Deadwood enrichment combining integrative and segregative conservation elements enhances biodiversity of multiple taxa in managed forests

Biological Conservation 228 (2018) 70–78

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Deadwood enrichment combining integrative and segregative conservation elements enhances biodiversity of multiple taxa in managed forests

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Inken Doerflera, , Martin M. Gossnera,b, Jörg Müllerc,d, Sebastian Seibolda, Wolfgang W. Weissera a Terrestrial Ecology Research Group, Department of Ecology and Ecosystem Management, School of Life Sciences Weihenstephan, Technische Universität München, HansCarl-von-Carlowitz-Platz 2, 85354 Freising, Germany b Forest Entomology, Swiss Federal Research Institute WSL, Zürcherstrasse 111, CH-8903 Birmensdorf, Switzerland c Bavarian Forest National Park, Zoology, Department of Conservation and Research, Freyunger Str. 2, 94481 Grafenau, Germany d Field Station Fabrikschleichach, Biozentrum University of Würzburg, Glashüttenstraße 5, 96181 Rauhenebrach, Germany

A R T I C LE I N FO

A B S T R A C T

Keywords: Beech forest Canopy cover Deadwood Integrative conservation Multidiversity Saproxylic species

Integrative management strategies that simultaneously aim for wood production and biodiversity conservation are considered crucial to protect biodiversity of forest species outside protected areas. In this study, we evaluated whether deadwood enrichment as an integrative strategy at a scale of 17,000 ha resulted in enhanced biodiversity of saproxylic and non-saproxylic taxa eight years after the implementation of the strategy. The strategy included active deadwood enrichment with harvest remnants, retention of deadwood, and nature forest reserves areas. The analysis was based on data on the occurrence of plants, fungi, beetles, true bugs and birds from directly before and after the implementation of the strategy. The implementation of the strategy resulted in an increase in the deadwood amount by an average of 90 ± 40 m3 ha−1 (mean ± SE) over this period. While deadwood amounts doubled in production forests (+90%), they increased even more in nature forest reserves (+160%). Multidiversity (species density of all taxa) increased with an increase in deadwood amount; this was a result of an increase in the multidiversity of saproxylic species as the non-saproxylic multidiversity did not respond. Among single taxon groups, fungal and beetle species density responded positively to the increase in deadwood amount, especially when only saproxylic species were analysed. Importantly, this effect was not only found in the nature forest reserves, but also in the production forests. We thus conclude that active deadwood enrichment in production forests and nature forest reserves is a promising tool to rapidly promote the protection of forest biodiversity.

1. Introduction Millennia of human activity have affected temperate forests in Europe and altered their structure considerably to optimize forests for timber production (Parviainen, 2005). During and after the Middle Ages the forest cover decreased strongly in Germany. The introduction of a sustainable silviculture in the middle of the 18th century stopped the deforestation (Röhrig et al., 2006). As part of this management, the extent of forests has increased but natural broadleaf tree species were largely replaced by conifers to promote timber production (Spielmann et al., 2013). The structure of the remaining broadleaf forests has been strongly altered and particularly the amounts of deadwood were reduced to, on average, < 10% of natural amounts (Müller and Bütler, 2010). Additionally, forests have become denser and darker due to an increasing growing stock (Schelhaas et al., 2003). Owing to these structural changes, managed forests across Europe harbour, in general, ⁎

a lower biodiversity, especially of deadwood dependent species, compared to old-growth forests, or forests set aside for conservation, but with differences among taxa (Chumak et al., 2005; Paillet et al., 2010). The reduced availability of certain habitats, such as large-diameter deadwood and open forest habitats, has caused a decline in populations of species associated with these habitats and thus many of them are today classified as threatened (Berg et al., 1994; Nieto and Alexander, 2010; Seibold et al., 2015b). One strategy to protect biodiversity of forest species is the establishment of forest reserves without timber production that develop according to natural forest dynamics (Bollmann and Braunisch, 2013). Such reserves can benefit a variety of species (Paillet et al., 2010), particularly species requiring high amounts of deadwood (Bässler and Müller, 2010). The protection of species exclusively in reserves has, however, been questioned because protected areas comprise only a small fraction of the overall forest area, and are usually small and

Corresponding author. E-mail address: inkendoerfl[email protected] (I. Doerfler).

https://doi.org/10.1016/j.biocon.2018.10.013 Received 19 November 2017; Received in revised form 27 September 2018; Accepted 8 October 2018 0006-3207/ © 2018 Published by Elsevier Ltd.

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might increase with deadwood enrichment combined with sun exposure (Brazee et al., 2014). However, sun exposure can also have negative effects, e.g., on species numbers of saproxylic fungi (Bässler et al., 2010). Since management can influence the size and dynamic of canopy gaps (Boncina, 2000), it is important to consider potential effects of the canopy openness when evaluating deadwood enrichment. Existing evaluations of deadwood enrichment strategies in temperate European forests are mostly based on experimental treatments (Seibold et al., 2015a). The drawback of experiments is that they are often implemented on a small spatial scale and that their documentation often focuses on single taxonomic groups. The evaluation of the success of deadwood enrichment for increasing biodiversity on a landscape scale is still lacking. In this paper, we evaluate the effects of an integrative forest management strategy, applied by the Bavarian State Forest Company (BaySF, 2016), that aims at increasing deadwood amounts to support biodiversity. The strategy is implemented by active deadwood enrichment with harvest remnants and the passive retention of deadwood combined with nature forest reserves as segregative element. Based on an assessment of deadwood enrichment before and after the implementation of the strategy (Doerfler et al., 2017), we measured changes in diversity of five taxonomic groups, i.e. plants, true bugs, birds, beetles and fungi, that span a range of trophic levels and differ in their association with deadwood. These measurements were taken on 68 plots directly before and eight years after the implementation of the strategy. We tested the following hypotheses:

embedded in a matrix of production forests or non-forest habitats and are thus poorly connected (Abrego et al., 2015). Additionally, the development of old-growth features is a slow process if reserves are established in former production forests (Christensen et al., 2005; Paillet et al., 2015). The protection of species that depend on features such as deadwood may thus not be guaranteed in these reserves as populations may disappear before the required habitat structures are established by natural processes (Gossner et al., 2014; Müller et al., 2005). In contrast, with active deadwood management the deadwood amounts can increase even over short time periods (Doerfler et al., 2017). Therefore, conservation need not be restricted to forest reserves but should also include production forests (Gossner et al., 2013a; Seibold et al., 2017). Integrative forest management strategies that include both biodiversity conservation and timber production aim to create and retain structures such as deadwood or canopy gaps known to benefit biodiversity (e.g. Bauhus et al., 2009; Fedrowitz et al., 2014). When pursued in large parts of the production forest matrix, these measures are thought to increase total habitat amount and thus population sizes of many species (Mason and Zapponi, 2016; Seibold et al., 2017). In temperate European production forests, one of the most important components of such integrative strategies is the creation and retention of deadwood during regular management operations. In Germany, this strategy is increasingly applied and supported by governmental incentives. For example, the Bavarian State Forest Company, which manages 30% of the Bavarian forest (Bundeswaldinventur, 2012), has successfully applied this strategy in the northern Steigerwald forest since 2006. This resulted in a significant increase in deadwood amounts at the landscape scale (Doerfler et al., 2017). This particular strategy includes the intentional retention of tree crowns and parts of trunks after harvests, and preserving existing, naturally developed deadwood. These measures are complemented by the protection of small reserves (30–180 ha) that were created in the 1980s where deadwood naturally accumulates. The active creation of deadwood has been shown to promote the diversity of saproxylic, i.e., obligatorily dead-wood dependent, taxa (Gossner et al., 2013b; Jonsell et al., 2004; Lassauce et al., 2011; Seibold et al., 2015a) and it is similarly effective in stands with previously low and high deadwood amounts, at least for saproxylic beetles (Seibold et al., 2017). The creation of deadwood, however, may also affect non-target organisms, i.e. taxa not obligatorily dependent on deadwood. Non-saproxylic taxa have been shown to respond positively but also negatively to experimental deadwood enrichment (Seibold et al., 2015a). For example, birds have been repeatedly shown to profit from an active creation of standing deadwood stems (snags) that are used as breeding sites or for foraging (e.g. Brandeis et al., 2002; Kroll et al., 2012). The effect of deadwood enrichment on plants is rarely studied. However, it was shown that for single species, such as Gymnocarpium dryopteris (L.) Newman or Lycopodium annotinum (L.), deadwood can be an important habitat, especially in old growth mountain forests (Dittrich et al., 2014) as deadwood for example facilitates the growth of tree seedlings (Zielonka and Niklasson, 2001). Some species groups however profit from deadwood removal during post-disturbance salvage logging (Thorn et al., 2018). These are commonly associated with open habitats including a number of spiders and carabid beetles (Thorn et al., 2018). Therefore, the response of species to deadwood development in forests can also be linked to changes in abiotic conditions mediated by changes in the canopy, because the death of a tree usually leaves a gap in the canopy. Non-saproxylic taxa, such as plants, are positively influenced by creation of canopy gaps because of increased light availability (Beudert et al., 2015). For saproxylic taxa, increased light availability may also have an effect as it increases the exposure of deadwood to sunlight. Sun-exposed deadwood promotes, for instance, a higher diversity of saproxylic beetles, as well as different beetle communities than deadwood under shady conditions (Gossner et al., 2016; Lindhe et al., 2005; Seibold et al., 2016). Additionally, fungal richness

(i) Deadwood enrichment promotes overall biodiversity (ii) Responses are most pronounced in saproxylic taxa, but neutral or even negative for non-saproxylic taxa (iii) Canopy openness promotes saproxylic and non-saproxylic biodiversity. 2. Methods 2.1. Study area and conservation strategy The study area is part of the northern ‘Steigerwald’ in southern Germany (Bavaria) (N 49° 50′ 53 E 10° 29′ 41). Elevation ranges from 300 to 450 m, mean annual temperature is 8.2 °C and mean annual rainfall is about 850 mm (Lischeid, 2001). The study was conducted in the forestry department Ebrach (17,000 ha), which belongs to the Bavarian State Forest Company. The managed area covers 16,494.2 ha (97% of the total forestry department). The remaining area (3%) is designated as six forest nature reserves, in which all management activities have stopped. The two first reserves were established in 1978, followed by two more reserves in both 1998 and 2010. Today, the sizes of the reserves range from 28 to 183.4 ha, due to extensions of the areas of the oldest reserves. The main harvesting practice in this department is shelterwood cutting and single tree harvests. In 2006, the forest management was reformed with two major components: abandonment of clear-felling and a focus on tree species native to Germany: European Beech (Fagus sylvatica L.), Sessile oak (Quercus petraea (Mattuschka) Liebl.), Scots pine (Pinus sylvestris L.) and Norway spruce (Picea abies (L.) H.Karst.). Additionally, a conservation strategy was implemented that aims to increase deadwood amount to 20 m3 ha−1 in broadleaf stands older than 100 years, and 40 m3 ha−1 in broadleaf stands older than 140 years (Doerfler et al., 2017). These goals are pursued by an active enrichment via retention of harvest remnants including logs, snags, branches and stumps (without the definition of a minimum diameter) and by passive enrichment, i.e., naturally developed deadwood, such as snags or windblown trees, are retained. For the active enrichment, a strong focus is on the preservation of crowns as harvest remnants, wherefore felled stems are cut at the first strong branch (~ > 15 cm diameter) and the upper part, including the stem and branches remains in the forest. Additionally, the lower part of the stem remains in the 71

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both years. If the top layer was differentiated in over- and understory in one plot, we summed the cover of the two layers for a plot, which can result in > 100% cover for one assessment.

forest if this is either rotten or crooked (Appendix 1, Fig. A1-1). 2.2. Study design and assessment of structural characteristics

2.3. Biodiversity assessment

We conducted the study on 68 plots in beech forests that are permanently marked by a magnet in the centre and were surveyed for biodiversity in 2004 (Müller, 2005). Forty-four plots were located in managed areas and 24 in forest nature reserves. All plots were dominated by beech, with a minor admixture of other tree species, displaying the goal of the silvicultural reformation. Hence, they are not affected by the abandonment of clear-felling or deliberate changes in the tree species composition but are part of the deadwood enrichment strategy. The plots in the forest nature reserves were located in three areas. One area with eight plots (Brunnstube) was established in 1978 with a size of 5 ha and was extended in 1997 to 48 ha. Three plots were located within the core area and five plots in the extension area. The core area in this reserve was strongly affected by a wind throw in 2011, partly affecting two of our plots. The second reserve (Waldhaus), with 12 plots, was established in 1978 with a size of 10 ha and extended to 90.7 ha in 1998. Eight plots were located in the core area and four plots in the extension area. A third reserve with 4 plots (Kleinengelein) was established in 2010 with 53.7 ha. The plot set is one less than the set studied by Müller (2005), due to an inconsistency in the deadwood inventory in one plot in the third reserve. The main tree species is beech, accounting for an average of 80% of the basal area. Oak and pine are the main admixed tree species with an average of 0.05 and 0.04% basal area, respectively, followed by spruce, contributing about 0.01% of the basal area. The age structure on plots in the production forests ranged from very young (10 years) to over-mature stands (210 years). The stands in the forest nature reserves where the plots were located had a mean age of 150 years that included old beech trees > 300 years old (Müller, 2005). A deadwood inventory was conducted in 2004 (Müller, 2005), i.e. two years before the implementation of the strategy. The second inventory was carried out in 2014, i.e. eight years after the implementation of the strategy (Doerfler et al., 2017). A detailed description of how changes in deadwood amounts were assessed is given in Doerfler et al. (2017) and Appendix 2. Briefly, the inventory plots were circles with an area of 1000 m2. The minimum diameter for measured objects was 12 cm. The diameter of logs (branches or parts of stems) and stumps was measured in the middle of the object, taking the average of two measurements 90° to each other. For upright deadwood objects > 1.3 m in height, we measured the diameter at breast height (DBH, 1.3 m). For complete lying trees, we measured the diameter at 1.3 m distance from the root plate (Fig. A2-1). The amount of piled and scattered fine woody debris (FWD, diameter < 12 cm) was not measured but recorded as the percentage cover on 1000 m2 plots. The calculation of deadwood amount was done separately for each class of deadwood objects (Doerfler et al., 2017). For logs, broken snags, stumps and crowns we used the formula for the volume of a cylinder, while for snags and complete lying trees we used the formula for living trees based on diametrical quotients following Kennel (1973) (Appendix 2). For scattered FWD we calculated the ground area covered by wood and calculated the volume by using the formula of a cylinder with a standardized height of a 1.5 cm, which is the mean height of scattered FWD in 2004. For data analysis, all volumes measured per object were added and standardized to m3 per hectare (Appendix 2). We included changes in canopy openness, as a surrogate for mortality by harvest or natural causes (wind, ageing) and light availability in our study. This was defined as the change in canopy cover of the top tree layer (> 10 m height) which was assessed on a 200 m2 square that had the same centre as the inventory plot, in 2004 and in 2014. The assessment was done visually during the assessment of vascular plants, after an agreement on the estimates between the people that assessed in

We examined the species densities, i.e. numbers of species per unit area, of five taxonomic groups in 2004 and 2014. The methods of the biodiversity assessment differed depending on the taxonomic group but were centred in the deadwood inventory plots for all biodiversity assessments (Appendix 3, Fig. A3-1). For sampling of flying beetles and true bugs, we used flight-interception traps. Traps consisted of two transparent plastic windows (40 cm × 60 cm) with a funnel and a jar filled with copper sulphate attached to the bottom, and a roof attached to the top (Appendix 4, Fig. A4-1). The traps were installed at the centre of each plot at 1.5 m height using a rope that was hung over large off-standing branches of trees, between two trees or, if the stand was too open, on a wooden construction. Traps were operated from March to October, with sampling vials replaced once a month. Species were determined by different experts in the two years (beetles: Ludger Schmidt, Johannes Bail). Trapping was supplemented by time-standardized (45 min per plot) hand-collection of beetles by an expert (Heinz Bußler) in both years, with a special focus on deadwood habitats to also represent less mobile species. The sampling was conducted three times per year (April/June, July/August, September) in both years and on 1000 m2 circular plots. For analysis, data sampled with both methods were combined into one dataset. We identified all specimens to species level in 2014 and excluded all individuals from families that were not identified to species level in 2004 from the analysis. Fungi were assessed by two different experts in the two years (Heinz Engel, Markus Blaschke) on circular plots covering an area of 1000 m2. We used a time-standardized method with 45 min spent within the plot area. During this time, the expert examined the ground and deadwood objects. Fruiting bodies of deadwood- and soil-inhabiting fungi that could be determined without microscopic analysis (macroscopic) were recorded. Sampling was conducted parallel to hand-collection of beetles. To reduce the effect of the different experts in the sampling years and hence increase the comparability of the datasets we applied a strict set of rules to include only species that experienced mycologists can examine, inspired by Baber et al. (2016). We therefore excluded from the analysis all species from our record that were irregularly fructifying, very rare, or very small and therefore easy to overlook (total 310 species). This was done to include as many species as possible but exclude species that could not be determined equally by the experts. The selection was based on expert knowledge with an agreement among experienced mycologists (see Acknowledgements). Birds were recorded within a 1 ha square five times in regular intervals (5–8 days) between March and June, using point-stop grid sampling (Rico Michaelis, Niclas Böhm, Jörg Müller, Volker Zahner, and Christine Franz). The surveyor recorded for a fixed time of 7 min, passing the grid along a central line (Appendix 4, Fig. A4-2) and mapping also during the walk. To record undisturbed activities the mapper spent 1 min at the two opposing edges and about 5 min in the centre, inclusive of walking time. All birds that could be seen (using binoculars) or heard in the plot were recorded. In cases of calls that did not allow a species identification, the surveyor used tape lures at a low volume to provoke further calls. Only definitely identified individuals were registered. The abilities of the surveyors were known from previous projects. The mapping was conducted during the morning (from dawn, i.e. between 5 and 7 am, until 11 am) and in the evening (5–7 pm), and only on days without rain or wind. To account for differences between surveys the order of plots was changed after each replicate. Vascular plants were recorded in the same 200 m2 square as canopy cover by trained students (Holger Hastreiter, Michael Seuß) and the 72

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first author. Due to low herb abundances, we did not set a standardized time for sampling. However, species that could not be identified in the field were sampled for identification. To avoid errors due to different people conducting the recording, we classified the cover of single species according to Braun-Blanquet (Wikum and Shanholtzer, 1978). This scale gives only rough estimates for high cover values but is more precise at the low end of the scale. We calculated this scale to percentages (Appendix 4, Table A4-1). We only recorded species of the herb layer (< 1.5 m height) for the analysis of species density. Two surveys were conducted (April and June). We separated the taxa into subgroups depending on their deadwood dependency (saproxylic, non-saproxylic). We used the definition from Stokland et al. (2012): “any species that depends, during some part of its life cycle, upon wounded or decaying woody material from living, weakened or dead trees”. Therefore, cavity breeding birds, all fungi fruiting on deadwood, and beetle species listed as saproxylic in Schmidl and Bußler (2004) were considered as saproxylic. All plants, soil-saprotrophic fungi and mycorrhiza fungi, non-cavity breeding birds and phytophagous beetles were considered as non-saproxylics in the analysis. Among the sampled true bugs only the genus Aradus can be considered saproxylic (Gossner et al., 2007; Heiss and Péricart, 2007). Species of this genus were included in the analysis of the total species density, but not analysed separately as saproxylics because they were rare (six individuals of two species).

2) the difference in percentage canopy cover between the years to account for mortality of trees (by natural causes and harvest) and possible changes in the light regime. To account for possible effects of species diversity in 2004 on the biodiversity in 2014, we used the multidiversity or the log-transformed species density from 2004 as offset variables in the analyses. We tested our models for spatial autocorrelation with the function spline.correlog (package ncf), with the randomization set = 1000. This function analyses the relationship between the model residuals and the geographic distance between plots (Appendix 6). In addition, we tested if the examined taxa had different species densities in forest nature reserves compared to production forests, using a GLM with a Gamma-error distribution and a log-link for multidiversity, and a Poisson-error distribution for species density in both sampling years (2004 and 2014). In this study, we aimed to evaluate the effects of a landscape-wide deadwood enrichment strategy that includes deadwood enrichment in reserves and production forests. Therefore, we included plots in both management types, but with an imbalanced data set (44 plots in production forests and 24 reserves). To give us the opportunity to discuss the effects of different types of enrichment on biodiversity, we repeated the analyses separately for forest nature reserves and production forests, to see if deadwood enrichment has different outcomes in these areas (for the model see Appendix 5). For these models, we used an interaction term that separates the samples into plots within and outside of reserves.

2.4. Statistical analyses All statistical analyses were conducted in R (R CoreTeam, 2015, version 3.3.3). For pairwise comparison of deadwood amounts within plots between 2004 and 2014, we used a paired t-test for normally distributed data or a paired Wilcoxon-rank test for non-normally distributed data. The distribution was tested beforehand, both visually and by using the Shapiro test. One major goal was to evaluate the effects of integrative forest management on overall biodiversity. We therefore calculated an index of multidiversity published by Allan et al. (2014) that combines the species richness (total number of species identified per plot) of all studied taxonomic groups while accounting for differences between total species numbers between groups. The index calculates the average proportional species richness over all the examined groups. This was done by scaling the species richness per group to the maximum value found in either study year across plots. Since we wanted to compare the diversities in 2004 and 2014, we scaled the species richness to the maximum that was found in one of the two years. We used the function without thresholds, since the use of thresholds would exclude very rare species. The function without thresholds calculates the mean of the scaled species numbers of the five taxonomic groups. To calculate the final multidiversity measure, we calculated an average of the scaled diversities with all groups weighted equally to avoid species rich groups determining the pattern. In addition, we calculated multidiversity of saproxylic and non-saproxylic species separately. Dependent variables were thus overall multidiversity, multidiversity of saproxylic and non-saproxylics, and species densities of each taxonomic group, also separated into saproxylics and non-saproxylics in 2014. These variables were analysed using generalized linear models (GLM; function glm, package stats). We used a Gaussian error term distribution for multidiversity, and a Poisson error term for the analysis of species density (Appendix 5). If the deviance was > 1.5 times larger than the degrees of freedom (indicating overdispersion), we used the quasi-Poisson error term. As predictor variables we used:

3. Results 3.1. Deadwood amounts and canopy cover Deadwood amounts increased in 53 of our 68 plots (Appendix 7, Fig. A7-1). On average, deadwood amount (including all types of measured objects) increased significantly by 112.5% from 39.2 ± 5.9 (mean ± SE) m3 ha−1 in 2004 to 122.5 ± 41.2 m3 ha−1 in 2014 (V = 494, p < 0.001). The increase differed between production forests and forest nature reserves (Appendix 7, Table A7-1). The amount in production forest increased by 90% from 18.8 ± 2.4 m3 ha−1 in 2004 to 54.6 ± 8.2 m3 ha−1 in 2014. The amount in forest nature reserves increased by 160% from 69.6 ± 12.4 m3 ha−1 in 2004 to 251.2 ± 114.3 m3 ha−1 in 2014. Canopy cover decreased on average by 15% from 83 ± 2.9% cover in 2004 to 69 ± 4.3% cover in 2014 (W = 3218.5, p < 0.0001, range = −90 to +85), with a higher number of plots (51) showing decreased cover than increased cover (17) (Appendix 8, Fig. A8-1). Deadwood enrichment (log ratio of deadwood amounts) was not significantly related to the change in canopy cover (Appendix 9, Table A91). Thus, both variables can be used as independent predictors in the subsequent analyses of diversity changes. 3.2. Species diversity In total, we found 810 species across all five taxonomic groups, 655 species before the implementation of the strategy, and 601 species after the implementation of the strategy (Appendix 10). The most species rich group was beetles, followed by fungi (Table 1). These two groups also had the highest percentage of saproxylic species (Table 1). The percentage of species occurring in both years was highest for fungi (influenced by our subset of the species) and smallest for true bugs (Table 1). Before the implementation of the strategy non-saproxylic beetles, saproxylic fungi and non-saproxylic true bugs had a higher species density in production forests, whereas birds were more common in reserves (Appendix 11, Table A11-1). After the implementation, only saproxylic fungi were more common in production forests (Appendix 11, Table A11-1).

1) the log-transformed ratio of the change in deadwood amount (the 2014 log scaled amount divided by the 2004 log scaled amount, without having to add a value of one since there are no plots without deadwood) to give small changes in deadwood amounts stronger weight, and. 73

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groups or subgroups responded to changes in canopy cover (Table 2, Appendix 12, Fig. A12-1).

Table 1 Species densities of the five examined species groups, the number of saproxylic species and the percentage overlap in all species between the two years.

Beetles Fungi Birds Plants True bugs

All species

Saproxylic species

2004

2014

2004

2014

340 117 47 94 57

283 126 40 90 62

280 75 19 0 2

237 80 16 0 2

% overlap between years, all species

3.4. Differences in biodiversity change between production forests and forest reserves

49 80 74 57 38

The effects of changes in deadwood amount on multidiversity and on single taxonomic groups did not differ between forest nature reserves and production forests (Appendix 13, Table A13-1). Thus, the positive effect of deadwood enrichment can be found independently in both forest reserves and production forests. Hence, the positive effect of the conservation strategy on biodiversity is not only due to the creation of forests reserves, but also to deadwood enrichment during harvests. The effects of changes in canopy cover on diversity did, however, differ between forest reserves and production forest plots, for fungi and birds. Whereas species density of fungi and therein especially saproxylics increased with increasing canopy cover only in reserves, the species density of birds increased with decreasing canopy cover, also only in nature reserves.

3.3. Drivers of biodiversity change Deadwood enrichment and multidiversity were positively correlated (Table 2, Fig. 1), indicating a positive effect of the implementation of the deadwood enrichment strategy on biodiversity. With increasing deadwood amounts, multidiversity of saproxylics increased, while multidiversity of non-saproxylics showed no change (Table 2, Fig. 1). Multidiversity was not affected by changes in canopy cover (Table 2, Fig. 1). Positive effects of deadwood enrichment on individual taxonomic groups were mostly driven by saproxylic beetles and fungi. For both taxonomic groups, all species and the saproxylic subgroups responded positively to increasing deadwood amounts (Fig. 2). Considering all species, the exponential of the coefficient of deadwood changes is 1.147 for beetles and 1.155 for fungi, so that a change of one cubic meter in deadwood per hectare produces approximately a change of 1.2 in species numbers per 0.1 to 0.5 ha – depending on the examined group (Table 2). For the saproxylic subgroups the coefficients are even higher, indicating a stronger increase with an increasing change in deadwood amounts (Table 2). Non-saproxylic beetles and fungi did not respond significantly to deadwood enrichment (Table 2, Fig. 2). Neither all birds, nor their saproxylic or non-saproxylic subgroups responded to deadwood enrichment (Table 2, Fig. 2). For plants and true bugs, we also found no significant relationships between overall species density or species density of the subgroups and deadwood enrichment (Table 2, Fig. 2). The decrease in canopy cover had a marginally positive effect on the species density of non-saproxylic birds. None of the other taxonomic

4. Discussion Our study shows that there is a positive log-linear relationship between deadwood enrichment applied at the scale of a forest management company, and an increase in biodiversity of saproxylic organisms. This supports the hypothesis that deadwood enrichment with an increase in the amount and diversity of deadwood habitats promotes biodiversity (Seibold et al., 2017). Importantly, we found this for both production forest plots and for nature reserves. Our study thus indicates that the combination of segregative elements (set aside areas), which are demonstrably important for many species (Bässler and Müller, 2010; Paillet et al., 2010), with integrative measures of deadwood enrichment and retention in production forests can lead to very high deadwood amounts and with that, an increase in biodiversity. Thus, our results demonstrate the feasibility of a landscape-wide design of these conservation measures. For the Bavarian State Forest Company this finding indicates a success of the measures implemented in the nature conservation strategy of the district Ebrach aimed at promoting and protecting forest biodiversity.

Table 2 Summary of a generalized linear model with the multidiversity and species numbers of the five examined taxa (beetles, fungi, birds, plants, bugs) separated into nonsaproxylic and saproxylic species as response variables; predictors are deadwood enrichment as log response ratio (log (2014 amount/2004 amount)) and the change in canopy cover (2014 cover - 2004 cover). For each model the exp (estimate), test statistic and p-value is given. The value behind the test statistic depends on the family term used in the model: Poisson (p) = z, quasi-Poisson (qp) = t, and Gaussian (g) = F. The family term that was used in the model is indicated in “error distribution family”. Variables in boldface are significant. Taxonomic groups that had no or few saproxylic species were not analysed separately. Species

Group

Change in:

Error distribution family

Canopy cover

Multidiversity Multidiversity Multidiversity Beetles Beetles Beetles Fungi Fungi Fungi Birds Birds Birds Plants True bugs True bugs

All Saproxylic Non-saproxylics All Saproxylic Non-saproxylics All Saproxylic Non-saproxylics All Saproxylic Non-saproxylics All All Non-saproxylics

Deadwood

Estimate

Statistic

p

Estimate

Statistic

p

1.001 1.001 1.000 1.001 1.001 1.000 1.002 1.002 1.001 0.999 1.001 0.998 0.997 0.999 0.999

1.39 1.57 −0.92 0.56 0.65 0.02 1.69 1.44 1.02 −0.94 0.49 −1.73 −1.22 −0.58 −0.55

0.17 0.12 0.36 0.58 0.52 0.99 0.10 0.16 0.31 0.35 0.63 0.08 0.23 0.56 0.58

1.058 1.065 1.006 1.147 1.189 1.010 1.155 1.230 0.991 1.012 1.015 1.011 1.071 1.046 1.040

3.33 5.66 0.78 3.63 4.00 0.32 4.64 6.04 −0.18 0.44 0.41 0.29 0.88 0.77 0.67

< 0.001 < 0.001 0.44 < 0.001 < 0.001 0.75 < 0.001 < 0.001 0.86 0.66 0.68 0.77 0.38 0.45 0.50

74

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and sites designated for production. Therefore, the approximately 19 m3 ha−1 of deadwood found in our study area before the implementation of the strategy might actually be higher than the average in production forests in Germany. The amount after the implementation of the strategy (54 m3 ha−1) is then distinctly above the average found in forests in Germany, in all the federal states (Appendix 7, Table A7-2). We found a major increase in deadwood amounts in reserves, which resulted on the one hand from an increase in mortality of old trees, which is a typical development for reserves (Christensen et al., 2005), but also from ‘catastrophic events’ such as windthrows. In natural forests of the temperate zone, ageing and small-scale catastrophes are supposed to be the main driver of deadwood amounts (Holzwarth et al., 2013), which are suppressed by forest management (Laarmann et al., 2009). Within the production forests, we found markedly lower amounts, though the relative increase in deadwood amounts was also strong in production forests. This shows that enrichment of deadwood is possible within production forests, but that the competition with wood extraction does not allow for enrichments comparable to those of ageing forests or windthrow sites in reserves. The higher amounts are caused by three major factors: (1) trees in the forest nature reserves have in general a higher volume, building up a higher amount of deadwood when they die, (2) trees in forest nature reserves remain as whole tree and (3) therefore the diameter of deadwood object is also larger (Appendix 7, Table A7-1). 4.2. Response of biodiversity to deadwood enrichment The positive relationship between biodiversity and deadwood amounts was significant in saproxylic species, particularly beetles and fungi. This was independent from the location from in production forests or nature forest reserves. Biodiversity change was assessed only eight years after the implementation of the strategy, and although there was considerable variation in absolute species density between the years, the increase in deadwood already resulted in an increase in biodiversity. The high differences within fungi and beetle species density between the years are presumably related to climatic differences between the years, which are shown to considerably influence species numbers (e.g. Preisler et al., 2012). Therefore, we cannot tell whether the decrease in gamma diversity will be a persistent effect and would recommend a long term observation of these effects. Cavity-breeding birds were the only saproxylic group that did not respond to deadwood enrichment in our study. This could be because our method might not capture the extent of the species' home ranges, which would make an analysis including landscape-level factors important (Paillet et al., 2010). Non-saproxylic species were neither positively nor negatively affected by deadwood enrichment over the time period considered in this paper. Therefore, our findings do not confirm the positive relationship between non-saproxylic species and deadwood found in experimental studies (Seibold et al., 2015a). However, other studies have shown that species numbers of non-saproxylic species can be negatively affected by deadwood increases, with some species profiting rather from deadwood removal (Thorn et al., 2018). We did not find such negative effects and may conclude that the way deadwood enrichment was carried out in the Steigerwald region does not negatively affect non-saproxylic species while promoting saproxylic species. It remains to be seen if a response from the non-saproxylic species will appear when this study is repeated over longer time periods. If not, additional actions should be considered. Our study show that the generally positive effects of deadwood enrichment on biodiversity found in experimental studies (Gossner et al., 2013a; Seibold et al., 2015a; Vanha-Majamaa et al., 2007) can be transferred to landscape level in the real world. This indicates the importance of deadwood enrichment through a combination of different methods as an integrative strategy. In particular, saproxylic multidiversity and species density of saproxylic beetles and fungi showed a

Fig. 1. The relationship between the fitted values of multidiversity and the change in deadwood amount (log response ratio, upper panel) and change in canopy cover (lower panel). Red circles and red line: multidiversity of total species numbers. Blue triangles and blue line: multidiversity of saproxylic species. Green square and green line: multidiversity of non-saproxylic species. Dashed lines display non-significant results. (For interpretation of the references to colour in this figure legend, the reader is referred to the web version of this article.)

4.1. Deadwood enrichment Active deadwood enrichment and passive retention in managed forests, in combination with assigning set aside areas, resulted in a strong increase in deadwood amounts in production forests as well as in reserves. The amounts of wood found in the study area before the implementation of the strategy were comparable to those in German forests, which are on average 20.6 m3 ha−1 (measurement with a minimum diameter 10 cm at the thicker end). However, deadwood amounts depend on the organization managing the forests. Deadwood amounts within the state forests owned by the federal states, especially in Bavaria (35.1 m3 ha−1), are higher than in private forests, communal forests or state forests owned by the German state (Bundeswaldinventur, 2012) (Appendix 7, Table A7-2). However, the plots from the national forest inventory in Germany cannot be separated into protected areas, such as nature reserves or national parks, 75

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Fig. 2. The relationship between fitted values of the species density of the five examined taxa (upper to lower row from left to right: beetles, fungi, birds, plants and bugs) and change in deadwood amount (log response ratio) and diversity. Blue triangles and blue line: multidiversity of saproxylic species. Green square and green line: multidiversity of non-saproxylic species. Dashed lines display non-significant results. (For interpretation of the references to colour in this figure legend, the reader is referred to the web version of this article.)

However, an integrative strategy that aims to increase the amount of deadwood will necessarily also increase deadwood diversity, making the strategy more likely to be successful in promoting biodiversity.

positive response to deadwood enrichment. This effect may be caused by higher resource availability, allowing rare species to coexist with dominant ones as proposed by (Fahrig and Triantis, 2013) (habitat amount hypothesis). However, deadwood amount is usually correlated to deadwood diversity (e.g. Müller and Bütler, 2010; Seibold et al., 2016). Higher deadwood diversity may provide a higher number of niches, and thus host more saproxylic species (habitat-heterogeneity hypothesis; Hutchinson (1959)). As deadwood amounts and deadwood diversity were highly correlated in our study (Appendix 6, Fig. A6-2), we could not disentangle which mechanism caused the increase of biodiversity.

4.3. Response of biodiversity to changes in the light regime Experimental studies that included light conditions found that several forest species are positively affected by an increase in light availability, e.g. saproxylic taxa such as beetles (Seibold et al., 2016) and fungi (Brazee et al., 2014). In old-growth forests where canopy gaps 76

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5. Conclusions and implications for nature conservation

develop due to senescent trees, plants, as non-saproxylic taxa (Burrascano et al., 2008), were also found to respond positively to canopy openings. However, our model including all plots could not confirm an overall positive relationship between canopy gaps and biodiversity. Compared to production forests, the relationship between deadwood amounts and canopy cover was stronger in nature reserves, and fungi as well as birds showed a significant response towards canopy gaps in reserves. Whereas birds were positively affected by a decrease in canopy cover in reserves, fungi were negatively affected. This suggests that the gaps that occurred on our study plots were different between production forests and reserves. In production forests, the gaps were probably too small to create an effect, which is also indicated by the weak relationship between deadwood amounts and canopy opening. There could be several reasons for the missing relationships: (1) the dominant harvesting methods in our study area were shelterwood cutting and single tree harvests, where only small gaps are created during harvests; (2) beech can close canopy gaps very quickly by expanding the crown (Zeibig et al., 2005); (3) most of the larger gaps we found in our study resulted from the harvest in the previous winter (2013). Therefore, gaps in production forests are probably small and short-lived, and the time period since creation of the gap and our study was probably too short for species such as plants to colonize the patches with high light availability. In the nature reserves, there was a significant effect of changes in canopy opening on biodiversity. This is probably due to the wider range of changes in light availability in the reserves due to wind throws. The plots in one reserve were located at the edge of a windthrow area that caused a large gap in 2011. Although a few trees remained on our plots, the deadwood amount and the amount of sun-exposed deadwood in this area was very high. In this study, we have focused on the effect of deadwood enrichment and included canopy gaps as a covariate in the analysis. Biodiversity in forests can be affected by a large number of additional factors, e.g. trees with specific microhabitats, hollows (Müller et al., 2014), overall climate (Bouget et al., 2014) or tree species diversity (Sobek et al., 2009). Also, deadwood amounts are influenced by the stand structure, tree species composition and many other variables, even after enrichment (Doerfler et al., 2017). Despite all these factors that might additionally influence biodiversity, also in our study, deadwood enrichment was a significant factor for the promotion of biodiversity.

Our study shows that deadwood enrichment on a landscape scale applied as a conservation strategy with integrative and segregative elements by a forestry department has a positive effect on saproxylic biodiversity and does not harm non-saproxylic species groups. This suggests that such a strategy can be an important tool for biodiversity conservation in forests. Furthermore, we saw that gap creation was independent of deadwood enrichment in production forests yet was important for birds in nature reserves. These findings suggest that gap creation to benefit light-demanding species cannot simply be achieved by deadwood enrichment during management. To create a mosaic of gaps with shaded and sun-exposed deadwood therefore requires separate management actions. Acknowledgements We thank Heinz Bußler and Markus Blaschke for the sampling of beetles and fungi. We are very grateful for the support by Markus Blaschke and Claus Bässler for their expert opinion on the fungi samples. Furthermore, we would like to express our gratitude to Rico Michaelis for his support with the bird survey, to Claudia Seilwinder and Moritz Pretzsch for their support with the deadwood inventory and Peter Biber for his support with the volume calculations. The project was financed by the Bayerisches Staatsministerium für Ernährung, Landwirtschaft und Forsten, grant L55. We thank the Bayerische Landesanstalt für Wald und Forstwirtschaft and the Bayerische Staatsforsten for their cooperation. In particular, we want to thank the forestry department Ebrach with its manager Ulrich Mergner and his whole team from the forest company Ebrach for their support. Appendix A. Supplementary data Supplementary data to this article can be found online at https:// doi.org/10.1016/j.biocon.2018.10.013. References Abrego, N., Bässler, C., Christensen, M., Heilmann-Clausen, J., 2015. Implications of reserve size and forest connectivity for the conservation of wood-inhabiting fungi in Europe. Biol. Conserv. 191, 469–477. Allan, E., Bossdorf, O., Dormann, C.F., Prati, D., Gossner, M.M., Tscharntke, T., Bluthgen, N., Bellach, M., Birkhofer, K., Boch, S., Bohm, S., Borschig, C., Chatzinotas, A., Christ, S., Daniel, R., Diekotter, T., Fischer, C., Friedl, T., Glaser, K., Hallmann, C., Hodac, L., Holzel, N., Jung, K., Klein, A.M., Klaus, V.H., Kleinebecker, T., Krauss, J., Lange, M., Morris, E.K., Muller, J., Nacke, H., Pasalic, E., Rillig, M.C., Rothenwohrer, C., Schall, P., Scherber, C., Schulze, W., Socher, S.A., Steckel, J., Steffan-Dewenter, I., Turke, M., Weiner, C.N., Werner, M., Westphal, C., Wolters, V., Wubet, T., Gockel, S., Gorke, M., Hemp, A., Renner, S.C., Schoning, I., Pfeiffer, S., Konig-Ries, B., Buscot, F., Linsenmair, K.E., Schulze, E.D., Weisser, W.W., Fischer, M., 2014. Interannual variation in land-use intensity enhances grassland multidiversity. Proc. Natl. Acad. Sci. U. S. A. 111, 308–313. Baber, K., Otto, P., Kahl, T., Gossner, M.M., Wirth, C., Gminder, A., Bässler, C., 2016. Disentangling the effects of forest-stand type and dead-wood origin of the early successional stage on the diversity of wood-inhabiting fungi. For. Ecol. Manag. 377, 161–169. Bässler, C., Müller, J., 2010. Importance of natural disturbance for recovery of the rare polypore Antrodiella citrinella Niemela & Ryvarden. Fungal Biol. 114, 129–133. Bässler, C., Müller, J., Dziock, F., Brandl, R., 2010. Microclimate and especially resource availability are more important than macroclimate for assemblages of wood-inhabiting fungi. J. Ecol. 98, 822–832. Bauhus, J., Puettmann, K., Messier, C., 2009. Silviculture for old-growth attributes. For. Ecol. Manag. 258, 525–537. BaySF, 2016. Forstbetrieb Ebrach. Bayerische Staatsforsten AöR. Berg, Å., Ehnström, B., Gustafsson, L., Hallingbäck, T., Jonsell, M., Weslien, J., 1994. Threatened plant, animal, and fungus species in Swedish forests: distribution and habitat associations. Conserv. Biol. 8, 718–731. Beudert, B., Bässler, C., Thorn, S., Noss, R., Schröder, B., Dieffenbach-Fries, H., Foullois, N., Müller, J., 2015. Bark beetles increase biodiversity while maintaining drinking water quality. Conserv. Lett. 8, 272–281. Bollmann, K., Braunisch, V., 2013. 1.1 To Integrate or to Segregate: Balancing Commodity Production and Biodiversity Conservation in European Forests. European Forest Institute, Freiburg. Boncina, A., 2000. Comparison of structure and biodiversity in the Rajhenav virgin forest

4.4. The financial aspect of a deadwood enrichment strategy The Bavarian State Forest Company faces a trade-off between the obligation to protect biodiversity and the need for profitable wood production. Integrative conservation strategies therefore need to be financially feasible. The cost calculation of this strategy, besides the income loss from set aside areas, might be best done by calculating the monetary value of deadwood. The wood that remains during active enrichment is of a low quality (crowns, discoloured stems, cut stumps) and would usually be sold as firewood to wood buyers who bear the cost of felling and cutting the wood themselves. The company estimates about 25 €/m3 (~30.9 $/m3 or ~22.3 £/m3) for deadwood of this quality, resulting in a loss of about 90 € of income per hectare and year (((54.6 m3 − 18.8 m3) / 10 years) × 25 €; Appendix 7, Table A7-1). However, this value might be overestimated, because the deadwood amounts per hectare include deadwood that cannot be sold, such as cut stumps or small diameter snags. We are therefore convinced that, for a fast restoration of deadwood habitats in production forests and with this a promotion of biodiversity and ecosystem services, the strategy evaluated in this paper is financially feasible. We found that clearly defined goals are very important for the implementation and the evaluation of the strategy, although the formulation could be more precise, because it is e.g. not clear if the targets of 20 and 40 m3 ha−1 are a mean or a minimum value (i.e. to be achieved in all plots), or which diameter classes of deadwoods, as well as sun exposure or tree species are to be enriched. 77

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