0 in the Habitat Directive: Council Directive 92/43/EC) where chosen as study sites as more of the grazed areas in Gribskov have these habitats in common. Furthermore, in a number of areas the fence intersects these habitats, and hence the impact of the grazing management can be investigated directly while other factors are assumed to be kept similar. For those areas not regarded as a habitat type in relation to the Habitat Directive (Habitat Directive: Council Directive 92/43/EC), mostly due to the plantation character with equally aged monocultures in rows, the habitat type codes were assigned relative to the assumed development potential with help from Fredshavn et al. (2011). In addition woodland swamps with Alnus incana were regarded as alluvial forest with Alnus glutinosa (Code 91E0 in the Habitat Directive: Council Directive 92/43/EC).

3.2.1. Kollerup Enghave The enclosure covers a 14.3 ha large, quite nutritious and spring influenced area. The 7.6 ha meadow located in the middle of the area is one of the largest in Gribskov and has from old time served as pasture for cattle, though the grazing regime was abandoned in 1990 (Rune, 2009). After this period the meadow has been mown yearly, but without removal of the dead plant material produced (Worm, pers. comm.).

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Since spring 2013 the area has been re-grazed by 15 Icelandic horses from end of April until end of October, but also several of the adjacent forested stands have since 2013 been within the enclosure. Within the fenced area a beech stand (Code 9110 in the Habitat Directive: Council Directive 92/43/EC) afforested in 1888 and intersected by the fence is found. The grazed part is 0.5 ha large and the ground vegetation is dominated by Deschampsia flexuosa and Carex pilulifera. Also a 1.5 ha large oak stand (Quercus robur) afforested in 1961 and dominated by Deschampsia cespitosa, Stellaria holostea and Oxalis acetosella, and a 1.3 ha large alder swamp (Code 91E0 in the Habitat Directive: Council Directive 92/43/EC), which has been alder swamp since the 19th century (Rune, 2009) with the last chopping in 1924, is found. The alder swamp is dominated by Urtica dioica, Phalaris arundinacea, Scutellaria galericulata and Galium palustre.

3.2.2. Store and Lille Hessemose Store and Lille Hessemose constituted until the 19th century one of the largest mires in Gribskov with areas of respectively 11.5 ha and 7.5 ha. The northernmost part of Store Hessemose was until the middle of the 19th century a rather dense alder swamp with large coppice stools and root suckers. Store Hessemose became strongly drained with more than 5 km draining ditches and the entire area was used for mowing until the 1890s where afforestation took place with Betula sp., Alnus glutinosa and Alnus incana, though several parts apparently still were too wet and remained open with meadow-vegetation, likewise did Lille Hessemose (Rune, 2009). The enclosures in Store and Lille Hessemose respectively cover an area of 5.8 ha and 9.2 ha, and the areas have been grazed since 2012 from end April until end of October by respectively 15 and 9 Icelandic horses. However, Store Hessemose was grazed throughout the year the first year of management. In Store Hessemose meadow constitutes the main fenced area, but also a beech stand (0.7 ha) afforested in 1967 (almost without any ground vegetation) and an alder swamp (0.3 ha) afforested in 1943 and dominated by Deschampsia cespitosa and Poa trivialis are found within the enclosure. The fence intersects the beech stand, and hence the grazing effect can here be studied directly. Another alder swamp is found just outside the enclosure that also was afforested in 1943, hence, other factors than grazing are regarded equal. In Lille Hessemose an oak stand (0.5 ha), afforested in 1961 and dominated by Deschampsia cespitosa, Stellaria holostea and Oxalis acetosella, is found. The stand is intersected by the fence.

3.2.3. Sandskredssøen The central lake in the area was originally a nutrient-poor bog without any tree vegetation. But already from the middle of the 19th century the bog was drained and afforested with spruce. In the 1990s it was decided to clear-cut the spruce stand and re-establish the former wetland. Thus, the trees were cut down in 1996, the drainage ditches were blocked, and the water level began to rise. Originally it has never been a lake, but due to decomposition of the peat, fatal shading of the sphagnum mosses and the weight of the trees, a lake is found today (Rune, 2009). The spruce stands around the lake has subsequently been deforested, the first part in 2006, to create an open area which visualizes the moraine landscape (Rune, 2009). The enclosure covers a total area of 45.6 ha where Icelandic horses since spring 2011 have grazed year-round – approximately 22 individuals in the summertime and 15 in winter. Within the enclosure a 0.5 ha oak stand (Code 9160 in the Habitat Directive: Council Directive 92/43/EC) afforested in 1818 and a 2.8 ha oak stand afforested in 1983 are found. The old oak stand is dominated by Deschampsia cespitosa, Rubus idaeus, Calamagrostis epigeios and Oxalis acetosella whereas the young is dominated by Deschampsia flexuosa and Carex pilulifera. A 1.6 ha beech stand, afforested in 1966 is also found within the enclosure. It is dominated by Deschampsia flexuosa and Carex pilulifera, but also by seedling and saplings

14 of Picea abies. Both the old oak stand and the beech stand are intersected by the fence and the grazing effect can directly be studied. The young grazed oak stand is compared to another young oak stand (afforested in 1961) found nearby.

Figure 3: The four study areas in Gribskov; Store Hessemose (SH), Lille Hessemose (LH), Kollerup Enghave (KE) and Sandskredssøen (S).

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4. Methods The field studies were carried out in the period July 7th 2014 to October 1st 2014. All methods used were based on the guidelines by Fredshavn et al. (2014a) for monitoring terrestrial habitats within the European Natura 2000 network.

4.1. Ground vegetation To measure the species richness of vascular plants five circular quadrats with a radius of five meters (henceforth 5 m circle), each covering an area of 78.5 m2, were laid out at each site. The quadrats were distributed randomly, but in a way that the whole study-area was attempted covered. The distance between the centres of two quadrats was at least 10 meters to prevent any overlap. Within each 5 m circles all species of vascular plants were identified. To make an objective measurement of the density of species and to measure the frequency of each species, 25 smaller circular quadrats, each with an area of 0.1 m2 (Raunkjær circle), were distributed randomly within each 5 m circle. In each Raunkjær circle all species of vascular plants were identified as well. Only individuals rooted within the 5 m circles as well as the Raunkjær circles was included. When identification to species level was impossible, e.g. because the individual was only found vegetative or withered, it was identified to as low a taxonomic level as possible. Identification was done with help from Frederiksen et al. (2012), Mossberg and Stenberg (2003), Schou et al. (2014), Schou (2006), Schou et al. (2010), Faurholdt and Schou (2012) and Poland and Clement (2009).

4.2. Forest structures Both vegetation heights and light conditions were measured in each 5 m circle. The vegetation height was given as the height of the lowest growth layer consisting of vascular plants in a continuous growth layer. A layer of woody species clearly separated from the lowest growth layer was not included in the measuring (Figure 4). If the vegetation was lying flat, e.g. due to trampling or influence from wind or precipitation it was tried re-raised before measuring. The vegetation height was noted in intervals of 1 cm when the vegetation was less than 15 cm, in intervals of 5 cm when the vegetation was 15-30 cm and in 10 cm intervals when the vegetation was above 30 cm. The vegetation height was measured in four places, each one meter from the centre of the 5 m circle, at a white board with the different intervals as black lines. When 50% or more of a line was free of vegetation and hence visible when viewed horizontally at a distance equally to an arm’s length, it was regarded as the vegetation height (Figure 4). If 50% or more of the lower edge of the plate was visible when the plate touched the ground, the vegetation height was equal to 0. The average of the four measurements was regarded as the vegetation height in the circle. The light conditions were measured as the overstory density using a convex densiometer after the guidelines by Lemmon (1957). Four measurements were carried out each at a distance of two meters from the centre of the 5 m circle. By assuming four squares within each of the 24 squared mirrors, each measurement is then based on 96 readings. With four measurements in each 5 m circle, the light conditions in each quadrat are therefore based on 384 readings.

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Figure 4: a) Plate for determination of the vegetation height corresponding to the height above ground level where 50% of the line is covered when viewed horizontally at a distance of an arm’s length. Examples of the lines location (vegetation height) in different vegetation types: b) grass- and herb vegetation, c) dwarf shrub vegetation, d) trees and shrubs raised above the ground vegetation, e) woodland without ground vegetation (Fredshavn et al., 2014a).

At each site, three of the five 5 m circles were further expanded to a 15 m circle (circular quadrat with a radius of 15 meters), each covering an area of almost 707 m2. If the site had a size or shape that made it impossible to lay out three 15 m circles or to ensure a minimum distance of 30 meters between the centres of two 15 m circles, the 15 m circles were given another shape to avoid any overlap. If another shape was applied, it was ensured that the area of 707 m2 was retained. Within the 15 m circles supplementary lists of woody species were made. As woody species shrubs, liana and trees were included, whereas Rubus ssp. and dwarf shrubs were not. The amount of deadwood and the number of trees with hollowness, decaying parts, occurrence of larger areas with bryophytes or lichens (<2 m above ground level), or woodpecker holes were counted as well within each of the 15 m circles.

For deadwood, only parts with a minimum length of two meters and a minimum diameter of 20 cm at breast height (dbh), equal to approximately 130 cm above the ground, were registered. Dead lateral branches with the same minimum dimensions were included. In addition to the dimensions, whether the deadwood was standing or lying (regarded as standing when the angel towards ground level was >45°) as well as the degree of decaying was noted. The dimensions were measured as truncated cones. The size of deadwood that could not be measured from the ground was estimated. For both lying and standing deadwood the degree of decay was estimated on a five-step scale:

1. Recently dead, typically within the last year. 2. The wood still hard (bark starting to fall off but typically still >50% bark). 3. The wood still hard but starting to become soft at the surface (often <50 % bark). 4. The wood soft at the surface and possibly all way through. The original wood structure starts to disappear. 5. The wood totally soft, very decayed and the original structure gone.

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Hollowness and larger decaying parts were only included if it occurred at the trunk or at lateral branches with a diameter >20 cm and >0.5 m above ground level. Hollowness were defined as a gap in the bark with underlying decay or hollowness to a depth of at least 5 cm. Larger decaying parts were areas of >100 cm2 with loosened bark or exposed wood with a clear decomposition process. Dead lateral branches >11 cm in diameter will leave decaying parts at the trunk (Fredshavn et al., 2014a), hence they were included. For alluvial forests with Alnus glutinosa and Fraxinus excelsior (Code 91E0 in the Habitat Directive: Council Directive 92/43/EC) the number of coppice stools larger than 70 cm in diameters were counted. Besides this, all individuals of woody species within the 15 m circles were registered according to their size with four size-classes: 1 = current year’s seedlings, 2 = saplings <0.5 m, 3 = saplings 0.5 – 2 m and 4 = trees >2 m (Buttenschøn & Buttenschøn, 1985; Buttenschøn & Buttenschøn, 2013). For all trees above 2 meters, the diameter at breast height was measured as well. Concurrent to the size the impact of browsing was estimated on a five-step scale for each individual: 0 referring to no browsing being observed, 1 to light browsing of few leaves and buds, 2 to medium browsing of leaves, buds and twigs, 3 to heavy browsing of leaves, buds and twigs, and 4 to nearly all leaves, buds and twigs within reach and some smaller or even larger branches being eaten or disrupted. In case of a very high density of a size-group of one species, the number of individuals and concurrent browsing impact was estimated from 60 quadrats of 1 m2 within the 15 m circles, hereby covering nearly 8.5% of the total area of a 15 m circle. But as the abundance of individuals and the browsing impact may be unequal at the circumference and the centre – as well as the different sides, the 60 quadrats were distributes evenly and in the same way for all 15 m circles (Figure 5).

Figure 5: In case of a very high density of a woody species the number of individuals and concurrent browsing impact were estimated from 60 quadrats of 1 m2 within the 15 m circle. The quadrats were distributes as shown here.

4.3. Data analysis To analyse the effect of the grazing management on the ground vegetation unpaired t-test were applied in all cases where a grazed and an ungrazed site could be compared pairwise. Prior to this, all compared data sets were tested for homoscedasticity (F-test) and in case the variances were found significantly unequal an appropriate transformation was chosen. In case homogeneity of variances could not be obtained by any

18 data transformation an unpaired t-test with Welch’s correction was carried out instead. In some situations it was more meaningful to compare the results between sites of the same habitat type (alder swamps, oak stands and beech stands), thus a one-way analysis of variance test (ANOVA-test) was applied. Also in these cases data were tested for homogeneity of variance (Levene’s-test), transformed if necessary or else the Welch variance-weighted ANOVA-test was used. When the ANOVA-test found a significant difference, the test was followed by a Tukey’s test, but only in cases where homoscedasticity was found or obtained by data transformation; otherwise no post hoc tests were performed.

In cases of few observations (N ≤ 5) it did not make any sense to test for a normal distribution. But as data were sampled from natural populations it was assumed that data were distributed more or less according to a Gaussian distribution. However, ANOVA-tests are not very sensitive to a moderate deviation from normality (McDonald, 2009). When data for the forest structure was compared pairwise, the numbers of observations were even less – only three observations from each site. Thus, in case of heteroscedasticity data was not attempted transformed, but Welch’s correction was applied directly.

The ecological factors were found significantly unequal between the sites and especially major differences in the overstory density appeared (Figure 12). Thus, it would obviously be most ideal to set up a test which would be able to disentangle the effect and importance of the different factors; grazing/non-grazing, location, overstory density, soil moisture, soil pH, plant available nitrogen and perhaps even woodland type, as well as the interactions between these factors, on the species richness and the species density. A factorial design (two-way ANOVA) with grazing/non-grazing and location as classification variables and the abiotic factors included as quantitative variables in the model, would be ideal. However, then sites without a pair to compare would have to be left out. Another major problem is the measure of the overstory density, which is an indirect measurement for the light availability, or more precisely a measurement for the amount of PAR (Photosynthetic Active Radiation). The issue is that the amount of PAR that reaches the forest floor not necessarily is equal in two different woodland types, even though the overstory density is the same. The amount of PAR penetrating the canopy is a matter of the LAI (Leaf Area Index) of the canopy. But as the LAI, and thus the amount of PAR that reaches the forest floor, depends on the tree species, the age of the trees and the soil concentration of plant available nutrients (Petersen & Vestergaard, 2006; Friis Møller, 2010), the overstory density will only be an appropriate indirect measurement for the amount of PAR in monoculture stands of the same age and with same soil fertility. Instead the relationship between the species composition of the ground vegetation and the environmental variables, grazing management and locations were explored by ordination techniques. However, the ordination method is not applicable when the effects of the variables are tried disentangled for a single species with four different size categories (the woody species). Thus, in these cases the factorial design (two-way ANOVA) with the overstory density as quantitative variable was applied anyway, but with the different habitat types (alder swamp, oak stand and beech stand) separately.

For data analysis Graphpad Prism (Ver. 6.0) and SAS Enterprise (Ver. 6.1) were used. All figures show data prior to any data transformation.

For analysis of the effects of grazing, ordination plots for each habitat type, based on the relative abundance of each plant species within the 5 m circles, were made. By plotting vegetation samples in relation to each other in terms of their similarity of species composition (as points in a coordinate system), a one-, two-, three-, or more dimensional ordination diagram depicting the dissimilarity between the

19 quadrats (with Sørensen’s dissimilarity index as distance measurement) in relation to the species composition and abundance can be generated. The method is primarily descriptive but can be used for data exploration and illustration of important patterns in the data. Hence, it gives insight to the structure of plant community and elucidates possible causal relationships between vegetation components and environment variables. A number of different ordination methods have been developed, each method with its own set of pros and cons and assumptions about the nature of the data. For example, the widely used PCA method (Principal Component Analysis) assumes a linear response between the species and environmental variables. In contrast the NMS method (Non-metric Multidimensional Scaling) is based on a unimodal species response curve (Kent, 2012). Regardless of which technique being used, the process of interpretation of ordination diagrams is similar. Points that are close together will represent quadrats which are similar in species composition; the further apart any two points are, the more dissimilar the quadrats will be.

Although the plant species in this study not were tested for linear or unimodal response curves, the NMS method was applied, as the prevailing and most likely relation between a species response and an environmental factor is unimodal (Kent, 2012). In NMS, a dissimilarity matrix is calculated between all pairs of quadrats. Then points representing the quadrats are positioned within a number of dimensions or ordination axes, so the distances between the points have the best possible correlation with the dissimilarity matrix. There is no exact way to do this, thus the points are at first placed randomly, then “pushed” a bit, and if the new position increases the correlation they are pushed further in the same direction. When the points stabilize, the coordinates of each point can be calculated as well as a coefficient of determination – indicating how well the inter-point dissimilarity is correlated with the matrix (Adsersen & Mølgaard, 2003). The measure for how good a fit or match occurs between the two is called the stress. Then, a Monte Carlo test was used; meaning that the original data were randomized and all the calculations were redone (in this case 250 times). The number of times the real data resulted in lower scores for stress than the randomized data constitutes a measure of the “goodness” of the result.

The ordinations were performed with the software PC-ORD (Ver. 6.0) and two-dimensional diagrams were chosen, simply because these plots are easiest to read and will show more information than a one- dimensional one.

5. Results In total, 16 sites were analysed for ground vegetation and forest structures. In every site five circles of 5 meters in diameters were established. Hereby a total of 80 circles and a total area of 6283 m2 were analysed for ground vegetation. Furthermore, a total of 48 circles of 15 meters in diameter were analysed as well, given a total area of almost 3.4 ha that were analysed for woody species and browsing impact, the amount of deadwood and the number of trees with hollowness, decaying parts, occurrence of larger areas with bryophytes and/or lichens and the number of woodpecker holes.

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5.1. Ground vegetation With certainty, a total of at least 160 species of vascular plants were identified within the 5 m circles (Appendix 1). However, the actual number of species is presumably somewhat higher as not all individuals with certainty could be identified to species level. For example individuals of Viola reichenbachiana or V. riviniana were found in several plots. The two species cannot be distinguished without the presence of either flowers or the seed capsules with attached sepals, but as none of the individuals found had neither petals nor sepals present, it is possible that both species actually are present. In addition, individuals of Taraxacum were not attempted identified to species level. This is due to the degree of difficulty as the season for this genus is very limited, often restricted to only a few weeks in the early summer and because of the huge species variation in relation to soil conditions and shade (Mossberg & Stenberg, 2003). This also applies to brambles which only were identified to the two sections; Rubus sect. Rubus and R. sect. Corylifolii. Hence, when species richness is mentioned in the following, it is a conservative estimate and the minimum number of species present.

5.1.1. Species richness and density With a total of 71 species the highest species richness of vascular plants is registered in the grazed alder swamp in Lille Hessemose, while the highest percentage difference in species richness is found in the old oak stand at Sandskredssøen, where the grazed area contained 114 % mores plant species compared to the ungrazed (Figure 6). In general, the species richness appears strongly increased by the grazing management in the alder swamp and the oak stands, except in the young oak stand at Sandskredssøen where only a minor increase is found. By contrast, the species richness seems to decrease by grazing in the beech stands except in Store Hessemose where four more species are found in the grazed site (Figure 6).

80 Grazed Ungrazed

s 60 s e n h c i r 40 s e i c e p

S 20

0 ) ) E H E y o H E S H K S K ( ( L K h S r r k S S k h c h e e a k k a c e c ld ld a a e e e A A O O e B e O O B B Figure 6: The total number of vascular plant species found within the five 5 m circles in each site.

To prevent pseudoreplication when the average species densities per 0.1 m2 are calculated, the mean density for each of 5 m circles were first calculated based on the species richness in the 25 Raunkjær circles. Next the average densities for the five 5 m circles were found.

21

The average species richness in the 5 m circles is found significantly higher in the pairwise comparable grazed sites in both the alder swamp in Stores Hessemose (P < 0.01), the old oak stand at Sandskredssøen (P < 0.01) and the oak stand in Lille Hessemose (P < 0.05) (Figure 7). However, the only significant difference in the species density (species richness per 0.1 m2) for these three sites is found in the oak stands in Lille Hessemose with a higher species density in the ungrazed site (P < 0.05) (Figure 8). But interestingly, the opposite is found in the old oak stands at Sandskredssøen, where the species density appears to be increased by grazing, though, not significantly. The only other significant difference in the species density is found in beech stands in Kollerup Enghave, where grazing is decreasing the species density (P < 0.05) (Figure 8).

) 50 Grazed s e

l Ungrazed c r i

c 40 ** * m

** 5 (

30 s s e

n NS

h 20 c i

r NS NS NS s e i 10 c e p

S 0 ) ) E H E y o H E S H K S K ( ( L K h S r r k S S k h c h e e a k k a c e c ld ld a a e e e A A O O e B e O O B B Figure 7: The average species richness (±SE) of vascular plant species within the 5 m circles. †P < 0.1; *P < 0.05; **P < 0.01; ***P < 0.001.

8 Grazed ) 2 Ungrazed m

NS 1 , * 0

6

r NS e p (

y t i 4 s NS n e d

s NS e 2 * i c e

p NS S 0 ) ) E H E y o H E S H K S K ( ( L K h S r r k S S k h c h e e a k k a c e c ld ld a a e e e A A O O e B e O O B B Figure 8: The average density of vascular plant species (±SE) per 0,1 m2. †P < 0.1; *P < 0.05; **P < 0.01; ***P < 0.001.

A condition for the direct comparison of the grazed and ungrazed sites is that all other factors are kept equal. Hence, to assess the factors influences on plant performance, Ellenberg’s indicator values for soil

22 moisture, soil pH and plant available nitrogen were used (Ellenberg et al., 1992). To give a better estimate the mean values were calculated by weighting the indicator values by the relative abundance of each species in the vegetation. However, as all species found in the beech stands in Store Hessemose were represented with very low abundances (even in one of the quadrats in both the grazed and ungrazed site no individuals were registered in any of the 25 Raunkjær circles) no Ellenberg-values are calculated for these sites. Each woodland type (alder, oak, beech) is analysed separately. A one-way ANOVA test finds the soil moisture to be significantly equal in all the alder swamps (with Welch’s correction) and all beech stands, while a significant difference occurs between the oak stands (P < 0.001). However, Tukey’s test does not find any of the pairwise comparable sites to be significantly different (Figure 9). This is also the case for both the soil pH (Figure 10) and the plant available nitrogen (Figure 11). However, it is important to notice the pronounced and significantly lower soil pH and plant available nitrogen content in the young oak stands at Sandskredssøen compared to the other oak stands. It is clear that these two oak stands are on much more acidic soils than the others.

Grazed 8 A Ungrazed A A A A A BC AB AB F 6 - C A A A

g A r e b

n 4 e l l E 2

0 Alder KE Alder SH Oak KE Oak S(y) Oak S(o) Oak LH Beech KE Beech S

Figure 9: Average weighted Ellenberg-values (±SE) for soil moisture.

Grazed 8 Ungrazed A AB

R 6 - B A g r A e A A A b

n 4 A e l l A E B B A A 2

0 Alder KE Alder SH Oak KE Oak S(y) Oak S(o) Oak LH Beech KE Beech S

Figure 10: Average weighted Ellenberg-values (±SE) for soil pH.

23

Grazed 8 Ungrazed

A A A A A N 6 - B g

r B B

e A A b A A n 4 C e

l C l E 2

0 Alder KE Alder SH Oak KE Oak S(y) Oak S(o) Oak LH Beech KE Beech S

Figure 11: Average weighted Ellenberg-values (±SE) for plant available nitrogen. 100 Grazed AB A A Ungrazed AB B B ) %

( A

A y t i 80 s n

e A d

y r o t s

r 60 e v O

40 ) ) E H E y o H E S H K S K ( ( L K h S r r k S S k h c h e e a k k a c e c ld ld a a e e e A A O O e B e O O B B Figure 12: Average overstory densities (±SE). A multiple comparison test is not performed for the oak stands due to the lack of homoscedasticity and the impossibility of obtain it by any data transformation.

No significantly differences for the overstory density are found in the alder swamps (one-way ANOVA-test) (Figure 12). In the beech stands, however, a significant difference is found (P < 0.01) (after the data has been square root transformed to obtain equal variances (Levene’s test)), but Tukey’s test reveals that no significant differences occurs between the pairwise comparable sites. A clear significant difference is found in the oak stands as well (Welch’s variance-weighted ANOVA: P < 0.001). Although a multiple comparison test cannot be performed due to the lack of homoscedasticity and the impossibility of obtaining it by data transformation, it appears that the pairwise comparable sites, at least those at Sandskredssøen, have clear different average overstory densities (Figure 12).

5.1.2. Implications of abiotic factors For further analysis of the effects of grazing in relation to species richness, consequences of variations in abiotic factors and effects on species level, ordination plots are made based on the relative abundance of each plant species within the 5 m circles. However, as the oak stands are clearly different in soil pH and plant available nitrogen, the plant communities will be different too (Watkinson et al., 2001). Moreover, the grazing management may potentially have different effects due to the different vegetation communities and the abiotic factors. Hence, the two stands on acidic soil are analysed separately, as well as the other woodland types. Ordination plots for all the analysed woodland types (alder swamp, oak on non-

24 acidic soil, oak on acidic soil and beech on acidic soil) are made and in all cases a Monte Carlo test (with 250 permutations) finds them to be significant (in all cases P = 0.004).

For the ordination plot with the alder swamps the coefficient of determination is calculated to be 0.868 – corresponding to 86.8 % of the variations in the dissimilarity matrix can be explained by the positions of the points in the ordination plot. It appears that the vegetation communities in the three sites are different in composition, and that the spatial vegetation heterogeneity is higher in the alder swamp in Kollerup Enghave than those in Store Hessemose. Even though the overstory density is higher in the ungrazed site, where the lowest species richness is found (Figure 7), it is not significantly correlated with the species richness (Table 1). By contrast, species richness is significantly positively correlated with the grazing history (P < 0.001) and significantly negatively correlated with the vegetation height (P < 0.05). Hence, grazing management appears to be the best single determinant for the species richness in the alder swamps. When it comes to the vegetation some species are clearly represented with a higher abundance in some sites. This is clearly the case for both Stellaria holostea and Milium effusum which have higher abundances in the ungrazed site (Figure 13). But whether this is due to the absence of large herbivores, the higher overstory density or a combination, or even another unknown factor, is not clear. Some species, especially Ranunculus repens and Lysimachia vulgaris are clearly, positively correlated with the soil moisture whereas some species (Poa trivialis, Deschampsia cespitosa and Glechoma hederacea) are negatively correlated with increasing pH (Figure 13). However, it is important to remember that the values for soil moisture, pH and plant available nitrogen are calculated from Ellenberg-values. The correlations between the edaphic conditions and certain plant species may thus be a natural result of this fact rather than a representation of the real world’s variations.

Both the ordination plots for the oak stands on non-acidic soil (coefficient of determination: 0.752) (Figure 14) and the ordination plot for oak stands on acidic soil (coefficient of determination: 0.926) (Figure 15) depict the individual sites separately that indicates that the grazing management has changed the species composition in both Lille Hessemose and Sandskredssøen. For both plots a clearly higher vegetation height is found in the ungrazed sites. For the stands on non-acidic soil this is strongly negatively correlated with the species richness (P < 0.001) (Table 1), which is found higher in the grazed sites. Interestingly, the species richness and the overstory density is positively correlated (Table 1), thus, the overstory density is probably not the determining inhibitor for the species richness in these systems. It is clear that some species are found at a much higher abundance in some sites. This is especially true for the grazed site at Sandsskredssøen where arrows for Moehringia trinervia, Galeopsis bifida, Brachypodium sylvaticum, Epilobium montanum, Deschampsia despoitosa, Galium aparine, Urtica dioica, Carex remota, Circaea lutetiana and Stellaria media and/or S. neglecta are projecting at (Figure 14). In the oak stands on acidic soil no species appear to increase abundances in the grazed sites, while Vaccinium myrtillus and Quercus robur have higher frequencies in the ungrazed site (Figure 15).

The ordination plot for the beech stands (coefficient of determination: 0.948) elucidate that the vegetation composition in neither Store Hessemose nor Sandsskredssøen has changed with the introduction of the grazing management (Figure 16). Furthermore, the grazing has decreased the spatial vegetation heterogeneity in both Sandskredssøen and Kollerup Enghave. Both the species richness and the species density are significantly positively correlated with the vegetation height, and significantly negatively

25 correlated with the overstory density (Table 1). All species represented by arrows have either no correlation or are negatively correlated with the overstory density (Figure 16).

Figure 13: NMS-ordination of the vegetation composition in the alder swamps

Figure 14: NMS-ordination of the vegetation composition in the oak stands on non-acidic soil

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Figure 15: NMS-ordination of the vegetation composition in the oak stand on acidic soil.

Figure 16: NMS-ordination of the vegetation composition in the beech stands.

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Overstory density Vegetation height Ellenberg-F Ellenberg-R Ellenberg-N Years with grazing Alder swamps Species richness -0,414 -0,628* 0,187 -0,306 0,484 0,820*** Species density -0,407 -0,421 0,102 -0,062 0,445 0,448 Overstory density 0,243 -0,279 0,408 0,001 -0,492 Oak stands (on non-acidic soil) Species richness 0,603** -0,732*** -0,472* 0,484* -0,214 0,693*** Species density 0,325 -0,276 -0,196 0,534** -0,033 0,050 Overstory density -0,614** -0,455* 0,225 -0,256 0,626*** Oak stands (on acidic soil) Species richness -0,553 -0,500 -0,218 0,468 0,251 Species density -0,029 0,252 0,268 0,283 -0,646* Overstory density 0,767** 0,518 0,048 -0,274 Beech stands Species richness -0,454* 0,642*** 0,618** 0,683*** 0,621** -0,062 Species density -0,775*** 0,785*** 0,456* 0,448* 0,430 -0,268 Overstory density -0,588*** -0,239 -0,062 -0,296 0,140 Table 1: Pearson’s correlation coefficients. The correlation coefficients for beech stands with Ellenberg-values are only based on data from Sandskredssøen and Kollerup Enghave.

5.1.3. Spatial heterogeneity The ordination plots have already indicated differences in the spatial vegetation heterogeneity, but to test any differences statistically the dissimilarities in the vegetation compositions between each of the five random samples from each site are calculated (the spatial species turnover or the β-diversity). With five samples, ten comparisons are made and the average of the ten values gives an expression of the spatial vegetation heterogeneity. All the pairwise comparable grazed oak stands exhibit a clear increased spatial heterogeneity in the vegetation community compared to the ungrazed, though, only significance is found in the old oak stand at Sandskredssøen (P < 0.05) and the oak stand in Lille Hessemose (P < 0.01) (Figure 17). In the alder swamps in Store Hessemose there are not found any differences. In the oak stand there is clearly a pattern with an increased spatial heterogeneity in all the grazed sites, though only significant in the old oak stand at Sandskredssøen (P < 0.05) and the oak stand in Lille Hessemose (P < 0.01). In the beech stands is the pattern opposite and more homogeneous vegetation develops by the grazing (Figure 17).

Significantly decreased vegetation heights in the grazed areas are found in both the alder swamps in Store Hessemose (P < 0.01), the young and old oak stands at Sandskredssøen (respectively; P < 0.05 and P < 0.001) and the oak stands in Lille Hessemose (P < 0.001) (Figure 18). In the beech stands where the vegetation is much more sparsely, the vegetation height in the ungrazed sites are much lower compared to the ungrazed sites in the other woodland types. No significant differences are found between any of the pairwise comparable sites (Figure 18).

28

Figure 17: Average spatial heterogeneity (±SE) in the vegetation, measured as the β-diversity (Sørensen’s dissimilarity). †P < 0.1; *P < 0.05; **P < 0.01; ***P < 0.001.

70 ** *** Grazed Ungrazed

) 50 *** m

c *

( 30

t h

g 10 i e

h 9

n 7 o i

t 5

a NS t 5 e

g 4 NS e 3 V 2 1 0 ) ) E H E y o H E S H K S K ( ( L K h S r r k S S k h c h e e a k k a c e c ld ld a a e e e A A O O e B e O O B B Figure 18: Average vegetation height (±SE). †P < 0.1; *P < 0.05; **P < 0.01; ***P < 0.001.

5.1.4. Diversity The term diversity is often used interchangeably with the species richness in a community. But diversity is basically a matter of predictability. The more diversity in a community, the less certainty there is about the species of a random individual from the community. Hence, even if no changes in the species richness are registered an alteration in the abundances among the species present may alter the diversity. As indicated by the ordination plots (Figure 13, Figure 14, Figure 15 and Figure 16), the presence of large herbivores has altered the vegetation composition, hence, the diversity may be modified too. Many different diversity indexes have been developed through time, all with different weight of the two factors; species richness and abundance. Here Simpsons Diversity Index (expressed as 1-D) is applied as it has the advantage that the values lies between 0 and 1, indicating the probability of two random individuals from the same community belongs to different species. However, for a given number of species the maximum diversity index will be reached if all species has equal abundances. Hence, it may be interesting to examine the evenness alone. Of these reasons, Simpsons Diversity Index and the diverted Simpsons Evenness Index are chosen.

29

Not only the species richness in the grazed alder swamp in Store Hessemose has increased (Figure 7), but the abundance among the species found does also have a clear tendency (P < 0.1) to be more even (Figure 20), thus, not surprisingly the diversity has increased significantly too (P < 0.05) (Figure 19). The same is true for the old oak stand at Sandskredssøen but for the young oak stand the picture is different. Although no significant differences in neither the species richness (Figure 7) nor the evenness are found (Figure 20), a significantly decreased diversity (P < 0.05) is present at the grazed site (Figure 19). For the beech stand in Kollerup Enghave the evenness in the grazed site has become significantly decreased (P < 0.05) (Figure 20) and the diversity as well (P < 0.05) (Figure 19).

1.0 Grazed

x * *** NS Ungrazed e d n

I * 0.8 y t i s r )

e NS

D * v i - 0.6 1 D

( s ' n o

s 0.4 NS p m i S 0.2 ) ) E H E y o H E S H K S K ( ( L K h S r r k S S k h c h e e a k k a c e c ld ld a a e e e A A O O e B e O O B B Figure 19: Average values (±SE) of Simpsons Diversity Index (1-D). †P < 0.1; *P < 0.05; **P < 0.01; ***P < 0.001.

1.0 † NS Grazed

x ** Ungrazed e NS

d * n

I NS 0.8 s s e * NS n n e

v 0.6 E

s '

n NS o

s 0.4 p m i S 0.2 ) ) E H E y o H E S H K S K ( ( L K h S r r k S S k h c h e e a k k a c e c ld ld a a e e e A A O O e B e O O B B Figure 20: Average values (±SE) of Simpson’s Evenness Index. †P < 0.1; *P < 0.05; **P < 0.01; ***P < 0.001.

5.1.5. Species composition and abundance From the ordination plots it is easily seen that the vegetation compositions in both the alder swamps and the oak stands including all the pairwise comparable sites are different from each other (Figure 13, Figure 14 and Figure 15). The differences are partly due to the colonization of new species as well as the disappearance of other in the grazed sites – but of course also due to alterations in present species abundances because of changes in the intraspecific competition between the species. Many of the species

30 found are represented with such a low frequency that comparisons are impossible, but with several other species it makes sense. However, only species where a significantly difference or a pronounced, although not significant, difference in their abundances are included here. Hence, the beech stands are not further analysed as the relative abundance for species of the ground vegetation present in these sites are either too low for any meaningful comparison or there are not found any tendencies for differences in their relative abundance, which is the case for e.g. Deschampsia flexuosa and Carex pilulifera. Although, the differences found for the species abundances may be caused by other factors than the effects of grazing alone, it is assumed that the pairwise comparable ungrazed sites act as a picture of the pre-grazing management period – a condition to be able to analyse and evaluate the effect of grazing. Of course it would be most correct to follow the development in the vegetation over time, with the fluctuations that may occur to be able to fully conclude any alterations. However, with only one field period for this study, and without any previous inventory of the plant species composition, this is clearly not an option. With the numerous pitfalls for drawing incorrect conclusions, the different grazed sites are instead attempted compared. Thus, timelines with changes in abundance after two to four years of management are tried established.

Deschampsia cespitosa Milium effusum Stellaria holostea Carex acutiformis Poa trivialis 25 *** Grazed *** * Ungrazed

e 20 c n

a NS d *** n 15 u b a

e

v 10 i t a l e

R 5

0 Urtica dioica Calamagrostis canescens Glechoma hedcracea Ranunculus repens Phalaris arundinacea 15 NS e c n

a 10 d n

u NS b a

NS e v i t 5 * a l * e R

0 ) ) ) ) ) ) ) ) ) ) rs rs rs rs rs rs rs rs rs rs a a a a a a a a a a e e e e e e e e e e y y y y y y y y y y 2 4 2 4 2 4 2 4 2 ( ( ( ( ( ( ( ( ( (4 e e e e e e e v s v s v se v s e e a a a a o v s h o h o h o h a o g m g m g m g m h m n e n e n e n e g e E s E s s E s n s s s E s s E s p e p e p e p e p e ru H ru H ru H ru H u e e e r H ll re ll re lle re ll re le e o o o o o o o to l r K t K t K t K o to S S S S K S

Figure 21: The average relative abundance (±SE) of vascular plant species found in the alder swamps based on 25 quadrats of 0.1 m2 in five circular quadrats (radius of 5 meters). Only species which displayed a significant or a clear, though not significant difference between the grazed and ungrazed sites are included. †P < 0.1; *P < 0.05; **P < 0.01; ***P < 0.001.

In the alder swamps in Store Hessemose the relative abundance of both Milium effusum and Stellaria holostea is significantly higher in the ungrazed compared to the grazed site (P < 0.001), but also a tendency, though not significant is found for Calamagrostis canescens and Carex acutiformis. In addition, there are tendencies for declines in the relative abundance over time for Milium effusum, Stellaria holostea and

31

Calamagrostis canescens when compared to the two years grazed alder swamp in Kolleup Enghave (Figure 21). In contrast, significantly increases in relative abundance in the grazed site are found for Poa trivialis (P < 0.05), Deschampsia cespitosa (P < 0.001), Ranunculus repens (P < 0.05) and Phalaris arundinacea (P < 0.05) and the same tendency is observed for Urtica dioica and Glechoma hedcracea, though not significant (Figure 21). For both grass species; Poa trivialis and Deschampsia cespitosa there are further a clear increase in the relative abundance with increasing time of grazing management (Figure 21).

Urtica dioica Moehringia trinervia 6 Grazed NS Ungrazed e c n

a 4 d

n † u

b NS a

e v i t 2 a l e

R NS

0 Deschampsia cespitosa 25 Calamagrostis epigejos Calamagrostis canescens * NS

e 20 c n

a NS *** d

n 15 † u b

a NS

e

v 10 i t a l e

R 5

0 Glechoma hederacea 15 Viola reichenbachiana/ V. riviniana Stellaria media/ S. neglecta NS e c

n † a 10 d n

u NS b a

* e v i t 5 NS a l e

R NS

0 ) ) ) ) ) ) ) ) ) rs rs rs rs rs rs rs rs rs a a a a a a a a a e e e e e e e e e y y y y y y y y y 2 3 4 2 3 4 2 3 4 ( ( ( ( ( ( ( ( ( e ) e e ) e e ) e v ld s v ld s v ld s a o o a o o a o o h ( m h ( m h ( m g n e g n e g n n e s n e s n e se E ø s E ø s E ø s p s e p s e p s e ru s H ru s H ru s H d e e d e d e lle re ll ll re ll lle re ll k i o k i k i o s L s L o s L K d K d K d n n n a a a S S S

Figure 22: The average relative abundance (±SE) of vascular plant species found in the not acidic oak stands based on 25 quadrats of 0,1 m2 in five circular quadrats (radius of 5 meters). Only species which displayed a significant or clear, though not significant difference between the grazed and ungrazed sites are included. †P < 0.1; *P < 0.05; **P < 0.01; ***P < 0.001.

The vegetation alteration in the oak stands are analysed in relation to habitat types, hence, the changes in the relative abundance for plant species found in the acidophilous stands (the young oak stands at Sandskredssøen) are not directly compared with the species found in the oak stands on mull soil, as the vegetation compositions are not alike. With data of the plant community of two times grazed and ungrazed pairwise comparable sites with oak stands on non-acidic soil, it has the advantage that an increase or decrease in the relative abundance for a given plant species in both comparable areas, with a higher certainty can be attributed to the grazing management and not caused by other factors. However, none of

32 the plant species displayed a significantly increase in their relative abundances in both of the comparable grazed sites. But for Viola reichenbachiana and/or V. riviniana, Utrica dioica, Moehringia trinervia, Deschampsia cespitosa, Glechoma hederacea and Stellaria media and/or S. neglecta a pattern with either a significant increase in one of the grazed sites, or at least a tendency, is observed (Figure 22). An interesting pattern is especially found for Glechoma hederacea, with the absence in the ungrazed sites and an enhanced relative abundance with increasing time of grazing management. Both Calamagrostis epigejos and C. canescens have lower relative abundances in the grazed sites, especially for C. epigejos which displays a significant difference in the site with three years grazing management and a significant tendency (P < 0.1) in the site with four years of management (Figure 22).

In the oak stands on acidic soil a significant difference in the relative abundance is only found for Vaccinium myrtillus with a negative impact of grazing (Figure 23). But a tendency towards a decrease in abundance in the grazed site is also observed for Molina caerulea and Trientalis europaea. Opposite, a significant tendency (P < 0.1) towards an increased abundance by grazing is found for Betula sp. and Calamagrostis epigejos (Figure 23). However, C. epigejos was not registered in the ungrazed site at all.

20 Grazed Ungrazed e

c 15 NS n a d n

u *** b NS

a 10

† e v i t a l

e 5 † R

0 . s a s a p u e o e s ll l j a a ti ru e l r e ig p tu y a p ro e m c e u B a s e m i ti is iu lin s l in o ta c o r n c M g e a a ri V m T la a C Figure 23: The average abundance (±SE) of vascular plant species found in the oak stands on acidic soil (Sandskredssøen (young)) based on 25 quadrats of 0,1 m2 in five circular quadrats (radius of 5 meters). Only species which displayed a significant or clear, though not significant difference between the grazed and ungrazed sites are included. †P < 0.1; *P < 0.05; **P < 0.01; ***P < 0.001.

As the different sample sites are spread over a wide area, they will naturally differ in abiotic factors e.g. due to variations in soil properties. These variations in abiotic factors may influence the species composition of the vegetation. Hence, a directly comparison of the abundances of specific species between the areas may be biased. Instead, a plant community can be viewed as a collection, not of species, but of groups or functional types. Here the so-called C-S-R system (Grime, 1979) is particularly efficient in the balance between the potential for interpreting and predicting vegetation and ecosystem properties (Hodgson et al., 1999) and the simplicity of its assumptions (Hunt et al., 2004).

33

The system is based on two groups of external factors, both affecting the performance of plants. The first group, stress, consists of factors placing prior restrictions on plant production, such as shortages of light, water and nutrients. In the second group, disturbances, consists of factors causing partial or total destruction of plant production after it has been formed. This group includes intensifications factors such as grazing, trampling, mowing and ploughing, but also extreme climate events such as wind-damage, frosting, drought, soil erosion and fire (Hodgson et al., 1999). Plant species associated with three of the four possible permutations of environmental extremes displays distinct sets of ecological traits: competitiveness in the case of low stress and low disturbance, stress-tolerant, in the case of high stress and low disturbance and ruderality in the case of low stress and high disturbance. High stress and high disturbance does not support life at all.

Each plant species are classified as a functional type, according to the emphasis between the three ecological traits (C-S-R). Based on the species present and their relative abundances a net position of the whole community can then be calculated to each of the C-S-R dimensions. This is done based on the relative abundance of each of the plant species found in each 5 m circle and the net position of the vegetation in the C-S-R space is found using the C-S-R signature calculator from UCPE Scheffield (ver. 1.2). For the C-dimension, the only significant difference, with a decrease by grazing (P < 0.05) is observed in the old oak stand at Sandskredssøen (Figure 24). However, a tendency towards the opposite is seen in the alder swamp in Store Hessemose where grazing slightly increases the C-dimension (P = 0.1). The alder swamp in Store Hessemose is also the only site where a significant difference in the S-dimension is found, with a strongly decrease in the grazed vegetation (P < 0.001) (Figure 25). The most characteristic changes in the vegetation composition, however, appear when it comes to rudeality (R-dimension). Here grazing has significantly altered the vegetation compositions to become more disturbance tolerant in both the alder swamp in Store Hessemose (P < 0.01), the old oak stand at Sandskredssøen (P < 0.01) and the oak stand in Lille Hessemose (P < 0.01) (Figure 26). In neither the young oak stand at Sandskredssøen nor any of the beech stands any significantly changes are found. The beech stands in Store Hessemose were not included as the numbers of individuals of plants simply were too few to make any meaningful calculations upon.

0.8 Grazed Ungrazed 0.6 n o i

s † * n e 0.4 NS m i NS d NS NS -

C 0.2

0.0 ) ) E H E y o H E S K S K ( ( L K h r r k S S k h c e e a k k a c e ld ld a a e e A O O e B A O O B Figure 24: The average net position (±SE) of the vegetation communities in the C-dimension of the C-S-R space. †P < 0.1; *P < 0.05; **P < 0.01; ***P < 0.001.

34

0.8 Grazed Ungrazed NS NS NS 0.6 n o i NS NS s n e 0.4 *** m i d

-

S 0.2

0.0 ) ) E H E y o H E S K S K ( ( L K h r r k S S k h c e e a k k a c e ld ld a a e e A O O e B A O O B Figure 25: The average net position (±SE) of the vegetation communities in the S-dimension of the C-S-R space. †P < 0.1; *P < 0.05; **P < 0.01; ***P < 0.001.

0.8 Grazed Ungrazed

0.6 n o i s n e 0.4 m i ** d

- **

R ** 0.2 NS NS NS 0.0 ) ) E H E y o H E S K S K ( ( L K h r r k S S k h c e e a k k a c e ld ld a a e e A O O e B A O O B Figure 26: The average net position (±SE) of the vegetation communities in the C-dimension of the C-S-R space. †P < 0.1; *P < 0.05; **P < 0.01; ***P < 0.001.

5.2. Forest structures Totally, at least 29 different woody species were registered within the 15 m circles (Appendix 2 and Appendix 3). However, many species are found in so few sites and/or represented with an abundance that prevents any meaningful comparisons. Hence, only species with an abundance which permits an appropriate comparison are analysed further at species level.

For some seedlings of woody species the registration were challenged by the absence of characteristic morphological features. This was for instance the case for species of Betula. Normally B. pendula is characterised by being hairless and the presence of resin warts and B. pubescens by being hairy and the lack of resin wards, but seedlings of B. pendula may be hairy and lack glandular resin warts on twigs (Poland & Clement, 2009). Hence seedlings of Betula were not identified to species level. Also individuals of Salix constituted a challenge with many hybrids (Poland & Clement, 2009) and the possibility of the lack of striae under the bark on seedlings and young saplings.

35

5.2.1. Species richness and regeneration of woody species With only few years of grazing management in all grazed sites, grazing is not assumed to have had any influence on the abundance or diversity of trees >2 m. The only expected way the horses can have an impact is by selective bark stripping, which were registered in all cases it occurred. However, bark stripping was very infrequently observed. Two individuals of Crataegus sp. were found stripped, both were saplings between 0.5 m and 2.0 m growing in grazed sites, one in the alder swamp in Kollerup Enghave and one in the oak stand in Lille Hessemose. Alnus incana was the species found most frequently striped for bark, in all cases it was individuals of saplings between 0.5 m and 2.0 m or trees >2 m with a dbh <10 cm. One individual was found in the alder swamp in Kollerup Enghave and 7 individuals in the grazed alder swamp in Store Hessemose, but also 13 individuals were found in the ungrazed alder swamp in Store Hessemose. Hence, the grazing regimes are apparently not affecting the frequency of bark stripping.

A general tendency for a higher species richness of woody species in the grazed sites appears, although the only significant difference is found in the oak stands in Lille Hessemose, with more species in the grazed site (P < 0.05) (Figure 27). Looking at the oak stands, a tendency towards an increased species richness of seedlings in the grazed sites appear, although only significant in the old oak stands at Sandskredssøen (P < 0.05) and in Lille Hessemose (P < 0.05). No differences are found for the saplings <0.5 m (Figure 29). For the saplings between 0.5 m and 2.0 m a tendency towards a higher species richness in the grazed site in Lille Hessemose is found (P < 0.1) (Figure 30) whereas the opposite is found in the old oak stand at Sandskredssøen. However, it may be important to have in mind that the species richness of seedlings and saplings may simply be a function of the overstory density and/or species richness of trees >2 m rather than the grazing regime itself. Thus, to test the significance of these factors an ANOVA-test with grazing as classification variable and overstory density and species richness of trees >2 m as quantitative variables is set up for the pairwise comparable oak stands on mull soil. But with only three observations in each site, and four sites, 11 degrees of freedom are available for the test. For this reason the test is optimised and only includes the factors themselves and the interactions where grazing are included, although it still means that only 6 degrees of freedom are left to test for significance. Another issue is the assumption of homogeneity of variance. With only three observations in each site, it is impossible to transform data to obtain homoscedasticity – particularly if all three observations in one site by change are equal. Hence, the results should be interpreted carefully. Neither any of the factors nor the interactions can significantly explain the variation in the species richness of seedlings nor of saplings of woody species in the oak stands on mull soil (Table 2). However, it is important to notice how the richness appears positively correlated with the richness of trees >2 m., particularly in Lille Hessemose (Figure 31).

For the beech stands no differences are found between any of the pairwise comparable sites for neither the richness of seedlings (Figure 28), saplings <0.5 m (Figure 29) nor the saplings between 0.5 m and 2.0 m (Figure 30). However, for both the grazed and ungrazed site in Store Hessemose a clear decrease in species richness with increasing size is observed. The sites have more or less the same richness of seedlings of woody species as the other beech stands but none saplings in the size group 0.5 – 2.0 m are found at all, although the highest richness of trees >2 m for the beech stands are found in the grazed stand in Store Hessemose (Appendix 3). However, the ANOVA test finds the overstory density to be able to significantly (P < 0.01) explain the variation in the richness of saplings <0.5 m and the variations in the richness of saplings between 0.5 m and 2.0 m (P < 0.05) (Table 2); hence, the lowered richness in Store Hessemose is simply a matter of out-shading.

36

15 Grazed * Ungrazed

f o NS s s e s i

e 10 c n e

h NS NS

p NS

c NS s i †

r

y s d e o i o c 5 e w p S

0 ) ) E H E y o H E S H K S K ( ( L K h S r r k S S k h c h e e a k k a c e c ld ld a a e e e A A O O e B e O O B B Figure 27: Average species richness (±SE) of woody species. †P < 0.1; *P < 0.05; **P < 0.01; ***P < 0.001.

10 Grazed 15 Grazed NS Ungrazed Ungrazed * NS 8 NS s s s s NS m

e e

5 10 n n , s NS h h 0 g 6 NS c c

* < n i i i

NS l r r NS

s d NS s s g

e NS e e n i i e i 4 l c c s p e e 5 NS a p p s NS S S 2

0 0 ) ) E H E ) ) E S E H E y o H E S H (y o H H K S K ( ( L K h S K S K ( L K h S r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld O a a O e e e A A O O e B e A A e B e O O B B O O B B Figure 28: Average species richness (±SE) of seedlings of woody Figure 29: Average species richness (±SE) of saplings <0.5 m of species. †P < 0.1; *P < 0.05; **P < 0.01; ***P < 0.001. woody species. †P < 0.1; *P < 0.05; **P < 0.01; ***P < 0.001.

5 Grazed 5 Grazed Ungrazed Ungrazed NS † 4 4 s m s

NS NS s s 0 e , e m n 2 NS † n

- h h

3 2 3 5 c c ,

i NS i >

r r 0

NS

s s s s e NS e e g i i 2 e

r 2 n

c NS c

i NS NS l T e e p p p a S S NS s 1 1

0 0 ) ) E H E y o H E S H E H E ) ) H E S H K S K ( ( L K S K (y (o L K r S h S K h S e r k S k h c h r r k S S k h c h d e a k k a c e c e e a k k a c e c l ld O a a O e e e ld ld a a e e e A A e B e A A O O e B e O O B B O O B B Figure 30: Average species richness (±SE) of saplings of woody Figure 31: Average species richness (±SE) of woody species species between 0.5 and 2.0 m. >2 m. †P < 0.1; *P < 0.05; **P < 0.01; ***P < 0.001. †P < 0.1; *P < 0.05; **P < 0.01; ***P < 0.001.

37

Oak woodland Beech woodland Richness of Richness of Richness of Richness of Richness of Richness of seedlings saplings <0.5 m saplings 0.5-2.0 seedlings saplings <0.5 m saplings 0.5-2.0 m m Grazing 0.320 0.115 0.099 0.290 0.721 0.080 Overstory density 0.458 0.052 0.096 0.468 0.006 0.034 Richness of trees 0.322 0.749 0.655 0.735 0.369 0.208 Grazing*Overstory 0.363 0.143 0.151 0.328 0.862 0.104 density Grazing*Richness 0.320 0.981 0.599 0.991 0.155 0.790 of trees Table 2: P-values for ANOVA-tests with grazing as classification variable and overstory density and richness of trees as quantitative variables. Significant values are shown in bold.

In addition to the richness of woody species it is of interest to analyse the effect of the management on specific species. Both how the abundance of individuals may change and how the browsing impact may differ between the grazed and ungrazed sites. This is partly done by a two-way ANOVA-test with grazing and location as classification variables and overstory density as qualitative variable. However, as the propagule pressure may be very different between locations the ANOVA-test are only performed for abundances of oak seedling and saplings in the oak stands on mull soil and the abundances of beech seedlings and saplings in the beech stand. Then the propagule pressure can be regarded as equal in the tests. However, these tests should also be interpreted carefully due to the same reasons as for the species richness of woody species.

Seedlings and saplings (<0.5 m) of oaks are as the only woody species found in all analysed sites (Figure 32). The highest numbers of individuals are found in the oak stands and with an average of more than 2,000 seedlings per 0.1 ha in both the grazed and ungrazed sites in Lille Hessemose. Here the highest average number of saplings <0.5 m (>500 individuals per ha) are found as well. Only in the young oak stands at Sandskredssøen (P < 0.01) and the beech stand in Store Hessemose (P < 0.05) grazing is found to significantly decrease the number of individuals of seedlings, though a clear but not significant difference with more individuals in the ungrazed site is found for the beech stand at Sandskredssøen as well. Furthermore, the number of saplings <0.5 m is also significantly lower in the grazed site in the young oak stand at Sandskredssøen (P < 0.05). No saplings between 0.5 m and 2.0 m are found in any sites. An ANOVA-test is not able significantly explain the variance in neither the number of oak seedlings nor saplings by any of the factors or the interactions (Table 3).

The obviously highest average number of seedlings of beech with a mean of <250 individuals per 0.1 ha is found in the grazed old oak stand at Sandskredssøen. Here, the density is found significantly increased compared to the ungrazed site (P < 0.01) (Figure 33). The beech stand in Kollerup Enghave has also a significant difference in the number of seedlings (P < 0.05) but with a decreasing by grazing. The same significant relationship is also found for the saplings <0.5 m in the beech stand at Kollerup Enghave (P < 0.05), while the difference for the larger saplings (0.5 – 2.0 m) not is found significant even though it seems evidently (Figure 33). The ANOVA-test finds neither any of the factors nor the interactions to be able to significant explain the variations in number of seedlings (Table 3). For the number saplings <0.5 are all the factors and interactions however significant. This may very well be due to the great variation between locations, the significantly difference between the grazed and ungrazed site in Kollerup Enghave and the fact that the lowest numbers of saplings <0.5 m are found in both the grazed and ungrazed sites in Store Hessemose (Figure 33) and that these sites at the same time have the highest overstory density (Figure 12).

38

Alnus incana was the dominant woody species in all samples in both the grazed and ungrazed swamps in Store Hessemose, while A. glutinosa were dominating in some of the samples in the alder swamp in Kollerups Engahve (Appendix 3). Thus a comparison is challenged. Nevertheless, clear differences in the abundance of saplings are found between the grazed and ungrazed swamps in Store Hessemose with more of the smaller saplings in the grazed site (P < 0.001) while more of the larger in the ungrazed site (P < 0.05).

Individuals of oak Individuals of beech Seedlings Saplings <0.5 m Seedlings Saplings <0.5 m Saplings 0.5-2.0 m Grazing 0.382 0.988 0.687 <0.001 0,046 Location 0.633 0.239 0.442 <0.001 0,151 Overstory density 0.799 0.944 0.088 <0.001 0,052 Grazing*Location 0.126 0.742 0.914 0.001 0,372 Grazing*Overstory 0.455 0.997 0.718 <0.001 0,051 Table 3: P-values for ANOVA-tests with grazing and location as classification variables and overstory density as quantitative variable. Significant values are shown in bold.

Quercus robur NS 5000 Seedlings5000 Saplings <0,5 m Grazed 3000 Ungrazed

2 3000 NS m 1000 1000 0 NS 0 350 350 0 300 1

r 250

e 250

p 200

**

s 150 l 150

a * 100 NS u

d 50 i 50

v * i †

d 10

n 10 NS I NS NS NS NS 5 NS 5 0 0 ) ) ) ) E H E y o H E S H E H E y o H E S H K S K ( ( L K h S K S K ( ( L K h S r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e O O B B O O B B Figure 32: Average numbers (±SE) of seedlings and saplings (<0.5 m) of Quercus robur, No saplings with a size between 0.5 m and 2.0 m are present in any plots. †P < 0.1; *P < 0.05; **P < 0.01; ***P < 0.001.

Fagus sylvatica 1500 Grazed Seedlings Saplings <0,5 m * Saplings 0,5-2,0 m Ungrazed 1000 NS 2 NS

m 500 **

0 100 NS 0

0 * 1 80 NS

r

e 60 NS p

NS s l 40 a u

d 20 i v i

d 6 NS NS

n NS I 4 NS NS † 2 NS 0 ) ) ) ) ) ) E H E y H E S H E H E y H E S H E H E y H E S H K S K ( (o L K S K S K ( (o L K S K S K ( (o L K S r S h r S h r S h e r k S k h c h e r k S k h c h e r k S k h c h d e a k k a c e c d e a k k a c e c d e a k k a c e c l ld O a a O e e e l ld O a a O e e e l ld O a a O e e e A A O O e B e A A O O e B e A A O O e B e B B B B B B Figure 33: Average numbers (±SE) of seedlings and saplings (<0.5 m and 0.5 m – 2.0 m) of Fagus sylvatica. †P < 0.1; *P < 0.05; **P < 0.01; ***P < 0.001.

39

Alnus incana 600 * Grazed Seedlings Saplings <0,5 m Ungrazed

2 Saplings 0,5-2,0 m

m 500

0 0

0 400 1

r

e *** p

300 s l a

u 200 d i v i d

n 100 I NS 0 ) ) ) ) ) ) E H E y o H E S H E H E y o H E S H E H E y o H E S H K S K ( ( L K h S K S K ( ( L K h S K S K ( ( L K h S r r k S S k h c h r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e A A O O e B e O O B B O O B B O O B B Figure 34: Average numbers (±SE) of seedlings and saplings (<0.5 m and 0.5 m – 2.0 m) of Alnus incana. †P < 0.1; *P < 0.05; **P < 0.01; ***P < 0.001.

5.2.2. Browsing impact Even though the browsing impact is measured on an ordinal scale, and the scale may be non-linear, the mean browsing impact for woody species in all quadrats are calculated. Theoretically it is not meaningful to calculate means on an ordinal scale, but the approach is applied for convenience to investigate the browsing impact. As the variances in the number of individuals are massive and to avoid that single individuals have too high a determination rate on the total browsing impact, the average browsing impact for a plot is given a relative weight in relation to the number of individuals prior to any comparison. The relative weight is given as LOG(individuals) and not as the directly number of individuals. Otherwise, the plots with most individuals would in several instances be the only determinant. The average browsing impact is only calculated for species found in three or more plots in both grazed and ungrazed sites. The browsing impact is not measured for any of the seedlings. With possibly only two leaves and one bud present the difference between light and heavy browsing will be not observable or not present. Furthermore, the possibly browsed individuals with only a tiny stem left will be almost impossible to detect and thus the browsing impact will be highly biased.

In the horse grazed areas the browsing impact is significantly increased for saplings of beech (P < 0.001), Alnus incana (< 0.001) and Euonymus europaeus (P < 0.01) (Figure 35). For all other species no significant differences are found. However, it may be of interest to examine whether the browsed individuals are more heavily browsed in the grazed areas. Hence, the same analyses are redone, but with all the unbrowsed individuals left out. In general, the same patterns are found. However, when only the browsed individuals are examined a tendency (P < 0.1) of heavier browsing of saplings of Sorbus aucuparia appears (Figure 36).

40

4 Grazed *** Ungrazed t

c 3

a *** NS ** p m

i *

g 2 * NS n i NS s NS w o

r NS

B 1 NS NS NS NS NS 0 m m m m m m m m m m m m m m m ,5 ,5 ,0 ,5 ,5 ,5 ,5 ,5 ,5 ,0 ,5 ,0 ,5 ,5 ,5 0 0 -2 0 0 0 0 0 0 -2 0 -2 0 0 0 < < 5 < < < < < < 5 < 5 < < < r a , la s ia . . s , a , s s . bu ic 0 u n r sp sp ie 0 n 0 u lu sp o t ca d ce pa s s b es a a ae tu x r va ti n s u u u a i nc an p e ri s yl a e e c g n a b i c o b a su s lv p b u e ru e a s in ur s L r s y la u a ta P ic a u e u ue u s tu p s ra P ce ln us s in Q g s e la bu C i A ln u rp a u B tu r P A m a F ag e o y C F B S n uo E Figure 35: Average browsing impact (±SE) on saplings in two size categories; <0.5 m and between 0.5 m and 2.0 m. The browsing impact is measured on a five-step scale: 0 referring to no browsing being observed, 1 to light browsing of few leaves and buds, 2 to medium browsing of leaves, buds and twigs, 3 to heavy browsing of leaves, buds and twigs, and 4 to nearly all leaves, buds and twigs within reach and some smaller or even larger branches being eaten or disrupted. Only species that occurred in three or more plots in both the grazed and ungrazed sites are included. †P < 0.1; *P < 0.05; **P < 0.01; ***P < 0.001.

4 Grazed *** Ungrazed **

t ***

c 3

a NS

p *

m † i

NS * g 2 NS n i NS NS s w o

r NS

B 1

0 m m m m m m m m m m m m ,5 ,5 ,0 ,5 ,5 ,5 ,5 ,5 ,5 ,0 ,5 ,5 0 0 -2 0 0 0 0 0 0 -2 0 0 < < 5 < < < < < < 5 < < r a , s ia . . s a , s . bu ic 0 n r sp sp ie n 0 lu sp o t ca ce pa s s b a a tu x r va ti s u u u a nc an e ri s l a e c g n a i c b a su sy lv b u e ru e s in s L r s y u a ta P ic u u ue u s p s ra P ln us in Q g s la bu C A ln rp a u tu r A a F ag e o C F B S Figure 36: Average browsing impact (±SE) on those saplings which are subject to browsing. The saplings are categorized into two size groups; <0.5 m and between 0.5 m and 2 m. The browsing impact is measured on a five-step scale: 0 referring to no browsing being observed, 1 to light browsing of few leaves and buds, 2 to medium browsing of leaves, buds and twigs, 3 to heavy browsing of leaves, buds and twigs, and 4 to nearly all leaves, buds and twigs within reach and some smaller or even larger branches being eaten or disrupted. Only species that occurred in three or more plots in both the grazed and ungrazed sites are included. †P < 0.1; *P < 0.05; **P < 0.01; ***P < 0.001.

As Sandskredssøen is grazed all year round, the browsing impact within this enclosure may be higher compared to the other grazed areas due to the prolonged period and the possibility of seasonal variations in the forage selection, thus, browsing may increase during the winter as the availability for other forage may be scarce. To make a fair comparison between the browsing impact at Sandskredssøen and in the other enclosures only woody species which are present with at least five individuals in all three quadrats are chosen – hence, only oak and beech. However, as the browsing impact on beech saplings appears to

41 depend on the size of the saplings (Figure 35 and Figure 36) the two size groups are not combined. Only saplings <0.5 m of beech are compared as the abundance of saplings between 0.5 m and 2.0 m are too low to meet the comparison criteria in all grazed sites at Sandskredssøen.

Even though major differences in the average percentage of individuals of oak seedlings subjected to browsing are observed, with more than 55 % in the young oak stand at sandskredssøen and just above 1 % in Lille Hessemose, no significant differences are found (one-way-ANOVA) (Figure 37). Moreover, no significant differences are found in the browsing impact, neither when all individuals are compared (Figure 38) nor when only those subjected to browsing are (Figure 39). For the beech seedlings <0.5 m, more than 99 % of the individuals in beech stand in Kollerup Enghave are subjected to browsing. This is significantly more than in the all year round grazed beech stand at Sandskredssøen (P < 0.05) where an average of 53 % of the saplings is browsed (Figure 40). This difference also results in a significantly higher browsing impact in Kollerup Engave (P < 0.01) (Figure 41), but also those individuals subjected to browsing are significantly more browsed in Kollerup Enghave (P < 0.01) (Figure 42).

100 2.5 2.0 A d A A e

80 2.0 t t s

c 1.5 c w a a o A A p p r m b i m

60 1.5 i

s A g l A g n a 1.0 i n i

u A s s d i

40 1.0 w w v i o o r d r B n B i 0.5 20 0.5 % A A A A 0 0.0 0.0 Oak KE Oak S(y) Oak S(o) Oak LH Oak KE Oak S(y) Oak S(o) Oak LH Oak KE Oak S(y) Oak S(o) Oak LH

Figure 37: %-individuals (±SE) of saplings Figure 38: Average browsing impact (±SE) Figure 39: Average browsing impact (±SE) <0.5 m of Quercus robur that are subject on those saplings <0.5 of Quercus robur on all saplings <0.5 of Quercus robur. to browsing. that are subject to browsing.

AB A 4 4 100 A A AB d e t 80 t AB s AB c c 3 3 w a a AB

o B p p r B

b B m m i i 60

s l g g

a 2 2 n n i i u s s B d i 40 w w v i o B o B d r r n i B B

1 20 1 %

0 0 0 Oak KE Oak S(o) Beech S Beech KE Oak KE Oak S(o) Beech S Beech KE Oak KE Oak S(o) Beech S Beech KE

Figure 40: %-individuals (±SE) of saplings Figure 41: Average browsing impact (±SE) Figure 42: Average browsing impact (±SE) <0.5 m of Fagus sylvatica that are subject on those saplings <0.5 of Fagus sylvatica on all saplings <0.5 of Fagus sylvatica. to browsing. that are subject to browsing.

5.2.3. Insect pollinated woody species Data of woody species pollinated by insects are extracted from the data of the 15 m circles to test the impact induced by the horses on the diversity. From all the woody species found (Appendix 2 and Appendix 3), species belonging to the families Rosaceae, Grossulariaceae, Salicaceae, Celastraceae, Rhamnaceae, Caprifoliaceae and Aceraceae are regarded as insect pollinated. The comparisons are made at genus level as seedlings and small saplings of some genera may be complicated to identify to species level – particulary species of Salix and Rosa. The only significant differences are found in the alder swamp in Store Hessemose

42

(P < 0.05) and the oak stand in Lille Hessemose (P < 0.05), in both cases with an increased diversity in the grazed sites (Figure 43). A tendency for higher genus richness (P < 0.1) in the grazed site is also found in the beech stand in Kollerup Enghave, though not significant. However, when looking at the seedlings and two sapling groups independently, it is only in the oak stand in Lille Hessemose where the increased richness is found in all three size categories (Figure 44). In the beech stand in Kollerup Enghave a tendency appears also for the seedlings, though not significant. However, the same low richness level is found for saplings <0.5 m, and none has yet survived to an size above 0.5 m (Figure 44). No more than two genera of insect pollinated trees >2 m were registered in any the sites. Hence, the richness within the sites does not appear to be the determinant.

8 Grazed * Ungrazed

s 6 s e n h c

i * † r

4 NS s u n

e NS NS G 2 NS

0 ) ) E H E H y o E S H K S K L ( ( K h S r r k k S S h c h e e a a k k c e c ld ld a a e e e A A O O e B e O O B B Figure 43: Average richness (±SE) of genera of insect pollinated woody species. †P < 0.1; *P < 0.05; **P < 0.01; ***P < 0.001.

8 Seedlings Saplings <0,5 m Saplings 0,5-2,0 m Grazed Ungrazed †

s 6 s e n h c

i NS r

4 * NS s

u NS

n † e NS NS NS G 2 NS NS NS NS NS NS NS NS NS 0 ) ) ) ) ) ) E H E H y o E S H E H E H y o E S H E H E H y o E S H K S K L ( ( K h S K S K L ( ( K h S K S K L ( ( K h S r r k k S S h c h r r k k S S h c h r r k k S S h c h e e a a k k c e c e e a a k k c e c e e a a k k c e c ld ld a a e e e ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e A A O O e B e O O B B O O B B O O B B Figure 44: Average richness (±SE) of seedlings and saplings of genera of insect pollinated woody species. †P < 0.1; *P < 0.05; **P < 0.01; ***P < 0.001.

5.2.4. Deadwood No significant differences in the average total amount of deadwood are found between any of the pairwise comparable sites (Figure 45). However, variations between the locations occur obviously and that the highest amounts are in general are found in the alder swamps and the nearly 200 year old oak woodland at

43

Sandsskredssøen. A distinct difference between the types of deadwood does also appear, with a much higher percentage of the total amount of deadwood being lying than standing. The only exception is in the grazed oak woodland in Lille Hessemose, where a higher amount is found standing than lying.

300 NS Grazed - Standing 200 Grazed - Lying

1 Ungrazed - Standing - 100 a

h 80 Ungrazed - Lying

3 NS m

d 60 o o w

d 40

a NS e D 20 NS NS NS

0 ) ) E H E y o H E S H K S K ( ( L K h S r r k S S k h c h e e a k k a c e c ld ld O a a O e e e A A O O e B e B B Figure 45: Average amount of deadwood (±SE) categorized in standing and lying. The bars are stacked, thus, the tops are equivalent to the average total amounts. Error bars represents the total amount of deadwood. †P < 0.1; *P < 0.05; **P < 0.01; ***P < 0.001.

5.2.5. Other forest structures; epiphytic bryophytes and lichens, hollowness, decay and woodpecker holes The number of trunks with decaying parts (Figure 46) and the number of trunks with hollowness (Figure 47) exhibits the same patterns as the amount of deadwood with higher abundance of these structures in the alder swamps and generally low abundances in the other areas. However, no significant differences are found between any of the pairwise comparable sites. No more than a maximum of one trunk with woodpecker holes is found per 15 m circle. Thus, any meaningful comparison can hardly be made. Nevertheless, it appears that the number of trunks with woodpecker holes is correlated with the number of trunks with decaying parts as a total of respectively two and three trunks with hole are found in the grazed and ungrazed alder swamp in Store Hessemose. In the alder swamp in Kollerup Enghave and the grazed old oak stand at Sandskredssøen are a single trunk with woodpecker holes found while one on in all the other sites. In general, very few trunks with epiphytic bryophytes and/or lichens are found in the beech stands. Significantly higher numbers of trunks with epiphytes were found in the ungrazed sites in the young oak stand at Sandskredssøen (P < 0.05) and the oak stand in Lille Hessemose (P < 0.05) (Figure 48).

44

25 Grazed 5 Grazed a a NS h Ungrazed h Ungrazed NS

1 1 , , s 0

t 20

0 4

r

r s r a e s e p p

e p

g n s

15 s

n 3 i k w k y n o n l l a u u r o c r t t

h

e f

10 f 2 d o h o

t

i NS r h r t

i NS e w e b b w 5 1 NS NS m m

u NS NS NS NS u N N 0 0 ) ) E H E y H E S H E ) ) K S K (o ( L K S H E o y H E S H r S S h K S K ( ( L K h S e r k k h c h r r k S S k h c h d e a k k a c e c e e a k k a c e c l ld O a a O e e e ld ld a a e e e A A O O e B e A A O O e B e B B O O B B Figure 46: Average number of trunks (±SE) per 0.1 ha with Figure 47: Average number of trunks (±SE) per 0.1 ha with occurrence of decaying parts (areas >100 cm2 and >0.5 m above hollowness, defined as a gap in the bark with underlying decay ground level). or hollowness to a depth of at least 5 cm. †P < 0.1; *P < 0.05; **P < 0.01; ***P < 0.001. †P < 0.1; *P < 0.05; **P < 0.01; ***P < 0.001.

s 15 Grazed a n NS h e Ungrazed

h 1 , c i l 0

r r

o * e / p

d 10

n s a k

n s u

e * r t t

y

f NS h o

p 5 r o e y b r NS b m NS NS u h t i N w 0 ) ) E H E y o H E S H K S K ( ( L K h S r r k S S k h c h e e a k k a c e c ld ld a a e e e A A O O e B e O O B B Figure 48: Average number of trunks (±SE) per 0.1 ha with occurrence of larger areas with bryophytes or lichens (<2 m above ground level). †P < 0.1; *P < 0.05; **P < 0.01; ***P < 0.001.

5.3. Forest condition Although the species richness and diversity of vascular plants may be increased by grazing of large herbivores, the trend is not necessarily positive. The species added to the species pool can potentially be invasive or in other ways undesirable. Denmark has developed a system for mapping and condition assessment of terrestrial habitats (Fredshavn & Ejrnæs, 2007; Fredshavn et al., 2008). Besides condition assessment the system can be used to set objectives and provide a basis for prioritizing the effort. A common reference scale for habitat conditions of the individual habitats is used – called a nature index. The endpoints of the reference scale are respectively corresponding to the best and worst examples of, in the case in point, woodland conditions in Denmark, based on existing knowledge about the habitats and their regional differences (Fredshavn et al., 2008). The reference scale is a continuous scale from 0 to 1, which can be translated into five woodland condition classes reflecting the Habitat Directives requirement for woodland structure and function, where only the first two classes meets the Habitat Directives

45 requirement for a favourable conservation status (Fredshavn et al., 2008). The condition class is calculated from a structure index and a species index. However, the structure indexes are not calculated as not all the necessary factors have been measured. Hence, neither the nature indexes are calculated.

The species indexes are calculated as the weighted average of a species score index and a species diversity index. Both indexes are calculated on basis of the composition of vascular plants species found in the 5 m circle as well as the composition of woody species found in the 15 m circle. The species contributes to the calculations by their specific values, a score between 1 and 7. High values are awarded species which are very sensitive towards a negative impact in the habitat, while species with a low value are more or less favoured by these impacts. A number of problem species are designated as well. Problem species are promoted by a strong negative impact in the habitat. In both indexes are problem species as well as invasive species given the value -1, while introduced and non-native species are given the value 0. The species score index is calculated as the average species score, no matter the number of species found. The species diversity index is calculated as the sum of the species scores adjusted for the average species diversity of the habitat.

Tendencies (P < 0.1) toward a decrease in the species index by the grazing management are found both in the oak stand in Lille Hessemose, the young oak stand at Sandskredssøen and in the beech stand at Sandskredssøen (Figure 49). The only significant difference is found in the beech stand in Store Hessemose where the grazing management has increased the species index (P < 0.01) (Figure 49).

1.0 Grazed Ungrazed 0.8 † NS ** x

e † † d

n 0.6 i NS NS s e i c

e 0.4 p S 0.2

0.0 ) ) E H E y o H E S H K S K ( ( L K h S r r k S S k h c h e e a k k a c e c ld ld a a e e e A A O O e B e O O B B Figure 49: Average values (±SE) of the species index of the sites studied. The colours refers to the five condition classes – red being the worst condition while only the green once meets the Habitat Directives requirement for a favourable conservation status. †P < 0.1; *P < 0.05; **P < 0.01; ***P < 0.001.

6. Discussion An import assumption in this study is that the background grazing and browsing impact performed by the deer population is equal in all the pairwise comparable sites. However, grazing management is often estimated to have negative influences on deer populations. The domesticated herbivores are more or less keeping deer away and the fences act as barriers to their activities (Buttenschøn & Holst, 1999). But the problem depends on the size of the fence, the opportunity for coverage and the choice of herbivore

46 species, where grazing with horses or cattle seem to be more pleasant for the deer population than grazing with sheep for instance (Buttenschøn & Holst, 1999). The fences in Gribskov consist of two wires and are possible to pass through by both roe deer and fallow deer (pers. obs.), thus they do not appear to pose any obstacle for the deer population. Moreover, all the fenced areas are of a relatively large size with plenty of coverage possibilities. On the positive side, creation and maintenance of a relatively short lawn structure by the horses will increase the tiller density and forage quality which are preferred by the deer (van Wieren & Bakker, 2006; Buttenschøn et al., 2009b). In this way grazing with large herbivores may also facilitate more selective herbivore species such as roe deer and red deer (Bokdam et al., 2001). In a study of a cattle grazed meadow and heathland it was the presence of preferred forage which to a large extent determined the deer’s use of the livestock grazed areas, no matter if the cattle were present or not (Buttenschøn et al., 2009b), and the same results have been reported for horses as well (Buttenschøn et al., 2009b). With all these information in mind, it seems very likely that the assumption of equal browsing impact by the deer population in the grazed and ungrazed sites can be approved.

The effects of the horses on both the ground vegetation and the composition and abundance of trees and shrubs, and thus the forest structures too, are complex. Multifaceted interactions exist between the dynamics of large herbivores, grazing and browsing intensity and timing, vegetation composition and the abiotic conditions. Even though grazing, browsing and trampling probably are the most important direct impacts of large herbivores and affect all key stages of plant development, through germination, establishment, growth and seed production (Hester et al., 2006), no straightforward generalizations are possible regarding the immediate effects of herbivores on plant growth and resource allocation. Consequences of tissue damage are under the complex control of plant genetics, intensity and frequency of herbivore effects, plant developmental stage at the time of herbivore impact and plant tissue that is affected (McNaughton, 1979; Huntly, 1991; Rosenthal & Kotanen, 1994). Hence, different plant species have different tolerances toward removal of plant material, some being exceptionally sensitive to even a low grazing intensity while others are much more disturbance tolerant. Thus, grazing and browsing by large herbivores may change the interspecific competition among the plant species present, altering the vegetation composition and ultimately the forest structures. But the effect of grazing is also highly influenced by the abiotic conditions and modified by factors as light intensity, temperature as well as water and nutrient availability (McNaughton, 1979; Rosenthal & Kotanen, 1994). Nevertheless, large herbivores may also alter these abiotic factors by their grazing and browsing activity (Adams, 1975; Putman, 1986; Hobbs, 1996; Bardgett, 2005; Skarpe & Hester, 2008; Holtmeier, 2015), which further increases the complexity in the herbivore-plant community interactions. Understanding these complex interactions provides insight into vegetation dynamics in herbivore-driven ecosystems. Hence, the differences and alterations in the interspecific competition and coexistence which is mediated through differences in plant resistance and impact of herbivory together with the abiotic conditions and potential changes in propagule pressure can explain the observed patterns and changes in the ground vegetation and forest structures in this study.

6.1. Ground vegetation In spite of the relatively short time with grazing management, the horses have in several cases had significant effects on the ground vegetation in one way or the other. In some sites an increase in vascular

47 plant species richness has followed since the introduction of the horses whereas in others no changes have occurred (Figure 6 and Figure 7). The altered richness as well as the changed abundances of some species are a matter of how the horses affect the individual plant populations and communities. Plant populations and communities contain mixtures of life stages, and in the latter case species, which compete for resources but exhibit a range of different traits, leading to both differential damage by the horses and differential responses to that damage (Hester et al., 2006). These interactions between the horses’ impacts and plant competition are the key to understanding the effects on the plant communities induced by the horses (Hester et al., 2006), the variations in the effect on species richness (Figure 6 and Figure 7) and the increased and decreased abundance of some species (Figure 21, Figure 22 and Figure 23).

6.1.1. Species richness and density Under conditions of low environmental stress or low intensity of management (e.g. grazing) productivity is high and plant species with a high competitive ability attain maximum vigour (Figure 50). This results in a low species density and species richness as a result of competitive exclusion (Grime, 1973). Although herbivory may kill small plants, the more common effect of the removal of leaf area, with a reduction in the photosynthetic active area, is a reduction in growth rate and resource uptake. But it may also cause the loss of significant amounts of nitrogen and/or carbon reserves, which further may inhibit growth (Putman, 1986; Rosenthal & Kotanen, 1994; Crawley, 1997; Hester et al., 2006). This affects competition between plants and can result in a replacement of highly impacted plants by species that are less affected (Crawley, 1997; Hester et al., 2006). If the horses are able to reduce the competitive interactions between the plant species, then their presence is likely to allow greater coexistence between species and thus increase the richness (Huntly, 1991; Olff & Ritchie, 1998; Hester et al., 2006). However, under prolonged heavy grazing species richness may decline, as only a small number of species are tolerant enough to withstand the form of damage (Olff & Ritchie, 1998; Hester et al., 2006). This concords with the classic Intermediate Disturbance Model (Grime, 1973) which predicts that plant species richness is highest at intermediate levels of disturbance as interspecific competition is reduced. Nevertheless, if competitive ability and resistance trait toward herbivory are positively related, then the presence of herbivores can increase the dominance of an already dominant species, which will reduce diversity by preferentially damaging those species that already suffer a competitive disadvantage (Huntly, 1991; Rosenthal & Kotanen, 1994; Olff & Ritchie, 1998). However, the latter does not seem to be the case in any of the grazed sites in this study, although occurrences where species with a high competitive ability have increased their relative abundance by grazing are found, e.g. Urtica dioica (Figure 21 and Figure 22) and Phalaris arundinaea (Figure 21). Although these species have increased their relative abundance they have, however, not become dominant, and most importantly, it has not happened at the sacrifice of the species richness which has increased significantly in all cases (Figure 7). The possible reasons for the increased abundance of these species will be discussed later.

48

Figure 50: Diagrams representing impact upon species density of (a) intensity of environmental stress and (b) intensity of grazing, mowing, etc. C, Species with high competitive ability. S, Species of high resistance to the prevailing stresses imposed by environment or management respectively. R, Remaining species (Grime, 1973).

Due to the intermediate disturbance hypothesis (Figure 50), a decrease in the communities’ competitiveness by grazing could be expected. This is however not the general condition found (Figure 24), as competitors with grazing tolerant traits, such as U. dioica and P. arundinaea, are also invading the grazed areas (Figure 21 and Figure 22). But with tendencies towards an increased evenness in those areas which benefits from the grazing management by enhanced species richness (Figure 7), it emphasizes that reductions in the interspecific competition have occurred. Moreover, as expected those species invading the area, and the species present, which increase their abundance, are in general more disturbances tolerant (Figure 26). However, the significant decrease in the S-dimension (Figure 25) in the grazed alder swamp in Store Hessemose, can neither be explained by differences in soil moisture, soil pH nor plant available nitrogen as no differences occur. Instead it may be explained by the significantly higher abundance of Poa trivialis which is a competitive-disturbance tolerant species and thus highly impact the S- dimension. The enhanced abundance may very well be due to the grazing management and the increased amount of light, as also (Vestergaard, 2007) reported a high increase in the abundance of P. trivialis in a mowed meadow after just three years management. Another species which also highly impact the S- dimension is the stress-tolerant Milium effusum. But whether the higher abundance of M. effusum in the ungrazed site (Figure 21) is due to the grazing management or the higher overstory density is, however, more difficult to disentangle. Thus, it is also difficult to disentangle whether the difference in the S- dimension is due to the overstory density or the grazing management.

The alterations, or the lack of it, in the evenness of the present plant species (Figure 20) are also reflected in Simpson’s diversity index (Figure 19), where e.g. no change in the diversity is found in the oak stand in Lille Hessemose, though a significantly increased richness, due to the unaltered species richness (Figure 7). The opposite situation is found in the young oak stand at Sandskredssøen. Thus, these results stress the importance of a common understanding of the term “diversity”, as it is often used interchangeably with

49 species richness. Nevertheless, in both the grazed alder swamp in Store Hessemose and the old oak stand at Sandskredssøen a higher evenness is found as expected. The unaltered evenness in e.g. the oak stand in Lille Hessemose, though dominant species like Calamagrostis canescens are decreasing abundance, may be due to the higher species richness of which some of the newcomers still have a low abundance. The higher diversity in the ungrazed oak stand in Kollerup Enghave is clearly due to the increased dominance of a few species in the grazed part.

The removal of aboveground biomass of the ground vegetation by the horses, which is obvious and significant in the alder swamps and the oak stands (Figure 18), will subsequently reduce the amount of litter produced. As the germination and establishment of plants are particularly sensitive to the presence of litter (Facelli & Pickett, 1991; Crawley, 1997; Xiong & Nilsson, 1999), the reduced vegetation height and, hence, reduced amount of litter, may partly be an important explanation for the increased species richness observed in all the grazed sites where the base was a rather tall ground vegetation (Figure 7), except in the young oak stand at Sandskredssøen where no increase in the species richness is found. The way the litter mat on the forest floor can affect the species richness and plant community structure are diverse and can both act directly and indirectly (Facelli & Pickett, 1991). The direct effects are through an inhibition of germination and establishment of plants (Xiong & Nilsson, 1999) as it constitutes a physical barrier for seeds, seedlings and shoots. Seeds retained in the litter may either have delayed or obstructed germination and seedlings and sprouts emerging from beneath a litter mat must spend significant amounts of energy and time to penetrate it (Facelli & Pickett, 1991), with the possibility of morphological and physiological characters that lessen their capacity to fix carbon in result (Knapp & Seastedt, 1986). Hence, seed size and shape (Facelli & Pickett, 1991), and seedling or shoot morphology and phenology may determine the ability of a plant to tolerate litter (Facelli & Pickett, 1991; Koorem et al., 2011). Species such as Pteridium aquilinum and Utrica dioica may achieve competitive dominance due to their shoots robust structure capable of penetrating a thick litter layer and expanding the leaves above (Grime, 1979). Seedlings originating from small seeds may be unable to penetrate the litter mat or tolerate shading, but those germinating from larger seeds (a character often found in late successional species) may successfully cope with dense litter mats (Facelli & Pickett, 1991). However the movement of large seeds through the litter may be impeded or delayed and the seeds may remain in an unsuitable germination site. Even if the larger seeds are able to penetrate the litter mat, seedlings originating from large seeds can be confined within the shady and wet environment of the litter mat, increasing the risk of fungal infection or herbivore attack (Facelli & Pickett, 1991). The indirect effects of the litter are through changes in the resource availability and through the effects on other biotic components. The presence of litter alters the micro-environmental conditions of the top of the soil. Litter intercepts incident light and rain and changes the surface structure, affecting the transfer of heat and water between the soil and the atmosphere (Facelli & Pickett, 1991). Shading by litter is followed by an exponential reduction of the light intercepted as the amount of litter increases (Facelli & Pickett, 1991) and the total amount of PAR below a dense mat of grass litter can be 1 to 5 % of the total PAR above the litter (Knapp & Seastedt, 1986; Facelli & Pickett, 1991). Many seeds are dependent on exposure to light to break their dormancy (Crawley, 1997), this is the case for e.g. Urtica dioica, Poa trivialis, Stellaria media and Deschampsia cespitosa (Grime, 1979) which partly may explain these species’ increased abundance by grazing in this study (Figure 21 and Figure 22). The lights spectral composition is also important. Many species’ germination are strongly inhibited by exposure to light with a low red:far-red ratio (Crawley, 1997; Thomas, 2014). And as this ratio is decreased when the light passes

50 through a leaf, the lowered vegetation height may increase this ratio for the light reaching the soil, beneficial for many seeds’ germination, even though the canopy may still retain a significant part of the light. Additionally, the insulating effect of the litter may also reduce the soils thermal amplitude which will impair the germination of seeds whose dormancy is broken by alternating temperatures (Facelli & Pickett, 1991; Crawley, 1997), this includes e.g. the grasses Poa trivialis, Holcus lanatus and Deschampsia cespitosa (Grime, 1979). The local extinction of many species from oldfields and ungrazed grassland may very well be due to this mechanism (Facelli & Pickett, 1991), thus, it may likely be the case in woodland habitats too. Nevertheless, litter can also modify environmental conditions to have positive effects on seedling growth by maintaining soil moisture, moderating soil temperature, providing nutrients during decomposition and reducing interspecific competition (Facelli & Pickett, 1991; Xiong & Nilsson, 1999; Koorem et al., 2011). However, Xiong and Nilsson (1999) made a meta-analysis among 35 independent studies and found that the negative effects of plant litter generally outweigh the positive ones. It may still be argued that in spite of the high vegetation a big amount of the litter produced within the stands originates from the trees and that this proportion remains unaltered. But the effects of litter depend, however, on the type of litter (Facelli & Pickett, 1991; Xiong & Nilsson, 1999; Koorem et al., 2011), where it has been found that grass and tree litter have differential effects on seedling establishment due to differences in structure (Koorem et al., 2011) and different extinction coefficients (Facelli & Pickett, 1991). Donath and Eckstein (2010) suggested that emergence from below oak litter may be easier compared to grass litter which forms dense mats on the ground, because seedlings may displace oak leaves during emergence. In addition, the decomposition of the litter may be speeded up by the horses, as their trampling will enhance the contact of the litter with the soil (Knapp & Seastedt, 1986; Facelli & Pickett, 1991).

The increased spatial heterogeneity found in this study (Figure 17) may be a consequence of an increased structural heterogeneity induced by the behaviour, selection and trampling effect of the horses. This increased heterogeneity may thus influence the surface-near microclimate by alteration of the vegetation height, density, coverage and albedo (Holtmeier, 2015), creating new microhabitats with new niches that can be occupied by specialized plant species. Ultimately, the lowered interspecific competition, the increased amount of light reaching the forest floor and the higher heterogeneity with more niches will increase the invasibility of the sites, thus increasing the species richness as well. The significantly negative correlation between the vegetation height and the species richness in the alder swamps and the oak stands on mull soil (Table 1) clearly shows that this is the case. However, the effects on habitat heterogeneity, vegetation structure and interspecific competition, and ultimately the species richness, strongly depend on the grazing intensity. In the beech stands the species richness and the vegetation height are found to be significantly positively correlated (Table 1) and the spatial heterogeneity is decreased by the presence of the horses (Figure 17). It indicates a much too high grazing intensity with negative impacts on the species richness. Nevertheless, no changes in the species richness in the beech stands are found (Figure 7). But as herbivory more commonly affects growth rates and nutrient uptake than it directly kills individuals, there may very well be a lag time between the initial grazing management and a species disappearance. This fact, together with a possible high propagule pressure, may explain why the species richness in the beech stands has not decreased yet. However, with the present conditions it appears very likely to happen in the nearest future.

Whether species richness increases may in many cases depend on the scale (both spatial and temporal) at which it is measured (Huntly, 1991; Crawley, 1997; Olff & Ritchie, 1998; Hester et al., 2000; Hester et al.,

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2006). It may be argued that what is measured in the small quadrats (0.1 m2) and the larger ones (78.5 m2) in this study are the same but at different scales. In contrast to the alder swamp in Store Hessemose and the old oak stand at Sandskredssøen, the average number of species in the small quadrates is found to be decreasing by grazing in the oak stand in Lille Hessemose (Figure 8), albeit an increase in the larger quadrats in all sites (Figure 7). Hence, the results indeed constitute a classical example of how diversity and richness is scale dependent. A local increase in the abundance of an already dominant species might locally decrease the species density significantly with an impact on the general density found. However, no species where competition and resistance traits towards herbivory are positively correlated appears to increase its abundance entirely in the grazed oak stand in Lille Hessemose (Figure 14). On the other hand, the variation in the vegetation height seems to increase slightly in the grazed area (Figure 18), and the trampling effect of the horses have indeed created paths with exposed soil without any vegetation (Figure 51). These vegetation-less areas will, if their frequencies are high enough, decrease the overall species density, which appears to be likely in this case, and thus explaining the decreased species density found.

Figure 51: The grazed oak stand in Lille Hessemose with vegetation-less patches.

The effect of horse grazing, however, does also strongly depend on how the horses influence the abiotic conditions and how this possible alteration affects the plant community. Grazing a tall plant will reduce the competitive advantage of that individual in relation to light, but might also increase the amount of light available to other surrounding plants which are not grazed or grazed less. But if grazing on the surrounding plants has a stronger negative effect on water or nutrient uptake, then it might negate the advantage of the increase in light availability. Hence, the degree of impact of herbivores on competition depends very

52 much on how they affect plant uptake of the most limiting resource(s), as well as the physical form of the plant itself (Grime, 1979; Huntly, 1991; Crawley, 1997; Hester et al., 2006). In the beech stands where the overstory densities are high (Figure 12) and limited amount of light penetrates the canopies due to a high LAI of beech trees (Packham et al., 2012), the most limiting factor is probably the amount of light. But as the competitive superior beech trees, and hence the overstory density are not affected by the horses then neither is the interspecific competition. Clear tendencies to differences in the overstory density are found in the oak stands and alder swamps. As the species richness normally will be positively correlated with increasing light in woodlands, due to reasons already explained, the significantly positive correlation between the overstory density and the species richness in the oak stands on mull soil (Table 1) indeed emphasizes the positive impact of the horses. In the alder swamp in Store Hessemose it is however difficult to disentangle the effect of the overstory density and grazing on the species richness. Truly some species may benefit from the lower overstory density, but the high richness found in the grazed alder swamp in Kollerup Enghave, which has the same overstory density as the ungrazed swamp in Store Hessemose, gives hard evidence that a high richness in the alder swamps are possible, despite the overstory density, due to grazing.

But the interactions between abiotic factors and herbivore impacts on plant diversity are not straightforward. The type of vegetation change in relation to herbivory does also vary with nutrient availability, as ratios of the supply of different plant nutrients determine productivity and the tissue characteristics of competitively dominant plants. Tissue characteristics may influence the palatability and thus whether herbivores will prevent competitive exclusion (Olff & Ritchie, 1998; Skarpe & Hester, 2008). Any plant is less palatable when growing in infertile than in fertile soil, due to lower protein content and a higher level of carbon-based defences such as tannins, resins and essential oils (Chapin III et al., 2011), and in addition, particularly in infertile environments, herbivory may lead to a decrease in preferred plant species and an increase in defended unpalatable species (Skarpe & Hester, 2008). In natural systems this would imply a reduction in future grazing or browsing, as animals would select other foraging areas or suffer reduced densities (Hobbs, 1996; Skarpe & Hester, 2008), but in enclosures they do not have these possibilities unless variations in the soil’s concentration of plant available nutrients occur within the grazed area. However, it also strongly depends on tolerance of the herbivore to a reduction in food quality, and may operate with a considerable time lag (Skarpe & Hester, 2008).

Not only the amount of plant available nutrients, but also the soil moisture can be crucial to explain patterns in herbivore effects on plant diversity (Olff & Ritchie, 1998). Milchunas et al. (1988) developed a generalized model of the effects of grazing on community structure which proposed that the primary factors explaining changes in plant species diversity and composition are the interactions that occurs along gradients of evolutionary history of grazing and environmental moisture (Figure 52). To measure the soil moisture as well as soil pH and plant available nitrogen, Ellenberg’s weighted indicator values (Ellenberg et al., 1992) were calculated based on the present flora and its relative abundance. However, the Ellenberg- values are subjectively assessed and determined based on flora in central Europe. The values are based on the ecological optimum for the species, but as some species may occur over wide ecological amplitudes, the values are biased, and they are probably better to determine the considerable differences between dissimilar habitats than differences within the same habitat. Hence, minor differences like some of those found for the soil moisture (Figure 9) may be due to smaller variations in the abundances of specific species rather than actual differences in the edaphic conditions. On the other hand, the values will, however, give

53 an indication of the ecological conditions that are not affected by any seasonal fluctuations. Nevertheless, it also means that since no significant differences are found for neither the soil moisture (Figure 9), the pH (Figure 10) nor the plant available nitrogen (Figure 11) in neither the alder swamps, the oak stands on acidic soil, the oak stands on mull soil nor the beech stands, and as the variations are small too, the significant correlations found between these ecological factors and the species richness and the species density (Table 1) are highly biased. The strong positive correlation between the species density and the relative abundance of Calluna vulgaris in the oak stands on acidic soil (Figure 15) demonstrates this very well. Calluna vulgaris has an Ellenberg-value for nitrogen of 1 (Ellenberg et al., 1992) – far lower than the average weighted Ellenberg-values of 3,5 and 3,1 in respectively the grazed and ungrazed oak stand on acidic soil. If C. vulgaris, however, is not preferred by the horses, either because of its content of secondary plant metabolites or because other plants within the vegetation are more preferred, as it has been reported for cattle (Bokdam, 2003), then other plant species may be protected from grazing by growing in the proximity of C. vulgaris via associational resistance, possibly increasing the overall species density where C. vulgaris are growing. Then a significantly negative correlation between the nitrogen concentration and the species density may hence be found, which indeed is the case (Table 1), although the species density are not necessarily due to the concentrations of plant available nitrogen but rather to the presence or absence of C. vulgaris, which in turn affects the Ellenberg-N value. Hence, conclusions based on the correlations between species richness, species density and the edaphic conditions may be highly biased.

Figure 52: Plant diversity of grassland communities in relation to grazing intensity along gradients of moisture and of evolutionary history of grazing. Increments on the diversity axis are equal in all cases, but equal specific values are not implied; that is, relative response, not absolute diversity is implied (Milchunas et al., 1988).

The model of Milchunas et al. (1988) (Figure 52) has been further expanded in the sense that the axis “moisture” can be substituted by “annual net primary production (NPP)” or “productivity” (Milchunas et al.,

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1988; van Wieren & Bakker, 2006). Although the productivity or biomass of the ground vegetation is not measured directly the vegetation height or alternatively the amount of light reaching the forest floor (determined by the overstory density), as this tends to be the most limiting factor in a woodland system (Petersen & Vestergaard, 2006), may be a good substitution for the “moisture axis” of Milchunas et al. (1988) generalized model. Hence, when “semiarid” is substituted with “low vegetation height” or “high overstory density”, this further emphasizes why the grazing management in the analysed beech stands more than likely will decrease the species richness by time. The lack of development in the species richness in the young oak stand at Sandskredssøen is, however, not explainable by this model as the productivity appears fairly high, though lower than in the stands on more rich soil (Figure 18).

Large herbivores themselves do also have multiple indirect effects on the abiotic factors. In the short term, they enhance nutrient availability by returning large quantities of undigested and non-assimilated nutrients to the soil as faeces, and assimilated nutrients in urine, which effectively short-cuts the litter decomposition pathway and increases the nitrogen mineralization (Adams, 1975; Putman, 1986; Bardgett, 2005; Chapin III et al., 2011; Holtmeier, 2015). Even through urine and faecal deposition directly onto plants can cause physical damage and local toxic effects (Hester et al., 2006), the main effect are indirect through nutrient recycling and seed dispersal (Hobbs, 1996; Hester et al., 2006). As up to 90% of the consumed nutrients are returned through urine and faeces (Buttenschøn, 2007; Friis Møller, 2010) to an area smaller than the original, the nutrients will be spatially heterogeneous concentrated which may increase regeneration sites and soil heterogeneity (Huntly, 1991; Olff & Ritchie, 1998), further increasing the species richness as already explained. This heterogeneity in the spatial distribution of nutrients within the landscape is further amplified by ungulate selection for habitats and patches. Spatial segregation of foraging and resting sites generates nutrient transport between habitats, and this spatial heterogeneous urine and faecal deposition may by time cause a reallocation of nutrients within the grazed area. It is however not yet possible to observe neither an increase nor a decrease in the nitrogen content in the soil in any of the grazed sites (Figure 11). Horses are often reported to deposit faeces in latrines (Mayle, 1999; Tolhurst & Oates, 2001; Buttenschøn, 2007), but the use of latrines and the degree of subsequent grazing avoidance of the area depends on the breed (Tolhurst & Oates, 2001). For many feral horse populations, latrine areas and avoidance are not reported (Kolter et al., 1999), an exception is in New Forest. But also here, during winters with food scarcity, the horses are foraging on the latrine areas, and thus keep the vegetation height relative short (Putman, 1986). An increased variation in the amount of nitrogen could indicate an accumulation in certain areas and thus deposition of faeces in latrines. However, this is not found (Figure 11). There has not been found any increased variation in the vegetation height (Figure 18) either, which could indicate avoidance of certain areas.

What has been discussed until now is how the horses can affect the species richness. But increased invasibility cannot increase the species richness alone. In basic terms plant species richness is the results of the balance between the local colonization and extinction rates of species, and the mechanisms that influence these rates (Crawley, 1997; Olff & Ritchie, 1998). Thus any impacts that horses have on these processes will have effects on the species richness, with the highest richness occurring when the local extinction rate are lower than the local colonization rate (Olff & Ritchie, 1998). But importantly, even though the horses may increase the local invasibility, the colonization rate is solely depended on the number and richness of propagules, which are a matter of the size and richness of the seed bank and the landscape in which the site is situated and thus the possibilities for propagules to be dispersed to the site.

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The lack of response in species richness to the horse grazing in the young oak stand at Sandskredssøen still miss an explanation. Relative to the intermediate disturbance hypothesis (Figure 50) it could be indicative of either a too low or too high grazing intensity – considering the change in the vegetation height (Figure 18) probably the latter. However, then a change in the vegetation composition would be expected in response, with a species turnover from species with low resistance to grazing to more disturbance tolerant species. Although an alteration in the vegetation composition has occurred (Figure 15) no changes in communities disturbance tolerance is found (Figure 24). Of the 20 species not shared among the two sites, 14 species were found with such low abundances that their lack of observation in the grazed site, very well may be due to the number of observations as well as to their absence (Appendix 1). Thus the explanation has to be found somewhere else and here the propagule pressure may indeed be an important determinant. The horses may undoubtedly increase the colonization rate by enhanced zoochorous seed dispersal, particularly where gaps in established vegetation occurs (Olff & Ritchie, 1998; Hester et al., 2006). The mechanisms of zoochory are further discussed in the next section, but a high proportion of the invading species may be dispersed form the open habitats within the enclosure, where new species will emerge after the initial grazing management. However, the open area within the enclosure at Sandskredssøen, which is on acidic soil, has, compared to the other open habitats within the other enclosures, not been open for several decades but was a dense plantation of Picea abies until 2006 where the first part was deforested. Since then most of the plantation has been deforested. It may ultimately have the consequence that most of the area including the seed bank has been more or less biologically reset. Thus, the species pool of the open habitat may not be very different from the tree-covered area, and the number of new plant species dispersed by the horses to the oak stand is hence very limited compared to the potential in the other enclosures. However, the old oak stand which is found within the same enclosure had an increase in the species richness of more than 100% (Figure 6). But as this stand is on non-acidic soil and is more than 200 years old, the seed bank will probably, the age in mind, be somewhat larger. Bülow- Olsen (1980) found similar results in a cattle grazed grassland dominated by Deschampsia flexuosa, where no alteration in the richness of vascular plants species was found four years after the initial grazing management. The few new species Bülow-Olsen (1980) found were typically associated with old dung-pats, indicating the important means of zoochorous seed dispersal within the enclosure. Moreover, the community dominated by D. flexuosa may also partly be a possible explanation. Even though Grime (1979) does not classify D. flexuosa as an competitor but as a stress-tolerant species, D. flexuosa is known to outcompete Calluna vulgaris in many heathlands in the eastern part of Denmark (Vestergaard, 2007). Jarvis (1964) also reported it as highly competitive and found likely indications for D. flexuosa excreting root exudates with strong allelopathic impacts on other plant species resulting in a reduction in growth rate, plant weight, leaf area increment, net assimilation rate and ratio of root to shoot growth. Due to the significantly lowered vegetation height, there is no doubt that the horses graze upon D. flexuosa, but importantly no decrease in the abundance has occurred yet (Figure 15 and Appendix 1). Assuming that the unaltered abundance means no major changes in NPP of D. flexuosa, and that the very low availability of nitrogen (Figure 11) are at least as inhibiting as the amount of PAR, then D. flexuosa will still immobilize the same amount of nitrogen and no alteration in the interspecific competition has thus occurred. Altogether it appears plausible that the low propagule pressure and the high abundance of D. flexuosa with its competitive abilities and allelopathic impacts can explain the unaltered species richness. In addition, the low soil pH (Figure 10) may also be an important factor, as the low pH will affect and slow down the decomposition. The consequence is an organic soil with an humus-rich top soil where many seeds are

56 impeded to germinate (Vestergaard, 2007). Moreover, the low soil pH will affect the availability for the most important plant nutrients, nitrogen and phosphor, by decreasing their availability but also increasing the solubility of toxic ions as aluminium (Al3+) (Bardgett, 2005). Conditions where only species that are able to exclude the absorption of these toxic ions or neutralize them in their tissue, are able to establish themselves. In C. vulgaris this is done by the associated ericoid mycorrhizas, in D. flexuosa the arbuscular mycorrhizas are probably important while Carex pilulifera does it by itself (Vestergaard, 2007). These conditions will, however, set a restriction for the number of species able to colonise the habitat which may result in a prolonged lag time between the initial grazing management and an increase in the species richness compared to more productive sites. Last, but by no means least, the different history of the grazed and ungrazed site may of course also be a determinant. This is illustrated by some of those species that the two sites do not have in common, and to which absence or presence cannot be directly attributed to stochastic variables (Appendix 1). For example Pteridium aquilinum, which was only found in the ungrazed site, and although horses in New Forest have been reported to graze significant amounts of P. aquilinum in the autumn when the toxins has declined (Putman, 1986; Putman et al., 1987), and that P. aquilium may suffer from the trampling effect (Mayle, 1999), no such decrease in abundance would be expected by the introduction of the horses. In addition Calamagrostis epigejos is only found in the grazed site, though its abundance is decreasing by grazing in both the old oak stand at Sandskredssøen and the oak stand in Lille Hessemose (Figure 22).

6.1.2. Species specific abundance alterations In this study were Poa trivialis, Deschampsia cespitosa, Ranunculus repens and Phalaris arundinacea found to be significantly more abundant on the horse grazed sites but also Urtica dioica, Glechoma hederacea, Moehringia trinervia, Viola reichenbachiana/V. riviniana, Stellaria media/S. neglecta and Betula sp. displayed the same trend. In contrast, significantly higher abundances of Milium effusum, Stellaria holostea, Calamagrostis epigejos, Calamagrostis canescens and Vaccinium myrtillus were found in the ungrazed site. The same trend was also found for Carex acutiformis, Trientalis europaea and Molinia caerula (Figure 21, Figure 22 and Figure 23). However, as the spatial distribution of populations often are aggregated and these aggregations randomly distributed within a patch (Petersen & Vestergaard, 2006), some of the results, especially those far from significance may be due to this distribution, and with enhanced uncertainty if the grazed and ungrazed site are even not sharing the same history as the case is for the young oak stand at Sandskredssøen. Only by monitoring the development in the vegetation over time or by increasing the number of samples, the underlying dynamics can be revealed. Nevertheless, if the same tendencies are found in more than one case the likelihood that the patterns are due to the management regime increases dramatically. Thus, as U. dioica, M. trinervia, D. cespitosa, G. hederacea, V. reichenbachiana/V. riviniana and S. media/S. neglecta increased their abundances in more than one case and that C. epigejos and C. canescens decreased (Figure 22), it is highly plausible that these patterns are due to the horses’ grazing. This also emphasizes that the higher abundance of C. canescens in the grazed young oak stand at Sandskredssøen (Figure 23) probably is not due to the present of the horses but rather a matter of the absence of C. canescens in the ungrazed part. How individual species are affected may be a matter of whether they are preferred by the horses and to which degree, but also the likelihood of zoochorous seed dispersal and the plant species sensitivity towards herbivory. For example have C. acutiformis (Larsen & Vikstrøm, 1995) and V. myrtillus (Crawley, 1997; Mayle, 1999; Van Uytvanck &

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Hoffmann, 2009) been reported as highly sensitive and R. repens as tolerant to trampling and grazing (Harper, 1957). In this section a smaller review is made of the mechanisms determines whether specific species are selected or avoided together with the factors for successful zoochory. Examples constitute the species found in this study which exhibit clear patterns to the grazing management, hence explaining these species specific abundance alterations (Figure 21 and Figure 23).

Plants can avoid herbivory by either escaping or defencing themselves. Escape mechanisms includes morphological traits such as having short stature, maintaining large proportions of the biomass below herbivore level or keeping the edible biomass above the reach of herbivores. In this way species with stolonifeous growth like e.g. G. hederacea can increase abundance in grazed areas (Hutchings & Price, 1999). But phenological traits that make the plant less apparent or attractive to large herbivores are also an escape mechanism. This includes deciduousness, reduction of the plant to underground organs and survival as seeds (Hester et al., 2006; Skarpe & Hester, 2008), where especially those species with a persistent seed bank like Urtica dioica, Poa trivialis, Deschampsia cespitosa, Ranunculus repens and Stellaria media (Grime, 1979) may take advantage and increase their abundance when the right conditions are present. But escaping is not only determined by the plant’s own characteristics but also by environmental influences. These include the microenvironment, which may affect phenotype and visibility, and the characteristics of the whole plant community, such as the density of the plant species or the neighbouring plants species (Huntly, 1991; Hester et al., 2006). Growing in proximity to neighbouring plants, may reduce the likelihood for herbivory either because the other species are less palatable or more palatable. This depends on the foraging pattern of the herbivore. In this way, associational resistance is an indirect mutualism, as individuals of different plant species, which most obviously may be regarded as competitors, may in fact have a net positive effect on another species. However, the effectiveness of associational resistance can depend on the foraging selectivity of herbivores (Huntly, 1991). Many plants are not grazed or browsed because they are protected either physically by thorns, prickles, thick cuticles, trichosomes or glandular hairs that keep herbivores away (Hester et al., 2006) or chemically by plant secondary metabolites. Plant secondary metabolites are diverse in both chemical structure and their allelopatic interactions with herbivores, and the same chemical defence affects different herbivores in different ways (Freeland & Janzen, 1974). They reduce the nutritive quality of plant material in a number of ways and may be the primary determinant of palatability of plants to herbivores (Huntly, 1991; Hester et al., 2006). Some plant secondary metabolites reduce digestive efficiency and others act at the tissue level following absorption and circulation in the blood. The most important digestibility reducers are the tannins, forming complexes with proteins and, to some extent carbohydrates, leading to reduced protein digestibility by a direct effect on dietary proteins and an indirect effect through suppression of digestive enzyme activity (Gordon & Prins, 2008; Chapin III et al., 2011). A range of other plant secondary metabolites exert negative effect on the digestion through anti-microbial effects in the rumen and hindgut by inhibiting microbial fermentation directly (Freeland & Janzen, 1974; Gordon & Prins, 2008). Still other plant secondary metabolites exert their effects following absorption and distribution to the tissues. These include the alkaloids which can act as neurotoxins (Chapin III et al., 2011) such as the pyrrolizidine alkaloids found in Senecio spp. which form stable pyrroles in the liver and thus act as cumulative hepatotoxins (Freeland & Janzen, 1974). The range of plant secondary metabolites is large but their physiological effects are known in only sketchy details for a few well-studies plants (Gordon & Prins, 2008). Moreover it may be difficult to predict the effects of the plant secondary metabolites as a herbivore which eats a variety of different plant species also may ingest a

58 battery of different active biocides which can have both additive, synergistic and antagonistic interactions (Freeland & Janzen, 1974). However, all resistance traits imply a cost for the plants, and only if the fitness cost of resistance is less than the cost for the herbivory it prevent, will plants with resistance traits have a competitive advantage compared to plants without such traits (Skarpe & Hester, 2008).

Among the species found increased by the horse grazing, Utrica dioica constitutes a classic example of physical defence. The stinging hairs are often assumed to function as a protection against mammalian herbivory, which partly may be an explanation for its increased abundance in the grazed sites in this study. On the other side, Taylor (2009) argues that there is no evidence that U. dioica is harmful to large herbivores, and gives examples of breeds of domestic cattle avoiding U. dioica while other breeds ate them. Also Lamoot et al. (2005) reported both cattle and horses to graze on the stinging plant and red deer have been found to have the same habit too (Krojerova-Prokesova et al., 2010), though it only constituted a minor part of the diet in the two latter cases. However, zoochorous seed dispersal may also be an important explanation for the increased abundance. Clones of U. dioica are able to produce 30,000 seeds per shoot in a pastureland, though fewer in woodlands (Taylor, 2009), and several studies have reported that the seeds of U. dioica are capable of surviving passage through the gut of large herbivores and of seedlings germinating from dung-pats (Grime, 1979; Cosyns et al., 2005; Jaroszewicz et al., 2009; Stroh et al., 2012; Buttenschøn & Buttenschøn, 2013). Cosyns and Hoffmann (2005) even found U. dioica to be the most abundant species in horse dung though it was not at all common in the vegetation. This emphasizes that endozoochory more than likely is an explanation for its increased abundance. Zoochory may also explains the increased abundance of some of the other species and may indeed be an important factor as more than two third of Danish plant species growing in (semi-)natural habitats can survive dispersal by mammals, either by epizoochory or endozoochory (Bruun & Fritzbøger, 2002), where horses have the ability to transport hundreds of thousands of seeds in each year (Stroh et al., 2012). This emphasizes that the horses undoubtedly affect the seed dispersal within the enclosures and that zoochory is an important explanation for the increased species richness. However, to which degree the horses act as vectors cannot be determined by this study, since it is dependent on the number of species within the enclosure, each plant species output of seeds and on the dispersability of the propagules. This again depends on the morphology of the dispersules and the plants (Bruun & Fritzbøger, 2002; Graae, 2002; Couvreur et al., 2005). The germination capacity of seeds passing through the digestive system may either increase decrease if the seeds get destroyed (Crawley, 1997; Gill & Beardall, 2001; Holtmeier, 2015). The success of endozoochorous seed dispersal depends on mechanical and/or chemical scarification of the seed-coat, which may depend upon chewing behaviour and gut retention time, and hence the herbivore species that have eaten the seed (Cosyns et al., 2005). But also plant species and the size and shape of the seed are important for the relative germination success (Gill & Beardall, 2001; Bruun & Fritzbøger, 2002; Cosyns et al., 2005; Cosyns & Hoffmann, 2005; Stroh et al., 2012). Species with small spherical seeds without appendages tend to pass faster through the intestinal tract and to retain viability better than larger seeds, seeds with shapes deviating from spherical and seeds with appendages (Bruun & Fritzbøger, 2002; Cosyns & Hoffmann, 2005; Stroh et al., 2012). However, germination experiments have indicated that many more plant species may be dispersed successful by endozoochory than only those with obvious morphological adaptations to this dispersal mechanism (Cosyns & Hoffmann, 2005). This is for example the case for a number of grass species with hooked or awned fruits, but also species with a mucilaginous sheet (often interpreted as an adaptation to epizoochorous dispersal) seem to be very well dispersed through animal

59 guts (Bruun & Fritzbøger, 2002). Of those species which increased their abundance by grazing in this study, have germinable seeds in of Poa trivialis (Cosyns & Hoffmann, 2005; Stroh et al., 2012), Ranunculus repens, Stellaria media (Sobey, 1981; Cosyns & Hoffmann, 2005) previously been found in horse dung, while germinable seeds of Viola riviniana (Buttenschøn & Buttenschøn, 2013), Viola reichenbachiana, Deschampsia cespitosa, Moehringia trinervia (Jaroszewicz et al., 2009) have also been found in dung form other large herbivores. Epizoochorous seed dispersal may also be an important aspect where fur characteristic is a central factor, and for example has the long undulated and open fur of Galloway cattle been proved to be much more suited for long-distance transport of seeds than the short, straight and closed fur of horses (Couvreur et al., 2005). But the surface of the seed and the height of the inflorescences are also determining characters for epizoochorous seed dispersal (Fischer et al., 1996; Bruun & Fritzbøger, 2002; Graae, 2002; Couvreur et al., 2005). Here may e.g. the hispids perianth segments of the fruits of Utrica dioica be well adapted to epizoochorous dispersal (Taylor, 2009), as well as the hairy seeds of Deschampsia cespitosa (Graae, 2002). However, seeds may also be transported by soil on hoofs, which is recorded as a factor in the dispersal of e.g. S. media (Sobey, 1981). However, if the species are highly preferred or very intolerant towards grazing then may zoochory be of minor importance as these species may never reach the phenological stage where they are able to set seeds. This may for example be the case for Calamagrostis epigejos, which are found negatively impacted by the horse grazing (Figure 22) albeit the seeds are reported to germinate from horse dung (Cosyns & Hoffmann, 2005; Stroh et al., 2012).

In resource-rich environments plants do often not avoid herbivory, but instead they develop tolerance traits to compensate for the harmful effects of herbivory (Skarpe & Hester, 2008). This may also be the case in situations where the plant defences are ineffective in preventing loss of biomass, e.g. if the herbivore is insensitive to the defences or if trampling is an important source of damage (Skarpe & Hester, 2008). Tolerance mechanisms are numerous and varied, but can be divided into intrinsic and extrinsic factors, meaning, those determined genetically or developmentally and those determined by external factors such as resource availability (Rosenthal & Kotanen, 1994). Intrinsic morphological factors promoting plant tolerance of herbivory include numerous protected meristems, wide distribution of leaves and buds, branching or tillering responses and seed numbers, viability, longevity and size (Rosenthal & Kotanen, 1994; Hester et al., 2006). Two important intrinsic physiological factors promoting plant herbivory tolerance are growth rate and growth plasticity (Hester et al., 2006). In general, slow growing plants should be less tolerant to herbivory because it takes longer for them to replace lost plant parts, particular in environments with few plant available resources where uptake is limited (Rosenthal & Kotanen, 1994). For example Vaccinium myrtillus has been found very slow to recover from herbivory (Crawley, 1997) and is thus very sensitive towards grazing. Indications of this were also found in this study (Figure 23). Growth plasticity relates to the ability of a plant to release dormant buds, modify nutrient uptake and allocation and increase photosynthetic activity (Hester et al., 2006). Amount of plant available nutrients are a key extrinsic factor. It seems logical that plants growing with more nutrients available should be more tolerant to large herbivore damage, through greater productivity and greater opportunity for nutrient uptake after tissue damage (Crawley, 1997; Hester et al., 2006). But this is not always the truth, perhaps because the interaction between plant available nutrient status and plant productivity can affect the probability of herbivore damage and the amount eaten (Hester et al., 2006). The plant vigour hypothesis (Hester et al., 2006) states that more productive plants should be more attractive to herbivores. On the other side, the stress hypothesis states that stressed plants actually may be more attractive to herbivores as they tend to have a

60 higher nitrogen content in the tissue and lower concentration of plant secondary metabolites (Hester et al., 2006). Plants with herbivory-tolerant traits are generally palatable to large herbivores, but in spite of the palatability they have been found to increase their productivity in resource-rich environments when moderately grazed or intense tramped (McNaughton, 1979; Putman, 1986; Hobbs, 1996; Skarpe & Hester, 2008; Chapin III et al., 2011). Such increase in productivity may be due to a number of factors. McNaughton (1979) lists nine possible mechanisms which may compensate for plant tissue loss from herbivory and may result in increased primary production following grazing or browsing. 1) Increased photosynthetic rate in residual tissue. 2) Reallocation of substrates from elsewhere in the plant. 3) Mechanical removal of older tissues functioning at less than a maximum photosynthetic level. 4) Consequent increased light intensities upon potentially more active underlying tissues. 5) Reduction in the rate of leaf senescence, thus prolonging the active photosynthetic period of residual tissue. 6) Hormonal redistributions promoting cell division and elongation and activation of remaining meristems, thus resulting in more rapid leaf growth and promotion of tillering. 7) Enhanced conservation of soil moisture by reduction of the transpiration surface and reduction of mesophyll resistance relative to stomatal resistance. 8) Nutrient recycling from urine and faeces. 9) Direct effects from growth promoting substances in ruminant salvia. Many of the McNaughton (1979) suggested mechanisms deals with increased photosynthetic activity in one or the other way. Likewise, Crawley (1997) suggested that after removal of tissue a reduction in NPP will occur when the LAI is low, but not necessarily when the LAI is high due to self-shading of leaves. At a high LAI the shaded leaves may have a negative NPP due to a higher respiration than GPP and then it is possible that herbivory can increase the rate of carbon-fixation by reducing self-shading (Crawley, 1997). Hence, the impact of a given level of defoliation depends critically on the specific tissue removed, where both age and tissue types are determinants for the impact (McNaughton, 1979; Huntly, 1991; Rosenthal & Kotanen, 1994; Crawley, 1997; Hester et al., 2006; Holtmeier, 2015). Rosenthal and Kotanen (1994) suggested that increased photosynthesis rates after defoliation probably are more common in grasses than other species, and Davy (1980) reported the dense tussocks of Deschampsia caespitosa and the substantial accumulation of litter results in a considerable degree of self-shading. Hence, grazing by the horses may partly be an explanation for the increased abundance of D. caespitosa (Figure 21 and Figure 22). On the other hand may the horses’ preference for grasses (Figure 1) cause a negative impact for preferred species. Bokdam et al. (2001) suggested that removal of large herbivores from woodlands would favour gaps dominated by e.g. Calamagrostis epigejos, Molina caerulea or tall Carex ssp. according to the site conditions. This suggestion appears very likely, at least are these species decreasing abundance by the presence of the horses in this study.

6.2. Forest structures Changes in forest structures come about as a result of changes in the abundance and composition of seedlings and saplings of trees, the size class that of course is most vulnerable to damage by large herbivores (Gill & Beardall, 2001; Gill, 2006). Selective browsing by horses during the tree establishment phase may dramatically alter the species composition of trees and shrubs, determine the botanical composition of the next woodland generation, which ultimately may have high impacts on the woodland biodiversity. It is well known that the overstory density has significant effects on the ground vegetation cover, composition and diversity (Mitchell & Kirby, 1990; Koorem et al., 2011). These effects occur through multiple interacting mechanisms such as changes in light availability, temperature, humidity conditions, the

61 effects of litter (Petersen & Vestergaard, 2006; Koorem et al., 2011) and the water balance, where tree species, age and structure of the stand will influence the balance of interception and evapotranspiration (Petersen & Vestergaard, 2006). Changes in the functional composition of the plant community due to selectivity may lead to changes in both the quantity and quality of litter inputs into soil (Hobbs, 1996; Bardgett, 2005; Skarpe & Hester, 2008) which may alter the soil development (Petersen & Vestergaard, 2006; Koorem et al., 2011). In woodland ecosystems where a significant amount of the litter produced originates from the trees, the tree species may often indirectly affect the ground vegetation composition through the soil formation (Petersen & Vestergaard, 2006). Litter from the most nutrient demanding species (Fraxinus, Ulmus, Tilia, Corylus) has a high content of nutrients, high pH and is easily decomposed. Beneath these species brown earth soil is always found. Beneath beech (Fagus sylvatica) and oak (Quercus sp.) it depends on the geological condition, whether and how quickly leaching and acidification occur, resulting in a potential podsolization. Litter from conifers decomposes slowly and podsolization of the soil will almost always occur – at least on the type of soils where these species are grown in Denmark (Petersen & Vestergaard, 2006). Repeatedly browsed saplings in a woodland may either get killed by the herbivores or suffer reduced competitive ability relative to other woody or herbaceous plants (Gill, 1992; Hester et al., 2006). Thus, impacts on competitive relations at the seedling and sapling stage are fundamentally important in determining the species composition of the mature tree layer, which, in turn, generally dominates the ecosystem processes in the forest for a considerable period of time (Gill, 2006; Skarpe & Hester, 2008).

6.2.1. Tree regeneration and browsing impact In general the direct and indirect effects of the horses on the regeneration of trees and shrubs are the same as for the ground vegetation. Some effects damage and delay the forest succession while others increase regeneration and thus tend to promote the succession. Compared to grasses and forbs, trees and shrubs often grow out of reach for large herbivores. Thus the size of the woody plant exposed to browsing is important in relation to its fate. Individuals already grown above the browse line (2 meters) are no longer exposed to the possibility of complete defoliation. Hence, it is mainly seedlings and saplings that are susceptible to browsing impacts and browsing of these size groups may have a serious impact on tree regeneration by impeding sapling growth or by filtering out browsing-sensitive species and thus changing the species composition (Gill & Beardall, 2001; Kuiters et al., 2006). Despite a tendency towards a higher richness of seedlings of woody species in the grazed sites in the oak stands, significant increases are only found in the old oak stand at Sandskredssøen and the oak stand in Lille Hessemose (Figure 27). However, the higher species richness in Lille Hessemose is probably not due to the grazing management but rather the higher richness of trees >2 meters in height found within the stand (Figure 31). In the old oak stand at Sandskredssøen it may by contrast very well be due to the grazing management. The positive effects on the natural regeneration of woody species induced by the horses result from grazing the herbaceous ground vegetation, which improve conditions for woody regeneration by reducing competition for light with vigorous herb and grass species, and by their hoof pressure which may break roots or rhizomes creating niches suitable for germination (Mitchell & Kirby, 1990; Gill & Beardall, 2001; Gill, 2006). The subsequently lowered amount of litter may also have indirect effects through a reduction in the population of small granivorous mammals (Putman, 1986; Mitchell & Kirby, 1990; Mayle, 1999) and thus a lowered seed predation and probability for ring barking of saplings by mice. These effects are, however, likely to be short-

62 lived if the seedlings are killed by browsing, trampling or shading from the trees already grown beyond browsing reach. And albeit the increased richness of seedlings in the old oak stand at Sandskredssøen there is indeed a tendency towards a lower richness of the larger saplings (Figure 30). However, it is difficult to disentangle whether this decrease is due to browsing or the higher overstory density (Figure 12). In contrast, it is clear that the overstory density is the most determining factor for the richness of both small (<0.5 m) and large (0.5 – 2.0 m) saplings in the beech stands (Table 2). Though the richness of seedlings are relatively high in the beech stand at Store Hessemose compared to all the other sites (Figure 28), only few species are surviving to the stage of sapling (Figure 29), and even none of them are growing to a size >0.5 m in height (Figure 30). It indicates that several species are able to germinate and maybe even survive the first year though the conditions are unfavourable, and that the factors that limit subsequent growth not necessarily are the same that limit germination. This is in general true for species with larger seeds which enable the seedlings to be more tolerant of low light, nutrient and water availability as they have bigger supply of energy (Vera, 2000; Thomas, 2014), and may in fact be the reason why oak seedlings as the only species are found in all the sites (Figure 32), in spite of oak’s high requirement for light (Vera, 2000; Friis Møller, 2010). However, it may also work the other way around, and for example saplings of Alnus spp. can be planted on remarkably dry soil but will never germinate (Thomas, 2014). Although the overstory density influence the richness of saplings significantly, it is obvious that also the horses affect the richness in the other beech stands and reduce the richness of woody species >0.5 m. Gill (2006) came to the conclusion that it is possible that browsing by large herbivores acts to increase species richness in openings but decreases in a more shaded understory. At least the latter is also confirmed by this study.

In this study it was only the combined positive and negative effects of the horses on regeneration which were measured – simply as the number of seedlings within the grazed and ungrazed stands. If these impacts are nearly equal in their magnitudes, no changes will be found. This may be the reason for the general lack of significant alterations in the species richness of woody species. To separate the browsing and trampling impact on the seedlings from the potential positive ones, an enclosure within the grazed stand could have been established in the autumn. The difference in the number of seedlings the next year between the enclosure and the surrounding grazed stand would then be equal to the difference between the negative and positive impacts on regeneration of woody species. However, other explanations are also possible. Though seedlings of Malus sylvestris (Buttenschøn & Buttenschøn, 1998), Crataegus sp. and Rosa sp. (Buttenschøn & Buttenschøn, 2001) have been found germinating from horse dung, reviews of endozoochorous seed dispersal among temperate tree species have revealed few examples (Gill & Beardall, 2001). This may be a result of the in general larger seeds of woody species, which thus are more likely to be crushed between the molars. Instead, birds and small mammals are likely to be the most important dispersal agents for larger seeds without adaptations for wind dispersal (Vera, 2000; Gill, 2006; Thomas et al., 2011; Packham et al., 2012). Hence, the number of species dispersed within the enclosure may not be affected by the presence of the horses, as the case is for the ground vegetation, but perhaps slightly disadvantage due to a possible lower population of small mammals (Putman, 1986; Mitchell & Kirby, 1990; Mayle, 1999). It is not uncommon among temperate woodland trees for germination to be temporarily delayed, but remarkably few of the dominant woodland trees in Europe accumulate persistent seed banks (Grime, 1979; Thomas, 2014). Thus it is not unlikely that the presence of horses actually is increasing the invasibility for woody species, as well as they do for the ground vegetation, and that net effects are positive but that the sites simply suffer from a lack of seeds dispersed into the site itself, and a pronounced lag time

63 between the increased invasibility and an increase in the species richness thus occur. For example, Euonymus europaeus is fairly shade tolerant, and even have low mortality in deep shade, though strongly inhibited (Thomas et al., 2011). Despite it is shade tolerant, it is often more abundant in open habitats. This is suggested more likely to be a matter of limited seed dispersal where birds are the main dispersers (Thomas et al., 2011).

However, it is important to make a point of the probably highly biased number of seedlings registered. Not only may seedlings be overlooked in the tall vegetation in some of the ungrazed sites, but also browsed seedlings are very easily missed when only a tiny stem are left present. Even though younger plants in general are less tolerant towards browsing than older more established individuals (Rosenthal & Kotanen, 1994), although they may be more defended (Huntly, 1991; Hester et al., 2006), the tolerance varies greatly between plant life forms, plant species and individuals of the same species (Rosenthal & Kotanen, 1994; Hester et al., 2006). Simulated grazing experiments have e.g. shown that oak seedlings are particularly resilient to browsing and only die when subjected to sustained defoliation (Mitchell & Kirby, 1990). It may be because of the hypogeal germination of oak where the cotyledons stays below ground, as very small seedlings are likely to be most vulnerable to damage if the cotyledons are lost before the first leaf forms (Gill, 2006). Thus browsed seedlings of oak may not necessarily be interchangeable with the death of the individual.

With certainty, the horses were found to browse on Fagus sylvatica, Alnus incana and Euonymus europaeus (Figure 35), but also browsed individuals of Carpinus betulus were more impacted within the enclosures (Figure 36). Though the horses may impact the other woody species as well, both negatively and positively, it is of high interest how they affect these apparently preferred species. However, for both E. europaeus and C. betulus the number of sites and the number of individuals in those sites in which they occurred were so few that any meaningful comparisons are difficult (Appendix 2). Individuals of C. betulus were primarily found in the old oak stand at Sandskredssøen with a higher abundance in the ungrazed site. Though the horses partly may cause the difference in the abundance of C. betulus, the primary reason is probably the presence of a large C. betulus tree in the ungrazed site (Appendix 3). Though Gill (2006) reported hornbeam to decrease under deer browsing is it often found in wood-pastures (Vera, 2000) and Samojlik and Kuijper (2013) reported hornbeam to be more tolerant to browsing than any other woody species found in Białowieża Forest, and found it to profit from browsing, however, mainly because browsing removes the competing species that are less resistant. The abundance of E. europaeus were higher in the ungrazed alder swamp in Store Hessemose compared to the grazed while the opposite was the case in the oak stand in Lille Hessemose. Thomas et al. (2011) consider E. europaeus as fairly resistant to browsing by mammals and reported it browsed by both roe and red deer and among goats it has been found particularly popular and totally browsed (Rupp, 2013). The plant is nevertheless toxic and poisoning of sheep, goats and horses has been reported in England (Cooper & Johnson, 1998). Why the horses apparently browse this toxic species selectively is more difficult to answer. But as it is not commonly found and nearly all the individuals are saplings <0.5 m in height (Appendix 2 and Appendix 3), the amount consumed will only constitute a minor proportion of the diet. The negative toxic effects may be ameliorated by ingestion of food with higher nutrient content (Bidlack et al., 1986), and the extent to which the plant secondary compounds are degraded may increase with the experience the gut flora has had with the compounds. For example, experienced deer can consume diets of up to 50% needles of Pseudotsuga menziesii, while deer rumen floras that had not previously experienced P. menziesii are severely inhibited by it (Freeland & Janzen,

64

1974). However, whether or not horses as hind-gut fermenters are able to degrade the toxic substances by a more experienced micro-flora may depend on where in the digestive system the uptake occurs. All the above listed explanations appear reasonable viewed in the light of the poisoning case from England where the horses were fed in stalls but one evening allowed out to graze a meadow, on the edge of which E. europaeus were growing and considerable amounts of the shoots were eaten (Cooper & Johnson, 1998).

The high abundance of saplings of Alnus incana found in Store Hessemose, albeit few seedlings (Figure 34), may be a result of asexual reproduction by root suckers which is common for Alnus species (Thomas, 2014). The significant higher abundance of small saplings in the grazed site and significant higher number of large sampling in the ungrazed site indicate that the horses are effectively impeding the growth rate. But although browsing almost invariably reduces height growth, some species may have dormant buds near the base of the main stem or on roots which are activated following damage to the main stem (Gill, 2006; Skarpe & Hester, 2008) which thus further may be a cause of the high density of small saplings in the grazed site. Apparently no or very few other studies have examined the browsing effect on A. incana induced by horses, but cattle have been reported to avoid it (Hauck & Popp, 2010).

Differences in the abundance of beech seedling induced by the horses are only found in the beech stand in Kollerup Enghave, as the higher abundance of seedlings in the grazed site in the old oak stand at Sandskredssøen (Figure 33) in all probability are due to the large beech trees within the site (Appendix 3). As young trees grow and the shoots become thicker and more lignified, digestibility declines and only the distal parts are likely to be browsed. Here the leading shoots and upper leaves are usually the most actively growing and nutritious part of young trees and shrubs, and are thus actively selected (Gill & Beardall, 2001). Packham et al. (2012) categorize beech to be more tolerant of browsing than other broadleaved tree species in its area of distribution. As a result browsing is more likely to affect only the growth rate or growth form rather than survival, and repeated browsing may thus keep the saplings within reach of browsing for years (Gill, 2006). Vera (2000) reports an example of a beech browsed by cattle. The beech was 1.4 m tall and had a diameter of 1 m, but proven to be 220-230 years old. The loss of growth depends on the severity or frequency of browsing and may vary considerably between tree species, depending on feeding selection and the ability of the woody species to recover from the damage (Gill & Beardall, 2001). In general, the more severe the tissue damage, the more severe the impact, but the limits of plant tolerance depend on a variety of factors including the availability of light, nutrients and other resources (McNaughton, 1979; Hester et al., 2006). Being stressed from depletion of light the saplings are vulnerable to other stress factors such as disturbance by browsing. Hence, removal of the same amount of plant material may have devastating consequences if the plant grows in an environment where the resource availability are close to its limits while it may have minor effects under other circumstances. The number of beech saplings found in this study varies widely in relation to location and light conditions (Table 3) where only few seedlings survive in the beech stand in Store Hessemose (Figure 33), which is the shadiest (Figure 12). This is even though beech is the most shade tolerant deciduous woodland tree species in Denmark (Friis Møller, 2010). Also the browsing impact induced by the horses has effects, especially in Kollerup Enghave where the high browsing impact (Figure 40, Figure 41Figure 42) in combination with the low light radiation (Figure 12) clearly is impeding the growth rate of the saplings (Figure 33 and Table 3), and even evidence occur for such severe browsing that the saplings are dying, as the total number of saplings and small trees (dbh <10 cm) appears to be much lower in the grazed part compared with the ungrazed (Appendix 3). Saplings may be expected to increase their growth rate when the ground vegetation is grazed

65 short as the competition for light decrease. However, there was no such increase in the growth rate observed in any sites. This observation may be because browsing by the horses has the same impact on growth rate as the taller ground vegetation has in the ungrazed site. Even though devastating consequences with certainty only can be observed in the beech stand in Kollerup Enghave, the horses may also have not yet observable consequences for the survival of small saplings in some of the other sites. Though beech saplings taller than 30 cm have been reported to survive several years of severe browsing under shade of >50% reduction in sunlight (Packham et al., 2012), the rate of survival may not be as high for those lower than 30 cm. The browsing impact may thus be of a degree so that the saplings are giving up after years browsing. Furthermore, the growth impeding effect may hide the fact that some of the very small saplings are not surviving. But though this is a possibility, there are evidences from both Epping Forest and New Forest, England, that beech has spread during historic times in the presence of intensive grazing and browsing of livestock (Kirby & Baker, 2013). It is thus most possible that neither the horses in Gribskov are able to fully stop the regeneration of beech except under very shady conditions as in the beech stand in Kollerup Enghave.

In addition, the impact on oak is also of high interest since oak is a widespread species in northwest Europe, and failure of natural regeneration within woodlands has been noticed for some time (Vera, 2000), which the lack of seedlings >0.5 m in height in this study clearly indicates. But neither grazing, overstory density, location, nor the interactions were able to significantly explain the abundance of oak seedlings and saplings in the oak stands (Table 3). However, some differences occurred. A higher abundance of oak seedlings were found in the ungrazed part of the beech stands in both Sandskredssøen and Store Hessemose (Figure 32). As no oak trees (>2 m) was found within the stand at Sandskredssøen, and the number and size of oak trees in Store Hessemose was equal for the grazed and ungrazed sites (Appendix 3), the propagule pressure must thus be the same for the pairwise comparable sites. Thus, the lower abundance of oak seedlings must be caused by the combined browsing and trampling effects of the horses. Although these effects cannot be separated here, the effect of trampling may be more pronounced in areas with a dense litter mat or high overstory density as these factors will increase the elongation of the oak seedlings while the stem weight decreases, which makes them more prone to mechanical damage (Facelli & Pickett, 1991; Ziegenhagen & Kausch, 1995). However, it appears that the horses do not have any effects on the regeneration of oak in the beech stands in the long-term, as the same number of individuals are surviving the first year, and neither in the grazed nor the ungrazed sites any saplings survive to a size >0.5 m. For the oak stands the only difference is found in the young stand at Sandskredssøen and as the only site the difference is consistent among size groups (Figure 32). However, Friis Møller (2010) states that oak in forest stands first sets seeds at the age of 30 years and Thomas (2014) even indicates an age of 40-60 years. The two stands have, however, not the same age; the grazed and ungrazed stand were respectively afforested 31 and 48 years prior to the fieldwork. With this fact together with the higher number of oak trees (>2 m) in the ungrazed site (Appendix 3), it cannot be ruled out that the difference in the number of seedlings and saplings are caused by differences in the propagule pressure. In general, it thus appears that the determinant for the survival of the oak seedlings is the amount of light radiation and that the horses cannot alter the regeneration in the long-term as long as the overstory densities remain unchanged. Also (Buttenschøn et al., 2009a) came to the same result in a cattle grazed oak forest.

It is often reported that more browse material may be consumed during winter months due to forage scarcity (Tolhurst & Oates, 2001), as it for example was observed for the horses in New Forest (Putman et

66 al., 1987). But no such indications are found in this study as neither the percentage of individuals of beech and oak browsed (Figure 37 and Figure 40) nor the browsing impacts (Figure 38, Figure 39, Figure 41 and Figure 42) were found significantly higher in the all year round grazed enclosure (Sandskredssøen). It may of course be because the horses prefer other forage during wintertime than beech and oak. For example the increased browsing impact during winter in New Forest was primarily due to the browsing on Ulex europaeus and Ilex aquifolium (Putman et al., 1987). However, it is also suggested that browsing during winter time on e.g. Betula sp., Sorbus sp. and Fraxinus sp. is less detrimental than summer browsing (Mayle, 1999). Moreover it has been demonstrated that in Betula sp., shoot size increases after winter damage compared to summer grazing (Hester et al., 2004). If this partly holds true for beech and oak, then the effect of browsing throughout the year together with possible altered preferences during wintertime may be difficult to observe during summer – especially after just a few years management. However, it would imply that saplings of beech and oak were only browsed during winter and due to the general high browsing pressure on particularly beech (Figure 35, Figure 41 and Figure 42) this does not seems likely. Nevertheless, it is also dangerous to assume that the rate of browsing on any species in a habitat will apply in another habitat. Preferences can depend on the vegetation composition and is thus likely to vary in relation to the abundance of more preferred species, so preferences cannot be expected to be consistent between sites (Gill, 1992; Hester et al., 2000). This is also elucidated in this study by the highly varying browsing impact between the sites, and even within sites (Sandskredssøen). The browsing impact in the all year round grazed enclosure may thus be higher than if the area was grazed only during summer time, but the effect may not be observable by comparison with other sides. This is however speculations and can be revealed only if the management is changed and then compared with the present.

Despite the vegetation composition, several other possibilities for the varying browsing impact on the same species between sites exist. It has been demonstrated that sapling density may influence browsing patterns (Hester et al., 2000), where for example the browsing impact on Quercus sp. and Sorbus aucuparia has been found decreasing with increasing density of saplings (Buttenschøn et al., 2009a). However, no such correlations appear to occur in this study, probably because the decreasing browsing impact observed in other studies is caused by increasing impenetrability and that such high sapling densities simply do not exist in any of the sites in this study. But it is also clear from studies of deer that the browsing impact can vary as much between sites or with vegetation than with density, or that the correlation with density is more pronounced in some habitats or tree species than others (Gill, 1992). The strength of the association between browsing impact and density therefore depends both on density itself as well as on habitat or tree related factors. There is also plenty of evidence that elevated levels of nitrogen in soil will increase a tree's susceptibility to browsing (Gill, 1992; Hester et al., 2000; Miller et al., 2007). However, this does not seem to be the case in this study where both the percentages of individuals of oak browsed and the browsing impact appear to be higher on the poor soil in the young oak stand than the more nutritious soil in the old oak stand at Sandskredssøen (Figure 37, Figure 38 and Figure 39). Thus, the stress hypothesis that states that stressed plants are more attractive to herbivores as they tend to have a higher nitrogen content in the tissue and lower concentration of plant secondary metabolites (Hester et al., 2006), rather appears to be true. This may also explain the high browsing impact in the beech stand in Lille Hessemose where the beech saplings may be stressed due to the high overstory density. On the other hand, the lack of alternative forage in the beech stand in Kollerup Enghave may also bring about the higher browsing impact, as shown by Miller et al. (2007). Furthermore, many studies have reported changes in palatability or nutritional

67 quality following browsing or other forms of damage, which in turn influence the likelihood for further browsing. There are species which become more palatable after browsing as a result of mobilization of nutrients and changes in shoot morphology or possibly plant architecture (Huntly, 1991; Gill, 2006). For many species, trees that have been damaged once have been found to be more likely to be damaged again (Gill, 2006). In contrast, some thorny species (e.g. Rubus fruticosus and Ilex aquifolium) have been found to grow shoots with more thorns after being browsed (Gill, 2006), indicating an ability to increase defence against further browsing. Changes in the palatability and nutrient level in the leaves may also explain why the larger saplings are browsed more than small (Figure 35 and Figure 36). In a cattle grazed woodland the same pattern was found for most woody species, whereas Cytosus scoparius with nitrogen fixation and root suckers from Populus tremula was exposed to the same browsing pressure regardless of sapling size (Buttenschøn & Buttenschøn, 2013).

6.2.2. Bark stripping In addition to the impact of foliage browsing, herbivores may severely damage trees and shrubs by bark stripping (Mitchell & Kirby, 1990; Gill, 1992; Gill, 2006; Hester et al., 2006; Kuiters et al., 2006). All large herbivores may strip the bark from trees and shrubs, though some herbivore species will be more prone to this (Suchant et al., 1998). The herbivores most commonly reported to strip bark are the large browsers or intermediate feeders (European bison (Bison bonasus), moose (Alces alces), red deer (Cervus elaphus), sika deer (Cervus nippon) and fallow deer (Dama dama)) (Gill, 2006). Although the frequency of bark stripping in this study was found almost negligible and the registered occurrences not necessarily are caused by the present of the horses, bark stripping by horses has been reported in other studies (Keenan, 1986; Kuiters et al., 2006). The impact of bark stripping can be severe because it affects not only saplings but also canopy trees. Debarking may thus potentially impair growth, and could result in partial or total crown die-back (Kuiters et al., 2006). If bark is removed from the entire circumference, phloem translocation is interrupted and death is normally inevitable, and where some tree species are stripped preferentially, the species composition may be altered. However, most stripping by ungulates removes bark from only a proportion of the stem circumference and trees usually survive and continue to grow (Gill, 2006). Sudden out-breaks of bark stripping has been reported for horses (Keenan, 1986; Kuiters et al., 2006) and other ungulates (Gill, 1992) and is a well-known and quite common phenomenon (Gill, 2006). Here learning might have an effect on bark stripping behaviour. Once some individuals have started to strip bark the habit is likely to be repeated and copied by others (Gill, 1992; Kuiters et al., 2006). Hence, a possible future out-break of bark stripping by the horses in Gribskov cannot be rejected even though it is not reported yet. Several factors contribute to the susceptibility for bark stripping (Gill, 1992; Kuiters et al., 2006). Specific morphological characteristics such as bark thickness, bark roughness, stem branchiness and the ease of bark removal have all been suggested or reported to be important determinants of damage (Mitchell & Kirby, 1990; Gill, 1992; Gill, 2006; Kuiters et al., 2006). But since many of these characteristics develop together as a stand ages it is difficult to isolate which of them are most important (Gill, 1992). Kuiters et al. (2006) found Icelandic horses to strip bark from beech where trees with a smooth bark and a dbh <40 cm were significantly more damaged compared to individuals with a rough bark structure and larger diameter. Trees species such as Quercus robur, Betula pendula and Pinus sylvestris showed no damage at all (Kuiters et al., 2006). In addition Kuiters et al. (2006) observed that a difference in the susceptibility for bark stripping of beech between stand and more solitary trees occurred. This is probably related to density-related differences in

68 bark thickness, since bark on suppressed trees is thinner than bark on dominant and more free-growing trees (Gill, 1992). A possible out-break of bark stripping of beech trees by the horses in Gribskov is, due to stand structure and tree size, most likely to occur in the beech stand at Sandskredssøen or in Store Hessemose. In the stand at Sandskredssøen the dbh for all individuals are <40 cm and this also applies to most individuals in the stand in Store Hessemose (Appendix 2). However, with a significantly higher overstory density in Store Hessemose (Figure 12) the thickness of the bark may be thinner here.

There does not appear to be a unifying explanation for the variability in levels of bark stripping damage nor the sudden outbreaks (Gill, 2006). It has often been suggested that nutrient shortages might be a main reason for bark stripping behaviour of ungulates (Gill, 1992; Mayle, 1999), but Kuiters et al. (2006) did not find the mineral composition of the bark tissue to explain the horses specific preference for bark of beech. Another suggestion made for ruminants is that that bark stripping behaviour sometimes coincides with a change to a more nutritious diet in spring or after supplementary feeding, giving the herbivore the need for a more fibrous-rich diet to balance rumen pH (Gill, 1992). With only very protein-rich and easily digestible forage, the digestion of roughage feeders will collapse after a while due to over-acidification (Holtmeier, 2015). Keenan (1986) suggested a similar explanation for a sudden outbreak of bark stripping by horses when allowed to graze an irrigated pasture with a low fibre and high nitrogen content. Bark stripping appears to occur mostly in winter in temperate regions, however, there is considerable variation in this pattern and it may occur at any time of year (Gill, 2006). There are even examples where it is more common in the summer and e.g. beech are usually only stripped in summer because the bark is more easily removed at this time of year (Gill, 1992).

6.2.3. Browsing and grazing induced succession alterations The scale of this study is indeed important, not only spatially, as seen in the case of the species richness of the ground vegetation in the oak stand in Lille Hessemose but also the temporally. What appears to be stable in the timescale of a few years may be highly unstable over a longer period. Similarly, a change over a short period may simply be a fluctuation over a longer time scale. Thus, the impacts on the regeneration found in this study have to be interpreted given the short time of management. The evidence found for horses causing damage and affecting survival rates in the beech stand in Kollerup Enghave may be turned upside down if gaps in the overstory are suddenly formed and the increasing amount of light is altering the competition. On the other hand, the apparently only delaying effect on many saplings may be devastating if a sudden outbreak of bark stripping occurs. It is thus difficult to predict how the landscape will develop in the future with the present management, not at least because of the general shortage of studies with direct evidence of the long-term consequences of these effects (Gill, 2006). There is however examples of landscapes from northwest Europe where large herbivores inhibit shrub and tree cover expansion, e.g. New Forest and Oostvaardersplassen (Schippers et al., 2014), but there are also examples where they are unable to maintain patches of open landscape (Samojlik & Kuijper, 2013). Another way to try to predict the future effects is to try to understand the role of large herbivores in the past. How forest landscape looked, and more specifically, the level of openness of the primeval forest of northwest Europe, have been questioned and discussed in recent decades by several authors (e.g. Bradshaw & Mitchell, 1999; Vera, 2000; Bradshaw, 2001; Birks, 2005; Mitchell, 2005; Samojlik & Kuijper, 2013; Vera, 2013).

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Using palaeoecological evidence such as fossil pollen records and plant remains from peat and lake deposits, the high forest hypothesis was proposed. It states that lowland Europe was originally covered by dense, closed-canopy primeval forest and that it was the human impact in Neolithic and onwards that opened up the forest (Birks, 2005; Mitchell, 2005). Vera (2000) however states that this hypothesis, with the climax vegetation being a closed-canopy forest, was constructed without taking any impact of large herbivores into account. Species such as aurochs (Bos primigenius) and tarpans (Equus ferus ferus) were absent because they were extinct. When the natural vegetation was theoretically reconstructed, it was well known that grazing livestock such as cattle (Bos taurus), horses (Equus ferus caballus) and sheep (Ovis aries) prevented regeneration of trees in the forest. Consequently, the aurochs were defined as a forest species that should not have any influence on the forest regeneration – otherwise, the natural conditions would not have been a closed-canopy high forest (Vera, 2013). As cattle create grassland in woodland, they were not found to act as an ecological functional equivalent of the aurochs, because the aurochs, it was believed, lived in a closed-canopy forest. Horses were excluded from the reconstruction of the natural vegetation because the only extant species all live in open grassy biotopes (Vera, 2000). Still existing wild herbivore species such as red deer, roe deer, elk and European bison can, at certain densities, have an impact on the development of forest (Vera, 2013). This well-known potential threat posed by large herbivores in high densities for forest regeneration lead to the conclusion that their natural densities must have been very low. The high forest hypothesis was widely accepted by forest ecologists but raised for discussion after the publication of Vera’s competing wood-pasture hypothesis based on a non-linear cyclical system (Vera, 2000). It is impossible to reconstruct the population densities of large herbivores in prehistoric times (Bradshaw & Mitchell, 1999) but whereas the traditional view implies that forest structure dictated the herbivore carrying capacity, the Vera (2000) hypothesis dictates that herbivore density controls forest structure. In the wood-pasture hypothesis the succession is cyclic and spatially asynchronized and hence, it is a dynamic shifting mosaic of grassland, shrub thickets and woodland patches, where the closed-canopy forest only exists in groves and are not self-perpetuating (Vera, 2000).

Vera (2000) proposed that the rich abundance of fossil oak (Quercus petraea and Q. robur) and Corylus avellana pollen in peat and lake deposits across Europe indicate that the primeval forest must have been relatively open to enable oak and hazel to regenerate. Oaks are light-demanding trees that can germinate in the shade but need open conditions for long-term survival and growth of the seedlings (Ziegenhagen & Kausch, 1995), hence, they are not able to regenerate under a closed canopy (Vera, 2000). Evidences for this fact are also highly supported by this study. The openness was, according to Vera (2000), a result of large herbivores grazing and browsing in woodland areas. After a gap in the canopy is formed large herbivores will prevent regeneration of trees. In this way the canopy opens up – a process further facilitated by fungi and drought that will kill more and more of the senile trees (Vera, 2013). Seeds of ground vegetation are brought in and dispersed by zoochory and as the grove become more open as more trees die, a grazed lawn develops. In this way, woodland change from the centre with the oldest trees, gradually into grassland. Seedlings and saplings of palatable tree species growing in grassland are only protected against grazing when they grow in close vicinity to thorny shrubs such as Prunus spinosa, Crataegus sp. or herbaceous species protected by chemical defences such as low digestibility or toxicity (associational resistance). These species are avoided by the large herbivores and hereby they defend the palatable tree seedlings and saplings (Bakker et al., 2004). The regeneration of trees fails beyond the nurse species because they are either eaten or trampled by the large herbivores (Vera, 2000; Bakker et al., 2004).

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In mature thorny shrubs, the annual shoots lack spines, and they are browsed by large herbivores. This browsing results in a divaricate branching, which increases the impenetrability of the herbivores, enhancing the protection of the undefended palatable tree species (Bakker et al., 2004). In case the protective shrubs do not expand clonally but mainly from seeds (e.g. Crataegus sp.), a single or few trees will grow up, but if the shrubs expand clonally by root suckers (e.g. Prunus spinosa) seedlings will settle again and again at the edge of the thicket (Bakker et al., 2004). The only settle here, because within the thicket the light density is too low (Vera, 2000). When the trees grow above the shrubs, the crowns join together, resulting in a grove. Over time a gap can be created within the forest, and the process can restart. In the whole process, large indigenous herbivores are steering the succession, especially the true grazers, cattle and horses (Vera, 2000).

There is still much debate on how large the impact of large herbivores really was on the pre-Neolithic landscape. Did they create half-open parklands or were their densities and impact too low to maintain open areas and was it instead closed forest that dominated (e.g. Birks, 2005; Mitchell, 2005)? The dominant role of large herbivores proposed by Vera (2000) does not seem to be supported by more recent studies (Kirby & Baker, 2013; Rackham, 2013). However, Holtmeier (2015) argues that the common buzzard (Buteo buteo) relying on open hunting grounds, some thermophilic reptile species needing sunny and warm places and some songbirds and butterfly species would never have been able to exist under a closed forest canopy. Thus, the hypothesis that only the alpine zone, coastal areas, flood plains, a few bogs and dunes as well as avalanche pathways and terrain affected by rockfall were originally devoid of forest can hardly be supported (Holtmeier, 2015). Also finding of pollen from long-lived plants like Succisa pratensis, that do not flower in shade, indicates that the pre-Neolithic landscape must have had at least modest permanent open areas (Rackham, 2013). On the other hand it can also be said that it is unrealistic and of limited value in relation to nature management to try to restore the forest ecosystem to a pre-farming era, which for Gribskov is more than 6,000 years ago (Rune, 2009; Overballe-Petersen et al., 2013a). Not least because not enough is known about the structure of the wildwood to form a basis for restoration (Rackham, 2013) but also because on a timescale of millennia, no static baseline conditions of natural forest have ever existed. Even without human intervention, the landscape and species composition would have continued to change in response to natural disturbances, fluctuations in climate, the maturation of soil, development of blanket bogs and the continued spread of species from post-glacial refuges (Kirby & Baker, 2013; Overballe- Petersen et al., 2014). Moreover, many changes like the loss of time and the globalisation of plant diseases, are beyond human ingenuity to reserve (Rackham, 2013).

Samojlik and Kuijper (2013) suggest that both views are probably correct, and that the pre-Neolithic landscape in Europe most likely consisted of large stretches of closed high forest dominated by browsing herbivores (Birks, 2005; Mitchell, 2005) interspersed by open or part-open landscapes dominated by grazing herbivores (Vera, 2000). Hence, it is not about choosing which view is right; it is all a matter of scale. Nevertheless, it is expected that the horses probably by time will allow a more dynamic development where the distinct boundaries created during the last 200 years management will be turned into transition zones with natural gradients between the forest and meadows or pastures. This natural dynamic will create varying degree of light and humidity in many possible combinations and result in more heterogeneous habitats. Furthermore, the browsing, though is not reverses the succession, may change the structure of the trees. Oak trees with larger crowns may develop. The impeded beech saplings may form bushes instead, where one or several shoots by time can grow out of reach of the horses. If several shoots grow at

71 the same time it will result in a beech tree with several trunks (Vera, 2000) beneficial for the biodiversity. And if the trees are not outcompeted by undergrowth veteran trees may develop creating important habitats for many organisms.

6.2.4. Other forest structures and the biodiversity Many organisms directly or indirectly modulate the availability of resources to other species by causing physical state changes in biotic and abiotic materials (Jones et al., 1994) and grazing herbivores also influence both invertebrates and vertebrates through their effects on vegetation structures (Mitchell & Kirby, 1990; Holtmeier, 2015). Both greater abundances of spiders and higher diversity of birds have been found in grazed compared to ungrazed woodlands (Mayle, 1999). Also the dung of the horses is an important invertebrate habitat, with large numbers of coleoptera, diptera, earthworms, nematodes, mites and collembola (Mayle, 1999). By increasing the species richness of the ground vegetation and by the preference for grasses horses may increase the richness and abundance of flowering herbs and shrubs beneficial for many butterflies, bees and other insects relying on nectar and pollen (Mayle, 1999; Potts et al., 2003). These pollinators are closely related with the flora community and as several species are oligolectic and thus only collect pollen from specific taxa of plants, a high species richness of insects is conditional on a high richness of flowering herbs and shrubs (Westrich, 1990; Roulston & Goodell, 2011). These pollinators may further modulate the supply of resources for seed predators (Jones et al., 1994). Perhaps only Ilex aquifolium and Hedera helix are the only fruit-bearing and insect pollinated woody species, capable of producing flowers and fruit in dense, dark forests (Green, 2013). Hence, if the horses are able to create more open landscapes with transition zones, they may increase the richness of insect pollinated woody species. Though Prunus avium can germinate in shadow and Crataegus sp., Lonicera periclymenum, Prunus spinosa, Prunus cerasifera, Rosa sp., and Sorbus aucuparia can germinate in semi- shade (Buttenschøn & Buttenschøn, 2013), they often require more light for establishment (Vera, 2000). The lack of the species in this study in sites where more light can penetrate the canopy is probably due to the litter. If they do germinate as it obviously happens for some individuals, the seedlings and smaller saplings often die because they are nibbled by rodents (Vera, 2000). Though a consistent increased richness of genera of woody species pollinated by insects was only found in the oak stand in Lille Hessemose (Figure 44), the richness may also increase in some of the other sites by time. Most of these species are fairly resistant to browsing of large herbivores as they form shoots from stools or have new growth from roots (Vera, 2000) and when reached a size >0.5 m, browsing will result in forming of a dense brush of twigs, in some species assisted by thorns, which will make them less susceptible to damage (Buttenschøn & Buttenschøn, 2013). Though they may establish they do not necessarily flower unless the amount of light is sufficient. For example does Crataegus sp. not flower in shade (Alexander, 2013). Hence, these species’ presence are not likely to benefit the insect pollinators unless sufficient amounts of light are available too.

Creating more open conditions, open-grown trees may develop which otherwise are lacking in closed- canopy forest. These trees are themselves important for many plant and animal species in combination with the open surroundings, as open-grown, light demanding trees and shrubs have a far greater and diverse assemblage of biodiversity both in number of species and mass than shade-tolerant tree species growing in dense, dark forests (Green, 2013). This applies especially to oak to which more species are connected than any other indigenous European tree (Vera, 2013). Furthermore, although shade-tolerant

72 trees are able to regenerate under high overstory densities and are better able to compete for light, they may only produce lateral branch growth under more open conditions, which may be important habitats for invertebrates, fungi and epiphytic lichens and bryophytes, particularly if veteran trees develop and deadwood occur.

Giving the short time of management the only way the horses could possibly have had any impact on the number of trees with decaying parts, hollowness and the amount of deadwood would be by bark stripping where a proportion of the wounds would become infected with microorganisms leading to decay. But as bark stripping not is reported in this study, it is of no surprise that neither any alteration of these structural elements are found (Figure 45, Figure 46 and Figure 47). Nevertheless, these elements are an important energy source and a condition for many different organisms and a high biodiversity in the woodlands. The registration of hollowness and larger parts with decay are indicators for a number of micro-habitats which may be much more difficult to register by field observations. Some of the damages leading to decay may also alter the growth processes, the structure and chemistry of the bark, beneficial for a number of epiphytes and decomposers (Nygaard et al., 2013). Nygaard et al. (2013) indicate that an adequate number of trees with either decay or hollowness is one tree in at least 75% of the 15 m circles and 3 trees in at least 50% of the circles in oak forests on nutrient-poor soil (Code 9190 in the Habitat Directive: Council Directive 92/43/EC), while respectively 2 and 5 trees in oak forests on mull soil (Code 9160 in the Habitat Directive: Council Directive 92/43/EC), beech forests (Code 9110 in the Habitat Directive: Council Directive 92/43/EC) and alder swamps (Code 91E0 in the Habitat Directive: Council Directive 92/43/EC). But since no discriminations in this study were made between trees with decaying parts and trees with hollowness it is not possible to count the number of trees due to the risk of counting the same tree twice. However, as the possibility for a tree with hollowness also having decaying parts are likely to be higher than the other way around, it appears most fair to use these numbers. In that case only the alder swamps in Store Hessemose have sufficient amounts of these structural elements (Figure 46 and Figure 47). Nevertheless, with only a total of four trees in two of the quadrats in the old oak stand at Sandskredssøen (Appendix 3) it appears unrealistic to reach the amount of trees set by Nygaard et al. (2013) in this site. This may also be the case in the grazed stands in the future depending on how open the horses are able to keep the woodland.

Deadwood is used both as a food source, habitat and shelter for winter chill and summer drought of many wood-living species (Harmon et al., 1986; De Jong et al., 2004). Old trees and deadwood are habitats for close to one third of all Danish woodland living fungi and animals (Bruun et al., 2009; Nygaard et al., 2013) and approximately half of the red-listed woodland species in Denmark are related to deadwood in one form or another (Bruun & Heilmann-Clausen, 2012). Hence, removal of old trees and deadwood is the most serious threat to these species (Bruun et al., 2009). Also many vertebrates are dependent on deadwood, including amphibians, reptiles, birds and mammals (Harmon et al., 1986), but the most pronounced wood- living vertebrates are probably woodpeckers. Their nesting-holes are crucial for other hole-nesting birds. Altogether, in Sweden about 40 species of birds and mammals are entirely dependent on deadwood (De Jong et al., 2004). Hence, deadwood is one of the most important conditions for biodiversity in the woodlands. The big question is how much deadwood is needed in the forest. The species richness and abundance of species increase with the amount of deadwood (Harmon et al., 1986; Bruun et al., 2009). But it is not just the volume of deadwood that is important. The quality of the deadwood is often a direct determinant for whether certain species occurs or not (Jonsell et al., 1998; De Jong et al., 2004). The most obvious factors for the quality of deadwood are the tree species, dimensions (diameter and length), degree

73 of decay and the environmental microclimate including degree of sun exposure (Harmon et al., 1986; Jonsell et al., 1998; De Jong et al., 2004). In a study of all red-listed saproxylic invertebrates in Sweden were e.g. 202 species found living on deadwood from Quercus, 140 species on Fagus sylvatica, and 114 species on Betula sp. Of the species in the study the highest number of specialist (29%) were living only at wood of Quercus sp., but almost all tree genera have some monophagous species (Jonsell et al., 1998). Even though many species can occur within a high spectrum of deadwood sizes and deadwood with small dimensions can be an important substrate for wood-living species, especially in managed forest where deadwood with a larger diameter is a rarity, deadwood with smaller dimensions cannot replace the larger as many endangered and rare species exclusively or mainly lives on deadwood of these sizes (Dahlberg & Stokland, 2004). In this way deadwood with a larger dimension can sustain a higher diversity. The explanation for this lies in a more heterogeneous habitat that takes longer to break down and, hence, provides a stable microclimate which favours some species and finally the presence of certain fungi that several insects depend on (Harmon et al., 1986; Jonsell et al., 1998; De Jong et al., 2004). Dahlberg and Stokland (2004) found, that in Sweden, which is comparable to Denmark, nearly 50 % of all species, within a variety of different taxa living on deadwood, was primarily present on trunks with a diameter >20 cm. In general, with fungi as the only exception most species occurred at deadwood thicker than 20 cm – for wood-living insects it was the majority (80 %), which preferred deadwood at these sizes. Thus it seems fair that only deadwood above this size was included in this study. The deadwood’s position (standing or lying) as well as the degree of decay has an influence on the diversity. A trunk with good soil contact gets a different humidity and hence a different decomposition process than a trunk with less soil contact or standing deadwood. This result in different species compositions (De Jong et al., 2004), where standing deadwood gives rises to the highest diversity of coleopteran, lichens and vertebrates while lying deadwood in general is preferred by dipteran, fungi and bryophytes (Dahlberg & Stokland, 2004). Gradually during the decomposition, from the death of a tree until the wood is completely decayed, the energy and nutritional quality becomes worse. This affects which species are present and a conspicuous succession of invertebrate species, bryophytes and polypore-fungi occurs (Jonsell et al., 1998). In general the largest number of invertebrate and fungi species as well as species belonging to other taxa is occurring intermediately in the decomposition (Jonsell et al., 1998; Dahlberg & Stokland, 2004). With all these specific qualities of deadwood in mind, and knowing that specific qualities must be in sufficient amounts to sustain vigorous populations of specialists, De Jong et al. (2004) draw the conclusion based on other studies that probably around 20 m3 deadwood per ha is needed to secure the more specialized species. But to cope with all species, core areas with a significantly higher amount of deadwood are needed. Nygaard et al. (2013) reports of a significant loss of diversity by values below 50 – 70 m3 deadwood per ha in productive forest types and hence concludes that the minimum adequate volume of deadwood are 20 m3 ha-1 in at least 75 % of the 15 m circles and 50 m3 ha-1 in at least 50 % of the circles in oak forests on nutrient-rich soil (Code 9160 in the Habitat Directive: Council Directive 92/43/EC), while respectively 15 m3 ha-1 and 45 m3 ha-1 in oak forests on nutrient-poor soil (Code 9160 in the Habitat Directive: Council Directive 92/43/EC) and alder swamps (Code 91E0 in the Habitat Directive: Council Directive 92/43/EC) and finally respectively 10 m3 ha-1 and 30 m3 ha-1 beech forests (Code 9110 in the Habitat Directive: Council Directive 92/43/EC). The only stands which can match the limits set by De Jong et al. (2004) are the alder swamps in Store Hessemose and the old oak strand at Sandskredssøen, while only the alder swamps in Store Hessemose can match the amounts set by Nygaard et al. (2013) (Figure 45). But even though a higher volume of deadwood will increase the probability of a higher heterogeneity in the quality of deadwood, there seems to be a lack of standing deadwood in the

74 grazed alder swamp in Store Hessemose where only 5% of the total volume are standing and only represented by the intermediate decay states (Figure 45 and appendix 4). This pattern appears to be the general, but also with pronounced variations in the total volume between sites and within sites. Also Nygaard et al. (2013) found a large variation countrywide where 58 % of the 2118 samples were without any deadwood. However, Harmon et al. (1986) argue that major variations in the spatial distribution of deadwood probably will occur naturally due to the fact that the causes of mortality, such as windthrow, insect and diseases exhibit highly aggregated spatial patterns, and furthermore that the intermediate decay classes will tend to compose the largest fraction of deadwood biomass, while the most and least decayed compromise the smallest fraction (Harmon et al., 1986). However, this may depend on the residence times of the classes.

If the browsing by the horses is able to delay and stop at least a part of the regeneration, veteran trees will develop over time without being outcompeted by regenerating undergrowth. While a dead trunk will decompose in a relatively short time hollowness and decaying parts in old living trees will act as habitats for longer time. Old trees with hollowness and other damages may be habitat for more than 100 species of fungi, lichens, mosses and insects throughout many decades (Bruun et al., 2009). A number of insects only breeds in wounds on living trees and others only live in hollow trees (Jonsell et al., 1998). Furthermore, these trees will also contain many other microhabitats that are rare and required by saproxylic species, e.g. thick branches, coarse bark and dead parts of the stem (Jonsell et al., 1998). Also the more open structure of the woodland will favour a number of wood-living species. Deadwood can often behave quite differently when produced from open-grown trees in comparison the dense forest tree. When constantly exposed to sunlight and wind, the conditions can lead to desiccation with virtually no or very slow decay. Consequently, there is a whole spectrum of decay rates depending on the degree of moisture conditions found in the woodland (Vera, 2013). A high proportion of saproxylic invertebrates are dependent on host trees having sufficient space to develop a full grown crown with natural death and decay of the various woody tissue produced (Alexander, 2013). While very few of the red-listed wood-living insects in Sweden require shaded conditions 59 % are classified as preferring sun-exposed sites or indifferent to light conditions (Jonsell et al., 1998). Thus, woodland grazing may highly increase the diversity of deadwood qualities where a number of the habitats are threatened today.

Since the grazing management itself have not had any major impacts on the forest structures yet, no distinct picture of the effects on epiphytic bryophytes and lichens has emerged (Figure 48). Many species of epiphytic bryophytes and lichens thrives optimally by a combination of light and constantly high humidity. Hence, these microclimate factors and the structural elements modifying these factors (overstory density, shrub layer and vertical structure of the canopy) are very influential for epiphytes (Király et al., 2013; Odor et al., 2013). This may also be the reason for the apparently positive correlation between overstory density and number of trunks with epiphytes in the young oak stand at Sandskredssøen and the oak stand in Lille Hessemose, the only two sites where significantly differences occurred. However, the result for the stand at Sandsskredssøen may also simply be a matter of the age of the stands where epiphytes have had more time for establishment in the ungrazed site in combination with the higher density of trees (Appendix 3) and, hence, more substrate for growth.

Even though it in many ways would be more meaningful to compare the richness and composition of epiphytic bryophytes and lichens, not at least because the richness of lichens has been in decline and at

75 least 20 epiphytic species have disappeared from Denmark within the last 150 years (Johannsen et al., 2013), it was only the number of trunks that contained larger areas (>100 cm2) with epiphytic bryophytes and/or lichens that were measured. It may though be argued that a higher area (higher number of trunks with bryophytes and/or lichens) may increase the possibility for a higher richness due to the species-area relationship. Neither were any discrimination between bryophytes and lichens made, though the two groups have different preferences for the amount of light and humidity (Odor et al., 2013). However, it is likely that the impacts of grazing in the long-term will have quite different effects for the two groups. Jönsson et al. (2011) found the species richness of lichens to be higher in grazed wooded meadows compared to unmanaged wooded meadows. The semi-open overstory and transitions zones between the forest and open habitats increases the amount of heterogeneity in the light conditions and humidity which is considered beneficial for lichens particularly for forest specialists (Jönsson et al., 2011; Johannsen et al., 2013; Király et al., 2013). In return, the more open conditions may have devastating effects for the bryophytes as they are in general less light demanding but more sensitive to desiccation than lichens (Király et al., 2013). Nevertheless, there are several factors that affect the two groups in the same way and which may be altered in the long-term due to the grazing management. Even-aged forests with one-layered canopy (Király et al., 2013) and straight trunks of trees in good growth (Johannsen et al., 2013) are for example rarely a good habitat for the rare species of epiphytes. Tree species is also an important factor for the diversity of epiphytes and in general the diversity of both bryophytes and lichens is greatly improved by increasing diversity of woody species (Király et al., 2013). Many epiphytic species have preferences for host trees, driven by the difference in bark texture, chemistry, water and nutrient supply of different tree species (Király et al., 2013). Oak trees have been proven to host a higher diversity of both epiphytes groups compared to beech (Király et al., 2013). This is due to the more mesotrophic, wrinkle-rich bark of oaks which provides wind-proof and moist microhabitats whereas the smoother bark of beech is more exposed to hardships of the environment (e.g. stemflows, sun exposure and desiccating winds) (Chapin III et al., 2011; Király et al., 2013). This difference may explain the very low abundance of epiphytes in the beech stands found in this study (Figure 48). Tree size and age are also important determinants of epiphyte diversity. Larger and older trees maintain more diverse assemblages than younger ones (Jönsson et al., 2011; Király et al., 2013; Odor et al., 2013). This is both because big trees provide larger colonization surface and old trees ensure longer time for the establishment and growth of local populations, but also because they provide higher microhabitat diversity. Large trees have cracked, decayed bark and deep bark fissures giving rise to a variety of microhabitats (Király et al., 2013).

6.3. Management The ground vegetation in a woodland ecosystems is particular heterogeneous; not only is it affected by edaphic factors, but there is also the influence of the tree canopy cover which varies both spatially and temporally (Mitchell & Kirby, 1990). From the results it appears that some areas within the enclosures more easily are overgrazed than others due to these variations, and that it may be a challenge to secure an optimum grazing intensity on all sites within the same enclosure when highly different carrying capacities occur. Hence, objective criteria for management of woodlands cannot be based solely in terms of animal numbers per hectare but must take into account the carrying capacity. However, the horses’ selection for sites within the enclosure is not uniform. In New Forest, horses spend about half of their time on the grassland communities although these accounted for less than 5% of the total area of the woodland

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(Mitchell & Kirby, 1990), and the same pattern was shown by Lamoot et al. (2005). This segregation in the use of the habitat may indeed be the key to secure that overgrazing does not occur in the forested areas. It is thus important that the forested areas within the enclosure not are too small compared to the size of the enclosure itself and the number of horses. Given that the grazing intensity in the open area in Lille Hessemose is adequate, the fence can with advantage be moved to include a larger part of the oak stand as the current site appears to be close to overgrazed due to the vegetation-less patches (Figure 51) and the lowered vegetation density (Figure 8). However, this method is not applicable to all habitats. An enlargement of the grazed area in the beech stands will not turn the devastating grazing effects to positive impacts. A higher biodiversity in the beech stands can only be obtained by compromise the silviculture of monocultures of beech trees with dense closed canopies. The carrying capacity can only be increased by opening up the overstory allowing more light to reach the forest floor. Thus, the ecosystem services of timber production and biodiversity enhancement by livestock grazing are incompatible in this woodland habitat type. And even though the grazing management by time may open up the beech stands it will take long time as the uniform forest formed during the last decades will continue to dominate the landscape for some decades to come. Hence, artificial reconstruction of the heterogeneity may be an important part of a management scheme if higher species richness in the ground vegetation within these stands should be accomplished. Furthermore, larger enclosures which include ecological similar areas may secure seed dispersal between sites. Though the deer population in Gribskov to a large extent also affects the seed dispersal, endozoochorous seed dispersal by the horses may be more successful. This is due to their larger excrements to a higher degree affect the competition from the surrounding vegetation together with the more constant micro-environment (Buttenschøn, 2007). And though many plant species are not adopted to zoochory, the horses in large enclosures may contribute to the possible rare event of long-distance dispersal, which may be ecologically important in species migration and meta-population dynamics (Bruun & Fritzbøger, 2002).

Evaluation of the effect of the grazing management has to be done in the light of the objective(s). If the aim of the grazing management is to increase the diversity, the success depends on how the diversity is measured. And even though the diversity may be found to increase, the new species colonizing the area may not all be desirable. With the higher species richness found in this study in the grazed sites, desirable species like Veronica officinalis, Viola reichenbachiana / V. riviniana and Epipactis helleborine, which all a indicators of a woodland in relatively good condition, (Fredshavn & Ejrnæs, 2007; Fredshavn et al., 2008) are also increasing their abundance in the grazed sites (Appendix 1). However, the higher invasibility in the grazed site may also lead to colonization of undesirable plant species. These species may either be undesirable because they are regarded as problematic in the habitat like Urtica dioica, Plantago major, Circium vulgare, Galium aparine, Poa annua, Ranunculus repens, Epilobium angustifolium and Stellaria media, because they are indigenous like Larix sp., Picea abies, Prunus cerasifera, Impatiens parviflor and Abies sp., or even invasive like Heracelum mantegazzianum. All the above listed species are found in higher abundances in grazed sites compared to the ungrazed (Appendix 1). Consequently, if the undesirable species are increasing their abundance more than the desirable, then the species index will decrease as observed for several of the sites (Figure 49). However, cautions may be taken before interpretation of the species index. The index may be affected by germination of seedlings of woody species that will never survive due to a high over story density and neither is the density of the vegetation a part of the index. Thus, the index may calculate values that are much higher, and hence better, than the real condition. This

77 seems especially to constitute a problem in the beech stands in this study. Furthermore, the species index is only based on the vascular plant species even though rare species of e.g. bryophytes or lichens may be present, giving the site a high conservation value in its own right. However, the opposite may also be the case. For example, the invasive moss species Campylopus introflexus was registered in the grazed beech stand at Sandskredssøen (pers. obs.), and even though carpets of C. introflexus potentially may threaten the regeneration of Calluna vulgaris (Weidema, 2000) the presence of C. introflexus has no influence on the species index. Thus, it is natural that the species index normally are interpreted in the combination of the structure index (Fredshavn & Ejrnæs, 2007; Fredshavn et al., 2008) to give a more correct indication of the woodland condition.

It is still difficult to predict how the horses will affect the dynamics in the woodland. The effects of the management on the ecosystem are best understood in the context of relatively long-term dynamic processes (Huntly, 1991) and with reference to the long life span of trees the processes may be understood only as systems that change over centuries. Thus, it is important that the management not only are focusing on maintaining nature types in certain succession states, but also are aiming for the creation of natural transition zones between open habitats and forest stands with naturally spatial and temporal dynamics and thus great variations in habitats. Grazing horses at the right density may play an important role in creating and maintaining such patch dynamics, but natural disturbances like windthrows, seasonal flooding and fire are also important dynamic processes which help to maintain the woodland biodiversity (Ejrnæs et al., 2011). As the beech are intolerant to even relatively short-term flooding and the thin bark provides little protection from fires (Packham et al., 2012), it is likely that the beech under natural dynamics are more impeded than what grazing by large herbivores can do alone. However, giving the size of Gribskov, and the several kilometres of draining ditches, all these naturally dynamics are not likely to occur. Hence, to re-establish natural processes and dynamics, grazing with livestock as horses cannot stand alone though it is an important part.

7. Conclusion Even after just a few years grazing management the horses in Gribskov have clearly affected the ground vegetation, though the effect is not constant in all areas. The effect varies in relation to the overstory density, and thus productivity in the ground vegetation, and the present species composition. A generally positive effect is found in the alder swamps and the oak stands where the ground vegetation without the presence of the horses is rather tall. By grazing, the horses have decreased the vegetation height and thus the interspecific competition, increased the heterogeneity and ultimately allowed more vascular plant species to germinate and establish. The grazed species communities have higher abundances of disturbance tolerant species. However, in the young oak stand on acidic soil no increase is found in the species richness though a significantly lower vegetation height and a higher heterogeneity. This result is presumably due to a general lower propagule pressure and the high abundance of Deschampsia flexuosa, which frequency has remained unaltered despite the grazing management. Thus, with assumed allelopathic effects and an unaltered assimilation rate of nitrogen, which occur in limited amounts, D. flexuosa may still retain its competitive abilities and thus the invasibility of the site has remained unaltered. In the beech stands the effect is however somewhat different. The competitive superior beech trees are unaffected and

78 the ground vegetation remains more or less vegetation-less. Although the vegetation composition and the species richness are unaltered, the positive correlation between the vegetation height and the species richness clearly indicate the horses’ negatively impacts.

The richness of woody species appear in general to be unaffected by the horses and are particular in the beech stands determined by the overstory density. Even though the horses are found to have a high browsing impact on beech, Alnus incana, Euonymus europaeus and Carpinus betulus is it only in the beech stand in Kollerup Enghave where the horses have a fatal effect on the regeneration of beech. This is due to the combination between the highest browsing impact found in this study and the dense canopy impeding light penetration. With the short time of management it appears yet not possible to conclude whether, and to which degree, the horses are able to impede and delay the succession in the other stands where more light are penetrating the canopies. However, the horses are able to delay growth rate of saplings of A. incana though the browsing apparently are increasing the reproduction by root suckers and thus actually may increase the density of saplings. Against the expectations no indications of a higher browsing impact in the all year round grazed area were found. As the frequency of bark stripping was negligible, and it was the only expected way the horses could have impacted the amount of deadwood, the number of trunks with decaying parts and hollowness, these structures were neither found affected by the presence of the horses. However, with the high browsing impact, especially on beech, it appears likely that the horses in the long- term somehow will affect the woodland structures, particularly in the transition zones between open habitats and forest stands with high overstory densities.

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References Adams, S. N. (1975). Sheep and cattle grazing in forests - a review. Journal of Applied Ecology, 12(1), 143-152. doi: 10.2307/2401724 Adsersen, H., & Mølgaard, P. (2003). Floraen på Isefjordens småholme i øbiogeografisk belysning. Flora og Fauna, 109(3+4), 71-79. Alexander, K. (2013). Ancient trees, grazing landscapes and the conservation of deadwood and wood decay invertebrates. In I. D. Rotherham (Ed.), Trees, Forested Landscapes and Grazing Animals. A European Perspective on Woodlands and Grazed Treescapes. Bakker, E. S., Olff, H., Vandenberghe, C., De Maeyer, K., Smit, R., Gleichman, J. M., & Vera, F. W. M. (2004). Ecological anachronisms in the recruitment of temperate light-demanding tree species in wooded pastures. Journal of Applied Ecology, 41(3), 571-582. doi: 10.1111/j.0021-8901.2004.00908.x Bardgett, R. D. (2005). The Biology of Soil. A Community and Ecosystem Approach: Oxford University Press. Bidlack, W. R., Brown, R. C., & Mohan, C. (1986). Nutrirional parameters that alter hepatic drug-metabolism, conjugation, and toxicity. Federation Proceedings, 45(2), 142-148. Birks, H. J. B. (2005). Mind the gap: how open were European primeval forests? Trends in Ecology & Evolution, 20(4), 154-156. doi: 10.1016/j.tree.2005.02.001 Bokdam, J. (2003). Nature conservation and grazing management. Free-ranging cattle as a driving force for cyclic succession. (PhD thesis), Wageningen University, Wageningen, The Netherlands. Bokdam, J., Cornelius, R., & Krüsi, B. O. (2001). A Conceptual Model for Nutrient-Mediated Successinal Grazing Mosaics. In B. Gerken & M. Görner (Eds.), Neue Modelle zu Maßnahmen der Landschaftsentwicklung mit großen Pflanzenfressern - Praktische Erfahrungen bei der Umsetzung: Natur- und Kulturlandschaft 4, Höxter/Jena, 2001. Bradshaw, R. (2001). Vegetation of the past - structure and composition Report from the conferennce (Summary/Discussion). The role of large herbivores in north-west European vegetation. . Bradshaw, R., & Mitchell, F. J. G. (1999). The palaeoecological approach to reconstructing former grazing-vegetation interactions. Forest Ecology and Management, 120(1-3), 3-12. doi: 10.1016/s0378-1127(98)00538-6 Bruun, H. H., Ejrnæs, R., Heilmann-Clausen, J., Aude, E., Poulsen, R. S., & J, P. (2009). Dødt ved og gamle træer - hvor meget og hvor mange i de danske skove? Jord og Viden(17), 50-53. Bruun, H. H., & Fritzbøger, B. (2002). The past impact of livestock husbandry on dispersal of plant seeds in the landscape of Denmark. Ambio, 31(5), 425-431. doi: 10.1639/0044-7447(2002)031[0425:tpiolh]2.0.co;2 Bruun, H. H., & Heilmann-Clausen, J. (2012). Hvordan sikre vi skovenes biodiversitet? In H. Meltofte (Ed.), Danmarks natur frem mod 2020 - om at stoppe tabet af biologisk mangfoldighed: Det Grønne Kontaktudvalg. Buttenschøn, J., & Buttenschøn, R. M. (1985). Grazing Expirements with Cattle and Sheep on Nutrient Poor, Acidic Grassland and Heath. IV: Establishment of woody species. Natura Jutlandica, 21(7), 117-140. Buttenschøn, R. M. (2007). Græsning og høslæt i naturplejen: Miljøministeriet, Skov- og Naturstyrelsen og Center for Skov, Landskab og Planlægning, Københavns Universitet. Buttenschøn, R. M., & Buttenschøn, J. (2013). Woodland grazing with cattle: Results from 25 years of grazing in acidophilus pedunculate oak (Quercus robur) woodland. In I. D. Rotherham (Ed.), Trees, Forested Landscapes and Grazing Animals. A European Perspective on Woodlands and Grazed Treescapes (pp. 317-329): Routledge. Buttenschøn, R. M., & Holst, T. (1999). Vildt i forhold til græsning og hegning: videnblade. Park- og Landskab, 6.0-14. Buttenschøn, R. M., Madsen, T. L., Madsen, P., & Olesen, C. R. (2009a). Effekten af græsning på skovbryn, skov og skovetablering. In N. Kanstrup, T. Asferg, M. Flinterup, B. J. Thorsen & T. S. Jensen (Eds.), Vildt & Landskab - Resultater af 6 års integreret forskning i Danmark 2003-2008: Skov- og Naturstyrelsen, Danmarks Miljøundersøgelser, Aarhus Universitet, Skov & Landskab, Københavns Universitet og Danmarks Jægerforbund. Buttenschøn, R. M., Madsen, T. L., Madsen, P., & Olesen, C. R. (2009b). Husdyr og vildt som naturplejere - er de lige gode naturplejere, og holder husdyr vildtet væk? In N. Kanstrup, T. Asferg, M. Flinterup, B. J. Thorsen & T. S. Jensen (Eds.), Vildt & Landskab - Resultater af 6 års integreret forskning i Danmark 2003-2008: Skov- og Naturstyrelsen, Danmarks Miljøundersøgelser, Aarhus Universitet, Skov & Landskab, Københavns Universitet og Danmarks Jægerforbund. Bülow-Olsen, A. (1980). Changes in the species composition in an area dominated by Deschampsia flexuosa (L) Trin - as a result of cattle grazing. Biological Conservation, 18(4), 257-270. doi: 10.1016/0006-3207(80)90002-6 Chapin III, F. S., Matson, P. A., & Vitousek, P. M. (2011). Principles of Terrestial Ecosystem Ecology (2 ed.): Springer.

80

Cooper, M. R., & Johnson, A. W. (1998). Poisonous Plants and Fungi in Britain. Animal and Human Poisoning (2. ed.). The Stationery Office. Cornelissen, P., & Vulink, T. (2001). Effects of cattle and horses on vegetation structure. Are cattle and horses browsers enough to stop development of shrubs and trees? In B. Gerken & M. Görner (Eds.), Neue Modelle zu Maßnahmen der Landschaftsentwicklung mit großen Pflanzenfressern - Praktische Erfahrungen bei der Umsetzung: Natur- und Kulturlandschaft 4, Höxter/Jena 2001. Cosyns, E., Delporte, A., Lens, L., & Hoffmann, M. (2005). Germination success of temperate grassland species after passage through ungulate and rabbit guts. Journal of Ecology, 93(2), 353-361. doi: 10.1111/j.1365- 2745.2005.00982.x Cosyns, E., & Hoffmann, M. (2005). Horse dung germinable seed content in relation to plant species abundance, diet composition and seed characteristics. Basic and Applied Ecology, 6(1), 11-24. doi: 10.1016/j.baae.2004.09.012 Couvreur, M., Verheyen, K., & Hermy, M. (2005). Experimental assessment of plant seed retention times in fur of cattle and horse. Flora, 200(2), 136-147. doi: 10.1016/j.flora.2004.06.003 Crawley, M. (1997). Plant ecology (2 ed.): Blackwell Science. Dahlberg, A., & Stokland, J. N. (2004). Vedlevande arters krav på substrat - sammanställning och analys av 3600 arter.: Skogstyrelsen. Davy, A. J. (1980). Deschampsia caespitosa (L.) Beauv. (Aira cespitosa L., Deschampsia cespitosa (L.) Beauv.). Journal of Ecology, 68(3), 1075-1096. doi: 10.2307/2259475 De Jong, J., Dahlberg, A., & Stokland, J. N. (2004). Död ved i skogen. Hur mycket behövs för att bevara den biologiska mångfalden? Svensk Botanisk Tidskrift, 98(5), 278-297. Donath, T. W., & Eckstein, R. L. (2010). Effects of bryophytes and grass litter on seedling emergence vary by vertical seed position and seed size. Plant Ecology, 207(2), 257-268. doi: 10.1007/s11258-009-9670-8 Ejrnæs, R., Wiberg-Larsen, P., Holm, T. E., Josefson, A. B., Strandberg, B., Nygaard, B., . . . Levin, G. (2011). Danmarks Biodiversitet 2010 - status, udvikling og trusler. (pp. 152). Danmarks Miljøundersøgelser, Aarhus Universitet. Ellenberg, H., Weber, H. E., Düll, R., Wirth, V., Werner, W., & Paulißen, D. (1992). Zeigerwerte von Pflanzen in Mitteleuropa (2. ed.): Göttingen Scripta Geobotanica 18: Göttingen. Facelli, J. M., & Pickett, S. T. A. (1991). Plant litter: Its dynamics and effects on plant community structure. Botanical Review, 57(1), 1-32. doi: 10.1007/bf02858763 Faurholdt, N., & Schou, J. C. (2012). Danmarks skærmplanter: BFN's forlag. Fischer, S. F., Poschlod, P., & Beinlich, B. (1996). Experimental studies on the dispersal of plants and animals on sheep in calcareous grasslands. Journal of Applied Ecology, 33(5), 1206-1222. doi: 10.2307/2404699 Frederiksen, S., Rasmussen, F. N., & Seberg, O. (2012). Dansk flora (2 ed.): Gyldendal. Fredshavn, J., & Ejrnæs, R. (2007). Beregning af naturtilstand ved brug af simple indikatorer (2 ed.): Danmarks Miljøundersøgelser, Aarhus Universitet. - Faglig rapport fra DMU nr. 599. Fredshavn, J., Ejrnæs, R., & Nygaard, B. (2011). TA-N03 Kortlægning af terrestiske naturtyper (1.04 ed.): DCE - Nationalt Center for Miljø og Energi. Aarhus Universitet. Fredshavn, J., Johannsen, V. K., Ejrnæs, R., Nielsen, K. E., & Rune, F. (2008). Beregning af skovtilstand - tilstandsvurdering af Habitatdirektivets skovtyper: Danmarks Miljøundersøgelser, Aarhus Universitet. - Faglig rapport fra DMU nr. 696. Fredshavn, J., Nielsen, K. E., Ejrnæs, R., & Nygaard, B. (2014a). TA-N01 Overvågning af terrestiske naturtyper (2 ed.): DCE - Nationalt Center for Miljø og Energi. Aarhus Universitet. Fredshavn, J., Søgaard, B., Nygaard, B., Johannson, L. S., Wiberg-Larsen, P., Dahl, K., . . . Teilmann, J. (2014b). Bevaringsstatus for naturtyper og arter. Habitatdirektivets Artikel 17 rapportering. DCE - Nationalt Center for Miljø og Energi. Aarhus Universitet. Freeland, W. J., & Janzen, D. H. (1974). Strategies in herbivory by mammals - the role of plant secondary compounds. American Naturalist, 108(961), 269-289. doi: 10.1086/282907 Friis Møller, P. (2010). Naturen i Danmark. Skovene: Gyldendal. Gill, R. (2006). The influence of large herbivores on tree recruitment and forest dynamics. In K. Danell, R. Bergström, P. Duncan & J. Pastor (Eds.), Large Herbivore Ecology, Ecosystem Dynamics and Conservation: Cambridge University Press. Gill, R. M. A. (1992). A Review of Damage by Mammals in North Temperate Forests: 1. Deer. Forestry, 65(2), 145-169. doi: 10.1093/forestry/65.2.145 Gill, R. M. A., & Beardall, V. (2001). The impact of deer on woodlands: the effects of browsing and seed dispersal on vegetation structure and composition. Forestry, 74(3), 209-218. doi: 10.1093/forestry/74.3.209

81

Gordon, I. J., & Prins, H. H. T. (2008). The Ecology of Browsing and Grazing: Springer. Green, T. (2013). Ancient trees and wood-pastures. Observations on recent progress. In I. D. Rotherham (Ed.), Trees, Forested Landscapes and Grazing Animals. A European Perspective on Woodlands and Grazed Treescapes. Grime, J. P. (1973). Competitive Exclusion in Herbaceous Vegetation. Nature, 242(5396), 344-347. doi: 10.1038/242344a0 Grime, J. P. (1979). Plant Strategies and Vegatation Processes: John Wiley & Sons, Ltd. Graae, B. J. (2002). The role of epizoochorous seed dispersal of forest plant species in a fragmented landscape. Seed Science Research, 12(2), 113-120. doi: 10.1079/ssr2002103 Harmon, M. E., Franklin, J. F., Swanson, F. J., Sollins, P., Gregory, S. V., Lattin, J. D., . . . Cummins, K. W. (1986). Ecology of coarse woody debris in temperate ecosystems. Advances in Ecological Research, 15, 133-302. doi: 10.1016/s0065-2504(08)60121-x Harper, J. L. (1957). Biological Flora of the British Isles: Ranunculus repens L. Journal of Ecology, 45(1), 314-325. doi: 10.2307/2257092 Hauck, M., & Popp, A. (2010). The Effects of Cattle Browsing on Woody Vegetation in Large-Scale Pastures of the Thuringian Rhoen and the South Black Forest. In H. Plachter & U. Hampicke (Eds.), Large-scale Livestock Grazing. A Management Tool for Nature Conservation: Springer. Hester, A. J., Bergman, M., Iason, G. R., & Moen, J. (2006). Impacts of large herbivores on plant community structure and dynamics. In K. Danell, R. Bergström, P. Duncan & J. Pastor (Eds.), Large Herbivore Ecology, Ecosystem Dynamics and Conservation: Cambridge University Press. Hester, A. J., Edenius, L., Buttenschøn, R. M., & Kuiters, A. T. (2000). Interactions between forests and herbivores: the role of controlled grazing experiments. Forestry, 73(4), 381-391. doi: 10.1093/forestry/73.4.381 Hester, A. J., Millard, P., Baillie, G. J., & Wendler, R. (2004). How does timing of browsing affect above- and below- ground growth of Betula pendula, Pinus sylvestris and Sorbus aucuparia? Oikos, 105(3), 536-550. doi: 10.1111/j.0030-1299.2004.12605.x Hobbs, N. T. (1996). Modification of ecosystems by ungulates. Journal of Wildlife Management, 60(4), 695-713. doi: 10.2307/3802368 Hodgson, J. G., Wilson, P. J., Hunt, R., Grime, J. P., & Thompson, K. (1999). Allocating C-S-R plant functional types: a soft approach to a hard problem. Oikos, 85(2), 282-294. doi: 10.2307/3546494 Hofmann, R. R. (1989). Evolutionary steps of ecophysiological adaptation and diversification of ruminants - a comparative view of their digestive system. Oecologia, 78(4), 443-457. doi: 10.1007/bf00378733 Holtmeier, F. (2015). Animals' Influence on the Landscape and Ecological Importance: Springer. Hunt, R., Hodgson, J. G., Thompson, K., Bungener, P., Dunnett, N. P., & Askew, A. P. (2004). A new practical tool for deriving a functional signature for herbaceous vegetation. Applied Vegetation Science, 7(2), 163-170. doi: 10.1658/1402-2001(2004)007[0163:anptfd]2.0.co;2 Huntly, N. (1991). Herbivores and the dynamics of communities and ecosystems. Annual Review of Ecology and Systematics, 22, 477-503. doi: 10.1146/annurev.ecolsys.22.1.477 Hutchings, M. J., & Price, E. A. C. (1999). Glechoma hederacea L. (Nepeta glechoma Benth., N. hederacea (L.) Trev.). Journal of Ecology, 87(2), 347-364. doi: 10.1046/j.1365-2745.1999.00358.x Janis, C. (1976). Evolutionary strategy of Equidae and origins of rumen and cecal digestion. Evolution, 30(4), 757-774. doi: 10.2307/2407816 Jaroszewicz, B., Pirozikow, E., & Sagehorn, R. (2009). Endozoochory by European bison (Bison bonasus) in Bialowieza Primeval Forest across a management gradient. Forest Ecology and Management, 258(1), 11-17. doi: 10.1016/j.foreco.2009.03.040 Jarvis, P. G. (1964). Interference by Deschampsia flexuosa(L) Trin. Oikos, 15(1), 56-&. doi: 10.2307/3564748 Johannsen, V. K., Dippel, T. M., Møller, P. F., Heilmann-Clausen, J., Ejrnæs, R., Larsen, J. B., . . . Hansen, G. K. (2013). Evaluering af indsatsen for biodiversiteten i de danske skove 1992 - 2012. 90 s. ill. Københavns Universitet. Institut for Geovidenskab og Naturforvaltning. Jones, C. G., Lawton, J. H., & Shachak, M. (1994). Organisms as ecosystem engineers. Oikos, 69(3), 373-386. doi: 10.2307/3545850 Jonsell, M., Weslien, J., & Ehnstrom, B. (1998). Substrate requirements of red-listed saproxylic invertebrates in Sweden. Biodiversity and Conservation, 7(6), 749-764. doi: 10.1023/a:1008888319031 Jönsson, M. T., Thor, G., & Johansson, P. (2011). Environmental and historical effects on lichen diversity in managed and unmanaged wooded meadows. Applied Vegetation Science, 14(1), 120-131. doi: 10.1111/j.1654- 109X.2010.01096.x

82

Kanstrup, N. (2013). Kronvildtet på Sjælland - bestandene anno 2013 og nogle bud på udviklingen. Institut for Geovidenskab og Naturforvaltning, Københavns Universitet. Kanstrup, N., Madsen, P., Stenkjær, K., Buttenschøn, R. M., & Jensen, A. (2014). Kronvildt på Sjælland. Resultatet af tre års praksisorienteret forskning og forvaltning. Institut for Geovidenskab og Naturforvaltning, Københavns Universitet. Keenan, D. M. (1986). Bark chewing by horses grazed on irrigated pasture. Australian Veterinary Journal, 63(7), 234- 235. doi: 10.1111/j.1751-0813.1986.tb03008.x Kent, M. (2012). Vegetation Description and Data Analysis : A Practial approach (2 ed.): Wiley-Blackwell. Király, I., Nascimbene, J., Tinya, F., & Ódor, P. (2013). Factors influencing epiphytic bryophyte and lichen species richness at different spatial scales in managed temperate forests. Biodiversity and Conservation, 22(1), 209- 223. doi: 10.1007/s10531-012-0415-y Kirby, K. J., & Baker, A. (2013). The dynamics of pre-Neolithic European landscapes and their relevance to modern conservation. In I. D. Rotherham (Ed.), Trees, Forested Landscapes and Grazing Animals. A European Perspective on Woodlands and Grazed Treescapes (pp. 87-98): Routledge. Knapp, A. K., & Seastedt, T. R. (1986). Detritus accumulation limits productivity af tallgrass prairie. Bioscience, 36(10), 662-668. doi: 10.2307/1310387 Kolter, L., Schach, C., & Weber, T. (1999). Habitat Use of Feral and Przewalski's Horses. In B. Gerken & M. Görner (Eds.), Europäische Landschaftsentwicklung mit großen Weidertieren - Geschichte, Modelle und Perspektiven: Natur- und Kulturlanschaft 3, Höxter/Jena 1999. Koorem, K., Price, J. N., & Moora, M. (2011). Species-Specific Effects of Woody Litter on Seedling Emergence and Growth of Herbaceous Plants. Plos One, 6(10), 7. doi: 10.1371/journal.pone.0026505 Krojerova-Prokesova, J., Barancekova, M., Sustr, P., & Heurich, M. (2010). Feeding patterns of red deer Cervus elaphus along an altitudinal gradient in the Bohemian Forest: effect of habitat and season. Wildlife Biology, 16(2), 173-184. doi: 10.2981/09-004 Kuiters, A. T., van der Sluijs, L. A. M., & Wytema, G. A. (2006). Selective bark-stripping of beech, Fagus sylvatica, by free-ranging horses. Forest Ecology and Management, 222(1-3), 1-8. doi: 10.1016/j.foreco.2005.09.019 Lamoot, I., Meert, C., & Hoffmann, M. (2005). Habitat use of ponies and cattle foraging together in a coastal dune area. Biological Conservation, 122(4), 523-536. doi: 10.1016/j.biocon.2004.09.009 Larsen, S. N., & Vikstrøm, T. (1995). Ferske enge - en beskyttet naturtype: Miljø- og Energiministeriet, Skov- og Naturstyrelsen. Latham, J. (1999). Interspecific interactions of ungulates in European forests: an overview. Forest Ecology and Management, 120(1-3), 13-21. doi: 10.1016/s0378-1127(98)00539-8 Lemmon, P. E. (1957). A New Istrument for Measuring Forest Overstory Density. Journal of Forestry, 55(9), 667-669. Mayle, B. (1999). Domestic Stock Grazing to Enhance Woodland Biodiversity. Forestry Commission. Information note. McDonald, J. H. (2009). Handbook of Biological Statistics (2 ed.): Sparky House Publishing, Baltimore, Maryland. McNaughton, S. J. (1979). Grazing as an optimization process - grass ungulate relationships in the serengeti. American Naturalist, 113(5), 691-703. doi: 10.1086/283426 Milchunas, D. G., Sala, O. E., & Lauenroth, W. K. (1988). A Generalized-Model of the Effects of Grazing by Large Herbivores on Grassland Community Structure. American Naturalist, 132(1), 87-106. doi: 10.1086/284839 Miljøministeriet. (2014). Naturplan Danmark - Vores fælles natur. Miljøministeriet. mim.dk. Miller, A. M., McArthur, C., & Smethurst, P. J. (2007). Effects of within-patch characteristics on the vulnerability of a plant to herbivory. Oikos, 116(1), 41-52. doi: 10.1111/j.2006.0030-1299.15331.x Mitchell, F. J. G. (2005). How open were European primeval forests? Hypothesis testing using palaeoecological data. Journal of Ecology, 93(1), 168-177. doi: 10.1111/j.1365-2745.2004.00964.x Mitchell, F. J. G., & Kirby, K. J. (1990). The Impact of Large Herbivores on the Conservation of Semi-natural Woods in the British Uplands. Forestry, 63(4), 333-353. doi: 10.1093/forestry/63.4.333 Mossberg, B., & Stenberg, L. (2003). Den nye nordiske flora: Gyldendal. Nygaard, B., Bruun, H. H., Heilmann-Clausen, J., Damgaard, C., Ejrnæs, R., & Nielsen, K. E. (2013). Vurdering af bevaringsstatus for skov. DCE - Nationalt Center for Miljø og Energi. Aarhus Universitet. Odor, P., Kiraly, I., Tinya, F., Bortignon, F., & Nascimbene, J. (2013). Patterns and drivers of species composition of epiphytic bryophytes and lichens in managed temperate forests. Forest Ecology and Management, 306, 256- 265. doi: 10.1016/j.foreco.2013.07.001 Olff, H., & Ritchie, M. E. (1998). Effects of herbivores on grassland plant diversity. Trends in Ecology & Evolution, 13(7), 261-265. doi: 10.1016/s0169-5347(98)01364-0

83

Overballe-Petersen, M. V., Nielsen, A. B., & Bradshaw, R. H. W. (2013a). Quantitative vegetation reconstruction from pollen analysis and historical inventory data around a Danish small forest hollow. Journal of Vegetation Science, 24(4), 755-771. doi: 10.1111/jvs.12007 Overballe-Petersen, M. V., Nielsen, A. B., Hannon, G. E., Halsall, K., & Bradshaw, R. H. W. (2013b). Long-term forest dynamics at Gribskov, eastern Denmark with early-Holocene evidence for thermophilous broadleaved tree species. Holocene, 23(2), 243-254. doi: 10.1177/0959683612455549 Overballe-Petersen, M. V., Raulund-Rasmussen, K., Buttenschøn, R. M., & Bradshaw, R. H. W. (2014). The forest Gribskov, Denmark: lessons from the past qualify contemporary conservation, restoration and forest management. Biodiversity and Conservation, 23(1), 23-37. doi: 10.1007/s10531-013-0582-5 Packham, J. R., Thomas, P. A., Atkinson, M. D., & Degen, T. (2012). Biological Flora of the British Isles: Fagus sylvatica. Journal of Ecology, 100(6), 1557-1608. doi: 10.1111/j.1365-2745.2012.02017.x Petersen, P. M., & Vestergaard, P. (2006). Vegetationsøkologi: Gyldendal. Poland, J., & Clement, E. (2009). The Vegetative Key to the British Flora - A New Approach to Plant Identification: John Poland. Potts, S. G., Vulliamy, B., Dafni, A., Ne'eman, G., & Willmer, P. (2003). Linking bees and flowers: How do floral communities structure pollinator communities? Ecology, 84(10), 2628-2642. doi: 10.1890/02-0136 Putman, R. J. (1986). Grazing in Temperate Ecosystems. Large Herbivores and the Ecology of the New Forest: Timber Press. Putman, R. J., Pratt, R. M., Ekins, J. R., & Edwards, P. J. (1987). Food and Feeding Behaviour of Cattle and Ponies in the New Forest, Hampshire. Journal of Applied Ecology, 24(2), 369-380. doi: 10.2307/2403881 Rackham, O. (2013). Woodland and wood-pasture. In I. D. Rotherham (Ed.), Trees, Forested Landscapes and Grazing Animals. A European Perspective on Woodlands and Grazed Treescapes (pp. 11-22): Routledge. Rosenthal, J. P., & Kotanen, P. M. (1994). Terrestrial plant tolerance to herbivory. Trends in Ecology & Evolution, 9(4), 145-148. doi: 10.1016/0169-5347(94)90180-5 Roulston, T. a. H., & Goodell, K. (2011). The Role of Resources and Risks in Regulating Wild Bee Populations. In M. R. Berenbaum, R. T. Carde & G. E. Robinson (Eds.), Annual Review of Entomology, Vol 56 (Vol. 56, pp. 293-312). Rune, F. (2009). Gribskov. Fredensborg: Forlaget Esrum Sø. Rupp, M. (2013). Creation of open woodlands through pasture: Genesis, relevance as biotopes, value in the landscape and in nature conservation in south-west Germany. In I. D. Rotherham (Ed.), Trees, Forested Landscapes and Grazing Animals. A European Perspective on Woodlands and Grazed Treescapes: Routledge. Samojlik, T., & Kuijper, D. (2013). Grazed wood-pasture versus browsed high forests: Impact of ungulates on forest landscapes from the perspective of the Białowieża Primeval Forest. In I. D. Rotherham (Ed.), Trees, Forested Landscapes and Grazing Animals. A European Perspective on Woodlands and Grazed Treescapes (pp. 143- 162): Routledge. Schippers, P., van Teeffelen, A. J. A., Verboom, J., Vos, C. C., Kramer, K., & WallisDeVries, M. F. (2014). The impact of large herbivores on woodland-grassland dynamics in fragmented landscapes: The role of spatial configuration and disturbance. Ecological Complexity, 17, 20-31. doi: 10.1016/j.ecocom.2013.07.002 Schou, J. C. (2006). Danmarks halvgræsser (2 ed.): BFN's forlag. Schou, J. C., Wind, P., & Lægaard, S. (2010). Danmarks siv og frytler: BFN's forlag. Schou, J. C., Wind, P., & Lægaard, S. (2014). Danmarks græsser (2 ed.): BFN's forlag. Skarpe, C., & Hester, A. J. (2008). Plant Traits, Browsing and Grazing Herbivores, and Vegetation Dymanics. In I. J. Gordon & H. H. T. Prins (Eds.), The Ecology of Browsing and Grazing: Springer. Sobey, D. G. (1981). Stellaria media (L.) Vill. Journal of Ecology, 69(1), 311-335. doi: 10.2307/2259833 Spencer, J. (2001). A Brief Ecological History of the New Forest Short Communication. In B. Gerken & M. Görner (Eds.), Neue Modelle zu Maßnahmen der Landschaftsentwicklung mit großen Pflanzenfressern - Praktische Erfahrungen bei der Umsetzung: Natur- und Kulturlandschaft 4, Höxter/Jena, 2001. Stroh, P. A., Mountford, J. O., & Hughes, F. M. R. (2012). The potential for endozoochorous dispersal of temperate fen plant species by free-roaming horses. Applied Vegetation Science, 15(3), 359-368. doi: 10.1111/j.1654- 109X.2011.01172.x Suchant, R., Türk, S., & Roth, R. (1998). Grazing problems in Germany: balance or imbalance between wildlife and habitat? In J. Humphrey, R. Gill & J. Claridge (Eds.), Grazing as a Management Tool in European Forest Ecosystems (pp. 36-44): Forestry Commission Technical Paper 25. Forestry Commission, Edinburgh. Taylor, K. (2009). Biological Flora of the British Isles: Urtica dioica L. Journal of Ecology, 97(6), 1436-1458. doi: 10.1111/j.1365-2745.2009.01575.x Thomas, P. A. (2014). Trees: Their Natural History (2. ed.): Cambridge University Press.

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Thomas, P. A., El-Barghathi, M., & Polwart, A. (2011). Biological Flora of the British Isles: Euonymus europaeus L. Journal of Ecology, 99(1), 345-365. doi: 10.1111/j.1365-2745.2010.01760.x Tolhurst, S. A., & Oates, M. H. (2001). The breed profiles handbook: Peterbourgh: English Nature / Grazing Animals Project. Van Dyne, G. M., Brockington, N. R., Szocs, Z., Duke, J., & Ribic, C. A. (1980). Large herbivore subsystem. In A. I. Breymeyer & G. M. Van Dyne (Eds.), Grasslands, systems analysis and man: Cambridge University Press. Van Uytvanck, J., & Hoffmann, M. (2009). Impact of grazing management with large herbivores on forest ground flora and bramble understorey. Acta Oecologica-International Journal of Ecology, 35(4), 523-532. doi: 10.1016/j.actao.2009.04.001 van Wieren, S. E., & Bakker, J. P. (2006). The Impact of Browsing and Grazing Herbivores on Biodiversity. In I. J. Gordon & H. H. T. Prins (Eds.), The Ecology of Browsing and Grazing: Springer. Vera, F. W. M. (2000). Grazing ecology and forest history. Vera, F. W. M. (2013). Can't see the trees for the forest. In I. D. Rotherham (Ed.), Trees, Forested Landscapes and Grazing Animals. A European Perspective on Woodlands and Grazed Treescapes (pp. 99-126): Routledge. Vestergaard, P. (2007). Naturen i Danmark. Det åbne land: Gyldendal. Vulliamy, B., Potts, S. G., & Willmer, P. G. (2006). The effects of cattle grazing on plant-pollinator communities in a fragmented Mediterranean landscape. Oikos, 114(3), 529-543. doi: 10.1111/j.2006.0030-1299.14004.x Watkinson, A. R., Riding, A. E., & Cowie, N. R. (2001). A community and population perspective of the possible role of grazing in determining the ground flora of ancient woodlands. Forestry, 74(3), 231-239. doi: 10.1093/forestry/74.3.231 Weidema, I. R. (2000). Nord 2000: 13 Introduced Species in the Nordic Countries. Nordic Council of Ministers. Westrich, P. (1990). Die Wildbienen Baden‐Württembergs - Spezieller Teil: Die Gattungen und Arten. Stuttgard, Germany: Eugen Ulmer. Xiong, S. J., & Nilsson, C. (1999). The effects of plant litter on vegetation: a meta-analysis. Journal of Ecology, 87(6), 984-994. doi: 10.1046/j.1365-2745.1999.00414.x Ziegenhagen, B., & Kausch, W. (1995). Productivity of young shaded oaks (Quercus robur L) as corresponding to shoot morphology and leaf anatomy. Forest Ecology and Management, 72(2-3), 97-108. doi: 10.1016/0378- 1127(94)03482-c

85

Appendix 1 – Ground vegetation

Registration of the ground vegetation within the 5 m circles. X = the species present. The numbers are referring to the relative abundance (%).

Alder swamps Site / plot number Species 1 2 3 4 5 1 2 3 4 5 1 2 3 4 5 X Acer pseudoplatanus 0 X Aegopodium podagraria 12 X X Agrostis canina 12 8 X Agrostis capillaris 4 X X X Alnus glutinosa 20 20 8 X X X X X X X X X X X X Alnus incana 8 20 12 4 0 4 4 32 12 28 36 8 X Alopecurus geniculatus 0 X X Anemone nemorosa 0 4 X Angelica sylvestris 0 X Arctium nemorosum 0 X X X X Athyrium filix-femina 4 8 0 0 X Bidens cernua 0 X X Brachypodium sylvaticum 16 12 X Bromopsis inermis 0 X X X X X X X X X X X X X Calamagrostis canescens 32 12 48 8 8 12 20 4 0 28 8 24 36 X X X X X X X X Calamagrostis epigejos 44 24 12 4 4 36 8 4 X X X X X X Cardamine amara 0 24 0 4 8 8 X X X Cardamine flexuosa 32 16 4 X Cardamine pratensis 12 X X X X X X X X X X X X X X Carex acutiformis 20 48 32 40 68 20 52 24 8 68 64 68 32 32 X X Carex elongata 0 0 X X Carex remota 0 12 X X X X Cerastium fontanum 0 20 24 0 X X Chrysosplenium alternifolium 4 4

86

X Circaea alpina 8 X X Circaea lutetiana 0 0 X X Circaea X intermedia 28 4 X X X X X X X X X X Cirsium palustre 56 8 4 0 0 4 4 28 8 16 X X Cirsium vulgare 0 0 X X X X X X Crataegus sp. 4 8 0 0 0 0 X Dactylis glomerata ssp. glomerata 0 X Dactylis glomerata ssp. lobata 0

X X X X X X X X X X X X X X Deschampsia cespitosa 4 0 8 12 84 100 92 84 100 56 20 16 28 0

X X X X X X X X X X X Dryopteris carthusiana 12 4 4 4 4 0 0 4 0 4 0 X Dryopteris dilatata 0 X X X Dryopteris filix-mas 0 0 0 X X X X X Epilobium hirsutum 0 0 0 0 0 X X X Epilobium montanum 0 0 0 X Epilobium palustre 0 X X Epilobium roseum 0 4 X X X X Epilobium sp. 4 0 4 4 X X X Equisetum arvense 4 0 0 X X X X Euonymus europaeus 4 0 0 4 X X X X X X Eupatorium cannabinum 12 0 0 8 4 0 X X Fagus sylvatica 0 4 X Festuca gigantea 0 X X X X X X X X X X X X Galeopsis bifida 8 0 0 16 20 0 0 4 8 12 4 36 X X X X X X X X X Galium aparine 0 0 0 0 0 0 4 20 16 X X X Galium palustre 8 88 80 X X Geum rivale 0 0 X Geum sp. 8 X X X X Geum urbanum 8 4 0 0

87

X X X X X X X X X X X Glechoma hederacea 4 16 84 44 8 32 48 84 4 36 92 X Heracleum mantegazzianum 0 X Holcus lanatus 4 X X Holcus mollis 36 4 X X Humulus lupulus 0 0 X X X Impatiens noli-tangere 0 0 0 X X X Impatiens parviflora 4 8 0 X Juncus articulatus 0 0 X X X X X X X X Juncus effusus 0 0 0 24 48 24 16 24 X X Lycopus europaeus 0 0 X X X Lysimachia vulgaris 8 0 16 X X X X X X Mentha aquatica 4 8 64 8 12 0 X X X X X Mercurialis perennis 32 0 16 0 24 X X X X X X X X X X X Milium effusum 12 64 0 0 12 12 92 92 96 96 96 X X X X X Moehringia trinervia 12 16 4 4 8 X Myosotis sp. 4 X X X X X X X X X X X Oxalis acetosella 12 40 12 12 0 4 8 52 24 4 8 X X X X X Persicaria hydropiper 28 28 64 40 32 X X X X X X X X X X X X X Phalaris arundinacea 48 88 32 0 8 24 8 12 12 28 4 0 12 X X Plantago major 0 0 X X Poa annua 12 0 X Poa pratensis 4

X X X X X X X X X X X X Poa trivialis 0 0 16 60 84 88 72 92 100 4 12 88

X Prunus cerasifera 0 X Pteridium aquilinum 0 X X X X X X Quercus robur 0 4 0 0 0 0 X Ranunculus flammula 0 X X X X X X X X X Ranunculus repens 8 0 60 0 16 0 12 20 16

88

X X Rosa sp. 0 0 X X Rubus caesius 0 0 X X X X X X X X X X X X X X Rubus idaeus 8 0 20 16 0 0 0 0 4 0 8 12 0 8 X X X Rumex conglomeratus 0 0 0 X X Rumex sanguineus 0 0 X Rumex sp. 0 X Salix cinerea ssp. Cinerea 0 X Salix sp. 0 X Sambucus nigra 4 X X X X Scutellaria galericulata 76 40 68 0 X Senecio sylvaticus 0 X Silene dioica 28 X X X X Solanum dulcamara var. dulcamara 4 0 4 0 X Sorbus aucuparia 0 X X X X Stachys sylvatica 8 8 0 0 X X X X X X X X X X X Stellaria holostea 8 76 0 4 4 4 68 48 36 64 56 X Stellaria media 0 X X X X X Stellaria media/ S. neglecta 16 12 24 4 20 X X X X X X X Taraxacum sp. 0 0 4 8 4 4 4 X X Trifolium repens 4 0 X X Tussilago farfara 0 0 X X X X X X X X X X X X X X X Urtica dioica 80 36 16 72 16 24 16 32 36 32 8 4 20 44 8 X X X X Veronica beccabunga 0 4 4 0 X X Veronica montana 0 0 X Vicia sepium 8 X Viola reichenbachiana/ V. riviniana 4

Oak stands

89

Site / plot number Kollerup Sandskredssøen Sandskredssøen Sandskredssøen Sandskredssøen Lille Hessemose Lille Hessemose Enghave Grazed (young) Grazed (young) (old) Grazed (old) Ungrazed Grazed Ungrazed Ungrazed Species 1 2 3 4 5 1 2 3 4 5 1 2 3 4 5 1 2 3 4 5 1 2 3 4 5 1 2 3 4 5 1 2 3 4 5

X Abies sp. 0

X Acer pseudoplatanus 0

X X X X Agrostis canina 12 8 8 0

X X X X X X X X X X X X X X X X X X X X X X X X X X X X Agrostis capillaris 56 8 20 68 48 4 0 4 4 4 52 44 32 28 24 8 0 12 68 28 8 8 20 28 84 0 32 12

X X Agrostis stolonifera 0 12

X X X X X X X X X X Anemone nemorosa 32 12 8 4 4 8 0 0 8 8

Anthoxanthum X odoratum 4 X X X X X X X Betula pendula 24 48 4 0 4 0 0

X X X X X X Betula pubescens 0 8 8 0 0 0

X X X X X X X X X X X Betula sp. 0 0 4 8 4 4 16 8 4 0 4

Brachypodium X X X X sylvaticum 0 4 12 20 X X X X X X X X X X X X X X X X X X X X X Calamagrostis canescens 16 56 28 28 20 4 16 0 40 84 52 8 0 0 12 24 36 60 68 64 72

X X X X X X X X X X X X X X X X X X X X X X X X X X X X Calamagrostis epigejos 28 40 16 48 52 12 4 0 16 28 44 64 76 80 64 96 80 92 88 84 4 8 8 44 20 0 68 76

X X X X X X X Calluna vulgaris 20 20 0 0 8 0 20

X X X Cardamine flexuosa 4 0 0

X Carex hirta 0

X Carex nigra var. Nigra 8

X Carex ovalis 0

X X X X X X X X X X X Carex pallescens 0 4 4 52 0 0 0 0 8 0 0

X X X X X X X X X X X X X X X X X X X X X Carex pilulifera 16 60 0 48 56 16 68 48 84 60 32 56 60 24 40 4 4 0 4 0 4

X X X X Carex remota 16 12 4 4

X X X X X X X X X X Carex sylvatica 0 4 8 0 0 4 24 4 0 0

X X X X X X X X X X X Carpinus betulus 0 0 4 4 4 8 8 4 44 8 0

X X X X X X Cerastium fontanum 0 20 0 8 4 4

X X Chaerophyllum temulum 0 0

Chrysosplenium X alternifolium 4 X X X X X X Circaea lutetiana 0 16 56 8 0 8

90

X Circaea X intermedia 4

X Cirsium palustre 0

X X X Crataegus laevigata 0 0 0

X Crataegus sp. 0

X X X X X X X X Dactylis glomerata 4 4 16 16 4 4 8 24

Dactylis glomerata ssp. X X X X X X X X X Lobata 4 4 0 0 0 0 0 0 0 X X X X X X X X X X X X X X X X X X X X X X X X X X Deschampsia cespitosa 84 12 96 64 76 0 28 84 96 68 72 12 68 32 80 84 28 72 68 36 60 40 48 40 12 48

X X X X X X X X X X X X X X X X X X X X X X X X X X X X X Deschampsia flexuosa 4 76 0 4 20 100 96 100 88 100 92 92 96 96 92 4 0 4 4 0 8 4 8 4 12 40 4 52 0

X Digitalis purpurea 0

X X X X X X X X X X X X X X X Dryopteris carthusiana 4 0 0 0 0 0 0 0 0 0 0 0 0 0 0

X X X Dryopteris dilatata 0 0 0

X X X Dryopteris filix-mas 0 0 4

X Epilobium angustifolium 0

X X X X Epilobium montanum 0 4 8 0

X Epipactis helleborine 0

X X X Equisetum sylvaticum 0 4 20

X Euonymus europaeus 0

X X X X X X X X X X X X X X X Fagus sylvatica 0 0 0 0 0 4 12 0 44 16 16 0 8 4 0

X Festuca gigantea 24

X Festuca rubra 4

X Frangula alnus 0

X X X X X X X X X X X X X X X Galeopsis bifida 0 0 0 0 8 8 0 4 0 4 0 0 0 0 0

X X X X X Galium aparine 8 20 4 0 4

X Galium odoratum 20

X X X X X X X X X X X X Galium saxatile 8 32 8 0 12 20 4 20 20 16 4 0

X X Geum urbanum 0 0

X X X X X X X X Glechoma hederacea 20 0 40 16 0 20 80 72

X X Holcus lanatus 4 4

91

X X X X X X X X X X X X X X Holcus mollis 4 0 8 8 4 4 60 16 32 96 4 84 28 80

X Hordelymus europaeus 0

X X X Hypericum maculatum 0 0 0

X Hypericum perforatum 4

X Hypericum sp. 4

X X Hypochoeris radicata 0 0

X X X X X X Juncus conglomeratus 4 0 4 4 4 0

X X X X X X X X X X X X X X X X X X Juncus effusus 4 20 16 16 0 0 0 4 0 0 4 4 0 0 12 0 0 28

X Larix decidua 0

X X Larix sp. 0 0

X X X Lonicera periclymenum 0 12 12

X X X X X X X X X X X X X X X X Luzula multiflora 4 0 4 8 0 0 4 4 0 0 0 0 0 12 0 0

X X X X X X X X X X X X X X Luzula pilosa 36 8 20 0 0 0 0 0 4 0 0 0 4 0

X Lysimachia vulgaris 0

X X X X X X X X X X X X Maianthemum bifolium 52 0 12 12 16 0 4 0 16 4 44 12

X X X X X Melampyrum pratense 0 0 0 32 4

X X X Melica uniflora 0 4 0

X X Mercurialis perennis 0 0

X X X X X X X X X X X X X X X X X X Milium effusum 28 4 16 12 0 4 16 0 8 16 24 20 28 64 32 12 36 24

X X X X X X X X X X X Moehringia trinervia 0 0 24 24 4 20 16 8 4 0 8

X X X X X X X Molinia caerulea 32 0 64 44 4 4 8

X X Mycelis muralis 0 8

X Myosotis sp. 4

X X X X X X X X X X X X X X X X X X X X X X X X X X X X Oxalis acetosella 96 60 72 24 60 0 16 8 56 100 80 96 100 100 100 76 100 36 44 68 76 92 80 100 100 100 100 100

X Persicaria hydropiper 0

X X Phegopteris connectilis 4 0

X X X X X X X X X X X X X X X X X X X X X Picea abies 0 0 12 12 8 12 4 12 0 4 0 4 0 4 0 0 0 0 0 0 0

X X X X Plantago major 0 4 0 0

92

X Poa annua 0

X Poa compressa 0

X X Poa pratensis 4 4

X X X X X Poa trivialis 16 4 4 12 12

X Polypodium vulgare 0

X Prunus cerasifera 0

X X Prunus spinosa 4 0

X X X Pteridium aquilinum 0 16 16

X X X X X X X X X X X X X X X X X X X X X X X X X X X X X X X X X X X Quercus robur 32 24 36 16 8 0 0 0 4 4 12 8 12 4 16 12 16 32 24 24 16 12 0 32 0 28 40 20 24 8 40 60 32 44 44

X Quercus rubra 0

X X X Ranunculus repens 0 12 4

X Rosa sp. 0

X X X X X X X X X X X X X X X X X X X X X X X X Rubus idaeus 4 12 4 36 36 0 36 88 96 100 56 72 64 72 60 72 4 4 8 4 0 0 0 0

X X X X Rubus sect. Corylifolii 0 0 0 4

X X X X X X X Rubus sect. Rubus 0 0 0 0 0 0 0

X X X X Rumex acetosella 4 0 4 8

X Rumex sp. 0

X X X Scrophularia nodosa 0 0 0

X X X X X X X X Senecio sylvaticus 0 0 0 0 0 0 0 0

X X X X X X X X X X X X X X X X X X X X Sorbus aucuparia 0 0 0 0 0 0 0 0 0 4 0 0 0 0 0 0 0 0 0 0

X X X Stachys sylvatica 8 0 0

X X X X X X X X X X X X X X X X Stellaria holostea 72 72 84 80 8 4 16 52 72 64 40 56 80 84 76 64

X X X X Stellaria media 44 56 0 12

Stellaria media/ S. X X neglecta 44 16 X Stellaria neglecta 0

X X X X X Taraxacum sp. 0 0 0 0 0

X X X X X X X X X X X X X X X X X X X Trientalis europaea 44 76 16 0 4 44 32 4 72 64 24 8 48 4 8 28 32 40 16

X X X X X X X X X X X X X X X X Urtica dioica 8 4 16 8 0 20 24 28 0 4 8 4 0 4 8 20

93

X X X X X X X X X Vaccinium myrtillus 0 8 12 4 52 44 24 24 44

X X Vaccinium uliginosum 0 0

X X X X Veronica chamaedrys 0 8 0 0

X X X X X Veronica montana 0 0 0 4 8

X X X X X Veronica officinalis 0 0 4 4 0

Viola reichenbachiana/ X X X X X X X X X X X X V. riviniana 0 0 0 40 24 24 16 12 20 60 12 4

Beech stands Site / plot number Kollerup Enghave Kollerup Enghave Sandskredssøen Sandskredssøen Store Hessemose Store Hessemose Grazed Ungrazed Grazed Ungrazed Grazed Ungrazed Species 1 2 3 4 5 1 2 3 4 5 1 2 3 4 5 1 2 3 4 5 1 2 3 4 5 1 2 3 4 5 X X Abies sp. 0 16 X X X X X X X X X X X Agrostis capillaris 0 0 0 4 0 4 0 0 0 4 0 X X Anemone nemorosa 0 4 X X X X X X Betula pendula 0 0 4 0 4 0 X X X Betula pubescens 0 0 4 X X Betula sp. 4 0 X X X X Calamagrostis canescens 0 0 0 0 X X X X X X X X Calamagrostis epigejos 4 4 24 4 4 16 0 0 X X X X X X X X Calluna vulgaris 4 0 8 0 0 0 4 0 X X X X X X X X X X X X X X X X X X X X X X X X Carex pilulifera 24 8 12 8 8 20 12 4 28 32 24 60 36 36 8 28 12 8 36 52 8 0 0 12 X Carex remota 12 X Carex sylvatica 0 X Chenopodium suecicum 0 X Crataegus sp. 0 X X X X X X X X X X X X X X Deschampsia cespitosa 4 0 36 0 0 0 4 0 4 12 0 0 0 0 X X X X X X X X X X X X X X X X X X X X X X X X Deschampsia flexuosa 0 24 28 28 40 48 84 8 32 20 20 76 20 40 68 44 32 40 56 72 8 4 4 0 X X X X X X X X X X X X Dryopteris carthusiana 0 0 8 0 0 0 0 0 0 4 0 0 X Dryopteris dilatata 0 X Euonymus europaeus 0

94

X X X X X X X X X X X X X X X X X X X X X X X X X X X X X X Fagus sylvatica 4 8 4 4 0 8 36 20 16 8 4 4 8 4 0 4 4 4 8 8 0 0 0 0 0 0 0 4 4 0 X X Fraxinus excelsior 0 0 X X X X Galium saxatile 0 0 0 8 X X Holcus mollis 0 8 X Impatiens parviflora 0 X X Juncus effusus 0 0 X X X X X X Larix sp. 4 0 8 0 4 0 X Lonicera periclymenum 0 X X X X Luzula multiflora 4 12 0 8 X X X X X X X X X X X X X X X Luzula pilosa 0 20 0 24 8 0 4 4 0 0 0 0 0 0 4 X X Maianthemum bifolium 0 0 X X X X X X X X X X Milium effusum 0 0 4 4 0 0 0 8 4 4 X X Moehringia trinervia 0 4 X X Molinia caerulea 8 12 X X X X X X X X X X X X X Oxalis acetosella 0 12 4 0 4 96 0 32 4 4 8 4 4 X X X X X X X X X X X X X X X X Picea abies 0 0 4 16 0 0 20 20 20 16 44 52 20 12 44 12 X Pinus sp. 0 X X X Poa nemoralis 0 0 12 X X X Poa trivialis 0 16 4 X Polypodium vulgare 0 X Prunus avium 0 X Prunus cerasifera 0 X X Prunus sp. 0 0 X Prunus spinosa 0 X Pteridium aquilinum 12 X X X X X X X X X X X X Quercus robur 0 0 0 0 0 0 0 4 0 0 0 0 X X X X X X X X Rubus idaeus 0 4 0 0 0 0 28 4 X Salix sp. 0 X X X X X X X X X Sorbus aucuparia 0 4 0 0 4 0 0 0 0

95

X X X X X X Stellaria holostea 12 0 48 0 0 4 X Stellaria media/ S. neglecta 0 X X Trientalis europaea 0 0 X X Urtica dioica 4 4 X Vaccinium myrtillus 4 Viola reichenbachiana/ V. X riviniana 0

96

Appendix 2 – Seedlings and saplings of woody species

Average numbers (±SE) of seedlings and saplings (<0.5 m and 0.5 m – 2.0 m) of all woody species found within the 15 m circles. Only those size categories where individuals were present are shown. †P < 0.1; *P < 0.05; **P < 0.01; ***P < 0.001.

Abies sp. 500 Seedlings NS Saplings <0,5 m Saplings 0,5-2,0 m Grazed 400 Ungrazed 2 300 m

0 200 0

0 100 1

10 NS r e p

8 s l

a 6 u d

i NS v i 4 d n I 2 NS NS NS NS 0 ) ) ) ) ) ) E H E y o H E S H E H E y o H E S H E H E y o H E S H K S K ( ( L K h S K S K ( ( L K h S K S K ( ( L K h S r r k S S k h c h r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e A A O O e B e O O B B O O B B O O B B

Acer pseudoplatanus 10 Seedlings Saplings <0,5 m Grazed Ungrazed 2 m

8 0 0 0 1

r 6 e p

s l

a 4 u d i NS NS v i

d 2 NS n

I NS NS 0 ) ) ) ) E H E y o H E S H E H E y o H E S H K S K ( ( L K h S K S K ( ( L K h S r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e O O B B O O B B

97

Alnus glutinosa 900 Seedlings Saplings <0,5 m Saplings 0,5-2,0 m Grazed 700 Ungrazed 2 500 m 300 0

0 100 0 1

50 r

e 40 p 30 s l 20 a

u 10 d i 10 v i

d 8

n 6 I 4 NS 2 NS 0 ) ) ) ) ) ) E H E y o H E S H E H E y o H E S H E H E y o H E S H K S K ( ( L K h S K S K ( ( L K h S K S K ( ( L K h S r r k S S k h c h r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e A A O O e B e O O B B O O B B O O B B

Alnus incana 600 * Grazed Seedlings Saplings <0,5 m Ungrazed

2 Saplings 0,5-2,0 m

m 500

0 0

0 400 1

r

e *** p

300 s l a

u 200 d i v i d

n 100 I NS 0 ) ) ) ) ) ) E H E y o H E S H E H E y o H E S H E H E y o H E S H K S K ( ( L K h S K S K ( ( L K h S K S K ( ( L K h S r r k S S k h c h r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e A A O O e B e O O B B O O B B O O B B

Betula pendula 9000 Seedlings NS Saplings <0,5 m Grazed 5000 Ungrazed 2 NS

m 1000

0

0 900 0 700 1 500 r NS e 300 NS p

100 s l 50 a

u 40 d i v i 30 NS NS d

n 20 I 10 NS NS NS 0 ) ) ) ) E H E y o H E S H E H E y o H E S H K S K ( ( L K h S K S K ( ( L K h S r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e O O B B O O B B

98

Betula pubescens 700 Seedlings † Saplings <0,5 m Grazed 600 Ungrazed 2 500 m

0 400 0

0 300 1

r

e 30 p NS

s l a

u 20 d i v

i NS d 10 n I NS NS 0 ) ) ) ) E H E y o H E S H E H E y o H E S H K S K ( ( L K h S K S K ( ( L K h S r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e O O B B O O B B

Carpinus betulus 250 Seedlings Saplings <0,5 m Saplings 0,5-2,0 m Grazed Ungrazed

2 200 NS m 150 0 0

0 100

1 NS

r 50 e p 30 s l

a NS u

d 20 i v i d

n 10 I † 0 ) ) ) ) ) ) E H E y o H E S H E H E y o H E S H E H E y o H E S H K S K ( ( L K h S K S K ( ( L K h S K S K ( ( L K h S r r k S S k h c h r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e A A O O e B e O O B B O O B B O O B B

Crataegus sp. 140 Seedlings Saplings <0,5 m Saplings 0,5-2,0 m Grazed NS Ungrazed 2 100 NS m 60 0 0

0 20 1

20 r e p 15 s l NS a u

d 10 i NS v i d

n 5 I NS NS NS NS NS 0 ) ) ) ) ) ) E H E y o H E S H E H E y o H E S H E H E y o H E S H K S K ( ( L K h S K S K ( ( L K h S K S K ( ( L K h S r r k S S k h c h r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e A A O O e B e O O B B O O B B O O B B

99

Euonymus europaeus 80 Seedlings Saplings <0,5 m Saplings 0,5-2,0 m Grazed Ungrazed

2 60 m 40 NS 0 0

0 20 1

20 r e p 15 s l NS a u

d 10 i v i NS

d NS

n 5 I NS 0 ) ) ) ) ) ) E H E y o H E S H E H E y o H E S H E H E y o H E S H K S K ( ( L K h S K S K ( ( L K h S K S K ( ( L K h S r r k S S k h c h r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e A A O O e B e O O B B O O B B O O B B

Fagus sylvatica 1500 Grazed Seedlings Saplings <0,5 m * Saplings 0,5-2,0 m Ungrazed 1000 NS 2 NS

m 500 **

0 100 NS 0

0 * 1 80 NS

r

e 60 NS p

NS s l 40 a u

d 20 i v i

d 6 NS NS

n NS I 4 NS NS † 2 NS 0 ) ) ) ) ) ) E H E y o H E S H E H E y o H E S H E H E y o H E S H K S K ( ( L K h S K S K ( ( L K h S K S K ( ( L K h S r r k S S k h c h r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e A A O O e B e O O B B O O B B O O B B

Frangula alnus 5 Seedlings Saplings <0,5 m Grazed Ungrazed 2 m

4 0 0 0 1

r 3 e p

s l

a 2 u d i v i NS d 1 n I

0 ) ) ) ) E H E y o H E S H E H E y o H E S H K S K ( ( L K h S K S K ( ( L K h S r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e O O B B O O B B

100

Fraxinus excelsior 30 Seedlings NS Saplings <0,5 m Grazed Ungrazed 2

m 25

0 0

0 20 1

r e p

15 s l a

u 10 d i v i d

n 5 I NS NS 0 ) ) ) ) E H E y o H E S H E H E y o H E S H K S K ( ( L K h S K S K ( ( L K h S r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e O O B B O O B B

Larix sp.

80 Seedlings Saplings <0,5 m NS Saplings 0,5-2,0 m Grazed 60 Ungrazed 2 m

40 NS 0

0 20 0 NS 1

10 r

e NS

p 8

s l

a 6 NS u d i v i 4 NS d NS

n NS I 2 NS NS 0 ) ) ) ) ) ) E H E y o H E S H E H E y o H E S H E H E y o H E S H K S K ( ( L K h S K S K ( ( L K h S K S K ( ( L K h S r r k S S k h c h r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e A A O O e B e O O B B O O B B O O B B

Lonicera periclymenum NS Seedlings Saplings <0,5 m Saplings 0,5-2,0 m Grazed 150 Ungrazed 2

m NS

0 100 0

0 *** 1

r 50 e p

s

l 40

a NS u 30 d i v i 20 d n I 10 0 ) ) ) ) ) ) E H E y o H E S H E H E y o H E S H E H E y o H E S H K S K ( ( L K h S K S K ( ( L K h S K S K ( ( L K h S r r k S S k h c h r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e A A O O e B e O O B B O O B B O O B B

101

Picea abies 7000 Seedlings NS Saplings <0,5 m Saplings 0,5-2,0 m Grazed 5000 Ungrazed 2

m 3000

0 1000 * 0 0 1

500 NS r 400

e NS NS p

300 s l 200 * a 100 u

d NS i 50 v i 40 d

n 30 I 20 NS † NS NS 10 NS NS NS NS 0 ) ) ) ) ) ) E H E y o H E S H E H E y o H E S H E H E y o H E S H K S K ( ( L K h S K S K ( ( L K h S K S K ( ( L K h S r r k S S k h c h r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e A A O O e B e O O B B O O B B O O B B

Pinus sp. 5 Seedlings Saplings <0,5 m Grazed Ungrazed 2 m

4 0 0 0 1

r 3 e p

NS s l

a 2 u d i v i NS

d 1 n I NS 0 ) ) ) ) E H E y o H E S H E H E y o H E S H K S K ( ( L K h S K S K ( ( L K h S r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e O O B B O O B B

Prunusavium 5 Seedlings Saplings <0,5 m Grazed Ungrazed 2 m

4 0 0 0 1

NS

r 3 e p

NS s l NS

a 2 u d i v i NS NS

d 1 n I

0 ) ) ) ) E H E y o H E S H E H E y o H E S H K S K ( ( L K h S K S K ( ( L K h S r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e O O B B O O B B

102

Prunuscerasifera 5 Seedlings Saplings <0,5 m Grazed Ungrazed 2 m

4 NS 0 0

0 NS 1

r 3 e p

s l

a 2 u d i v i NS NS NS

d 1 n I

0 ) ) ) ) E H E y o H E S H E H E y o H E S H K S K ( ( L K h S K S K ( ( L K h S r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e O O B B O O B B

Prunus spinosa

60 Seedlings Saplings <0,5 mNS Saplings 0,5-2,0 m Grazed 50 Ungrazed 2 m

40 0

0 30 0 1

20 r e p

NS s l

a 10 u d i v i NS d 5 n I NS NS NS NS NS 0 ) ) ) ) ) ) E H E y o H E S H E H E y o H E S H E H E y o H E S H K S K ( ( L K h S K S K ( ( L K h S K S K ( ( L K h S r r k S S k h c h r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e A A O O e B e O O B B O O B B O O B B

Pseudotsugamenziesii 7 Seedlings Saplings <0,5 m Grazed Ungrazed 2 6 m NS 0

0 5 0 1

r 4 NS e p

s l 3 a NS NS u d i 2 v i

d NS n

I 1 0 ) ) ) ) E H E y o H E S H E H E y o H E S H K S K ( ( L K h S K S K ( ( L K h S r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e O O B B O O B B

103

Quercus robur NS 5000 Seedlings5000 Saplings <0,5 m Grazed 3000 Ungrazed

2 3000 NS m 1000 1000 0 NS 0 350 350 0 300 1

r 250

e 250

p 200

**

s 150 l 150

a * 100 NS u

d 50 i 50

v * i †

d 10

n 10 NS I NS NS NS NS 5 NS 5 0 0 ) ) ) ) E H E y o H E S H E H E y o H E S H K S K ( ( L K h S K S K ( ( L K h S r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e O O B B O O B B

Quercusrubra 15 Seedlings Saplings <0,5 m Grazed Ungrazed 2 m

12 0 0

0 NS 1

r 9 NS e p

s l

a 6 u d i v i

d 3 n I NS 0 ) ) ) ) E H E y o H E S H E H E y o H E S H K S K ( ( L K h S K S K ( ( L K h S r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e O O B B O O B B

Ribesspicatum 30 Seedlings Saplings <0,5 m Grazed Ungrazed 2

m 25

NS 0 0

0 20 1

r e p

15 s l a

u 10 d i v i

d NS

n 5 I

0 ) ) ) ) E H E y o H E S H E H E y o H E S H K S K ( ( L K h S K S K ( ( L K h S r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e O O B B O O B B

104

Rosa sp. 5 Seedlings Saplings <0,5 m Saplings 0,5-2,0 m Grazed Ungrazed 2 m

4 0 0 0 1

NS

r 3 e p

s l

a 2 u d i v i NS NS

d 1 n I 0 ) ) ) ) ) ) E H E y o H E S H E H E y o H E S H E H E y o H E S H K S K ( ( L K h S K S K ( ( L K h S K S K ( ( L K h S r r k S S k h c h r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e A A O O e B e O O B B O O B B O O B B

Salixaurita 5 Seedlings Saplings <0,5 m Grazed Ungrazed 2 m

4 0 0 0 1

r 3 e p

s l NS

a 2 u d i v i

d 1 n I

0 ) ) ) ) E H E y o H E S H E H E y o H E S H K S K ( ( L K h S K S K ( ( L K h S r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e O O B B O O B B

Salixcinerea 10 Seedlings Saplings <0,5 m Grazed Ungrazed 2 NS m

8 0 0 0 1

r 6 e p

s l

a 4 u d i v i

d 2 n I

0 ) ) ) ) E H E y o H E S H E H E y o H E S H K S K ( ( L K h S K S K ( ( L K h S r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e O O B B O O B B

105

Salixspp. 15 Seedlings Saplings <0,5 m Grazed Ungrazed 2

m † 12 0 0 0 1

r 9 e p

s l

a 6 u d i v i NS d 3 NS

n NS I NS 0 ) ) ) ) E H E y o H E S H E H E y o H E S H K S K ( ( L K h S K S K ( ( L K h S r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e O O B B O O B B

Sambucus nigra 10 Seedlings Saplings <0,5 m Saplings 0,5-2,0 m Grazed Ungrazed 2 m

8 0 0 0 1

r 6 e p

s l

a 4 u d i v i

d 2 n

I NS NS 0 ) ) ) ) ) ) E H E y o H E S H E H E y o H E S H E H E y o H E S H K S K ( ( L K h S K S K ( ( L K h S K S K ( ( L K h S r r k S S k h c h r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e A A O O e B e O O B B O O B B O O B B

Sorbus aucuparia 100 Seedlings Saplings <0,5 m Saplings 0,5-2,0 m Grazed Ungrazed

2 60 * m

0 20

0 20 0

1 NS

r

e 15 p NS

s l a

u 10 NS NS d

i NS v

i NS d †

n 5 I NS NS NSNS NS NS NS 0 ) ) ) ) ) ) E H E y o H E S H E H E y o H E S H E H E y o H E S H K S K ( ( L K h S K S K ( ( L K h S K S K ( ( L K h S r r k S S k h c h r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e A A O O e B e O O B B O O B B O O B B

106

Ulmusglabra 5 Seedlings Saplings <0,5 m Grazed Ungrazed 2 m

4 0 0 0 1

r 3 e p

s l

a 2 u d i v i NS NS d 1 n I

0 ) ) ) ) E H E y o H E S H E H E y o H E S H K S K ( ( L K h S K S K ( ( L K h S r r k S S k h c h r r k S S k h c h e e a k k a c e c e e a k k a c e c ld ld a a e e e ld ld a a e e e A A O O e B e A A O O e B e O O B B O O B B

107

Appendix 3 – Trees >2 m The distribution in dbh-categories in cm of woody species with a height >2 m.

Alder swamp – Kollerup Enghave

10 25 25 a a a Alnus glutinosa Alnus incana Crataegus sp. h h h

1 1 1 , , , 0 0 0

8 20 50 20 r r r Grazed e e e p p p

s s s Ungrazed l l l a a a 6 15 15 u u u d d d i i i v v v i i i d d d 4 10 40 10 ** n n n s i i i

f f f ) s o o o

r r r p e e e e 2 5 5 b b b n m m m m u u u h a N N N 0 0 30 0 c i

10 30 50 70 90 +100 10 30 50 70 90 +100 w 10 30 50 70 90 +100 r

s

DBH (cm) DBH (cm) DBH (cm) s r e e i 20 d c l e A p ( Alder swamp – Store Hessemose S 10

50 50 5 5 a a Alnus incana Alnus incana a Crataegus sp. a Crataegus sp. h h h 0 h

1 1 1 1 , , , , 0 0 0 0

40 40 4 4 r r r e e r e e e s v e p p p p

a

s s o s s l l l l

a a h 30 30 a 3 m a 3 u u u g u d d d d

i i e i n i v v v v i i i s i d d 20 20 d 2 s E d 2 n n n n i i i i

e

f f p f f o o o o

u H r r r r r e e 10 10 e 1 e e 1 b b r b ll b m m m m

u u o o t u u N N 0 0 S N 0 K N 0 10 30 50 70 90 +100 10 30 50 70 90 +100 10 30 50 70 90 +100 10 30 50 70 90 +100 DBH (cm) DBH (cm) DBH (cm) DBH (cm)

5 a Betula pubescens h

1 , 0

4 r e p

s l

a 3 u d i v i

d 2 n i

f o

r

e 1 b m u

N 0 10 30 50 70 90 +100 DBH (cm)

Oak stand – Kollerup Enghave

10 50 5 a Quercus robur a Fagus sylvatica a Crataegus sp. h h h

1 1 1 , , , 0 0 0

8 40 4 r r r e e e p p p

s s s l l l a a 6 30 a 3 u u u d d d i i i v v v i i i d d 4 20 d 2 n n n i i i

f f f o o o

r r r e e 2 10 e 1 b b b m m m u u u N N 0 0 N 0 10 30 50 70 90 +100 10 30 50 70 90 +100 10 30 50 70 90 +100 DBH (cm) DBH (cm) DBH (cm)

108

Oak stand – Sandskredssøen (young)

25 25 10 10 a a a Quercus robur Quercus robur Picea abies a Picea abies h h h h

1 1 1 1 , , , , 0 0 0 0

20 20 8 8 r r r r e e e e p p p p

s s s s l l l l a a a 15 15 6 a 6 u u u u d d d d i i i i v v v v i i i i d d d 10 10 4 d 4 n n n n i i i i

f f f f o o o o

r r r r e e e 5 5 2 e 2 b b b b m m m m u u u u N N N 0 0 0 N 0 10 30 50 70 90 +100 10 30 50 70 90 +100 10 30 50 70 90 +100 10 30 50 70 90 +100 DBH (cm) DBH (cm) DBH (cm) DBH (cm)

5 5 5 5 a Betula pendula a Abies sp. a Quercus rubra a Fagus sylvatica h h h h

1 1 1 1 , , , , 0 0 0 0

4 4 4 4 r r r r e e e e p p p p

s s s s l l l l a a a 3 3 3 a 3 u u u u d d d d i i i i v v v v i i i i d d d 2 2 2 d 2 n n n n i i i i

f f f f o o o o

r r r r e e e 1 1 1 e 1 b b b b m m m m u u u u N N N 0 0 0 N 0 10 30 50 70 90 +100 10 30 50 70 90 +100 10 30 50 70 90 +100 10 30 50 70 90 +100 DBH (cm) DBH (cm) DBH (cm) DBH (cm)

Oak stand – Sandskredssøen (old)

a 5 a 5 5 Quercus robur Quercus robur a Fagus sylvatica h h h

1 1 1 , , , 0 0 0

4 4 4 r r r e e e p p p

s s s l l l a a 3 3 a 3 u u u d d d i i i v v v i i i d d 2 2 d 2 n n n i i i

f f f o o o

r r r e e 1 1 e 1 b b b m m m u u u N N 0 0 N 0 10 30 50 70 90 +100 10 30 50 70 90 +100 10 30 50 70 90 +100 DBH (cm) DBH (cm) DBH (cm)

Oak stand – Lille Hessemose

25 25 a a 5 5 Quercus robur Quercus robur a Fagus sylvatica a Fagus sylvatica h h h h

1 1 1 1 , , , , 0 0 0 0

20 20

r r 4 4 r r e e e e p p p p

s s s s l l l l a a 15 15 a 3 a 3 u u u u d d d d i i i i v v v v i i i i d d 10 10 d 2 d 2 n n n n i i i i

f f f f o o o o

r r r r e e 5 5 e 1 e 1 b b b b m m m m u u u u N N 0 0 N 0 N 0 10 30 50 70 90 +100 10 30 50 70 90 +100 10 30 50 70 90 +100 10 30 50 70 90 +100 DBH (cm) DBH (cm) DBH (cm) DBH (cm)

109

5 5 25 5 a a Alnus incana a Alnus incana Crataegus sp. a Prunus spinosa h h h h

1 1 1 1 , , , , 0 0 0 0

4 4 20 4 r r r r e e e e p p p p

s s s s l l l l a a a 3 3 a 15 3 u u u u d d d d i i i i v v v v i i i i d d d 2 2 d 10 2 n n n n i i i i

f f f f o o o o

r r r r e e e 1 1 e 5 1 b b b b m m m m u u u u N N 0 N 0 0 N 0 10 30 50 70 90 +100 10 30 50 70 90 +100 10 30 50 70 90 +100 10 30 50 70 90 +100 DBH (cm) DBH (cm) DBH (cm) DBH (cm)

Beech stand – Kollerup Enghave

50 50 a Fagus sylvatica a Fagus sylvatica h h

1 1 , , 0 0

40 40 r r e e p p

s s l l

a 30 a 30 u u d d i i v v i i

d 20 d 20 n n i i

f f o o

r r

e 10 e 10 b b m m u u

N 0 N 0 10 30 50 70 90 +100 10 30 50 70 90 +100 DBH (cm) DBH (cm)

Beech stand – Sandskredssøen

25 25 5 a a Fagus sylvatica Fagus sylvatica a Picea abies h h h

1 1 1 , , , 0 0 0

20 20 4 r r r e e e p p p

s s s l l l a a 15 15 a 3 u u u d d d i i i v v v i i i d d 10 10 d 2 n n n i i i

f f f o o o

r r r e e 5 5 e 1 b b b m m m u u u N N 0 0 N 0 10 30 50 70 90 +100 10 30 50 70 90 +100 10 30 50 70 90 +100 DBH (cm) DBH (cm) DBH (cm)

Beech stand – Store Hessemose

25 25 5 5 a a Fagus sylvatica Fagus sylvatica a Quercus robur a Quercus robur h h h h

1 1 1 1 , , , , 0 0 0 0

20 20 4 4 r r r r e e e e p p p p

s s s s l l l l a a 15 15 a 3 a 3 u u u u d d d d i i i i v v v v i i i i d d 10 10 d 2 d 2 n n n n i i i i

f f f f o o o o

r r r r e e 5 5 e 1 e 1 b b b b m m m m u u u u N N 0 0 N 0 N 0 10 30 50 70 90 +100 10 30 50 70 90 +100 10 30 50 70 90 +100 10 30 50 70 90 +100 DBH (cm) DBH (cm) DBH (cm) DBH (cm)

110

5 5 5 a Prunus spinosa a Corylus avellana a Picea abies h h h

1 1 1 , , , 0 0 0

4 4 4 r r r e e e p p p s s s l l l

a 3 a 3 a 3 u u u d d d i i i v v v i i i d 2 d 2 d 2 n n n i i i f f f o o o r r r

e 1 e 1 e 1 b b b m m m u u u

N 0 N 0 N 0 10 30 50 70 90 +100 10 30 50 70 90 +100 10 30 50 70 90 +100 DBH (cm) DBH (cm) DBH (cm)

111

Appendix 4 – Deadwood

Average amount of deadwood categorized in lying and standing and decay classes.

Deadwood standing (m3 ha-1) Deadwood lying (m3 ha-1) Total 1 2 3 4 5 Total 1 2 3 4 5 Total Alder KE 1.56 0.00 5.15 1.39 0.00 8.10 0.00 0.00 11.11 17.79 3.76 32.66 40.77 Grazed Alder SH 0.00 0.00 6.72 3.04 0.00 9.76 7.10 11.44 98.84 38.88 12.85 169.11 178.87 Grazed Alder SH 0.00 7.49 10.09 0.00 0.00 17.57 40.07 30.92 26.63 27.49 1.57 126.68 144.26 Ungrazed Oak KE 0.00 0.00 0.00 0.00 0.00 0.00 0.00 7.92 0.00 0.00 0.00 7.92 7.92 Grazed Oak S(y) 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 Grazed Oak S(Y) 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 Ungrazed Oak S(O) 0.00 0.00 0.00 0.00 0.00 0.00 1.31 32.73 0.00 0.00 0.00 34.04 34.04 Grazed Oak S(O) 0.00 0.00 0.00 0.00 0.00 0.00 5.87 23.32 1.63 0.00 0.00 30.82 30.82 Ungrazed Oak LH 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 1.48 0.00 1.48 1.48 Grazed Oak LH 0.00 3.83 3.00 0.00 0.00 6.83 0.00 0.00 1.76 0.00 0.00 1.76 8.59 Ungrazed Beech KE 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 7.59 3.36 10.95 10.95 Grazed Beech KE 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 13.13 6.62 0.00 19.75 19.75 Ungrazed Beech S 0.00 0.00 0.00 0.00 0.00 0.00 0.00 2.53 4.96 0.00 0.00 7.49 7.49 Grazed Beech S 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 Ungrazed Beech SH 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 Grazed Beech SH 0.00 0.00 0.00 0.00 0.00 0.00 0.00 7.50 0.00 0.00 0.00 7.50 7.50 Ungrazed

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