Ecosystem Management of the Boreal Forest
Summary and Keywords
Boreal countries are rich in forest resources, and for their area, they produce a disproportionally large share of the lumber, pulp, and paper bound for the global market. These countries have long-standing strong traditions in forestry education and institutions, as well as in timber-oriented forest management. However, global change, together with evolving societal values and demands, are challenging traditional forest management approaches. In particular, plantation-type management, where wood is harvested with short cutting cycles relative to the natural time span of stand development, has been criticized. Such management practices create landscapes composed of mosaics of young, even-aged, and structurally homogeneous stands, with scarcity of old trees and deadwood. In contrast, natural forest landscapes are characterized by the presence of old large trees, uneven-aged stand structures, abundant deadwood, and high overall structural diversity. The differences between managed and unmanaged forests result from the fundamental differences in the disturbance regimes of managed versus unmanaged forests. Declines in managed forest biodiversity and structural complexity, combined with rapidly changing climatic conditions, pose a risk to forest health, and hence, to the long-term maintenance of biodiversity and provisioning of important ecosystem goods and services. The application of ecosystem management in boreal forestry calls for a transition from plantation-type forestry toward more diversified management inspired by natural forest structure and dynamics.
The boreal forest consists of about 30% of the global forest area and is distributed across the northern hemisphere in Canada, Alaska, Russia, and Scandinavia. The boreal forest contains a large share of the world’s terrestrial carbon and provides habitats for numerous species adapted to cold climatic conditions. Although large regions of unmanaged forest still exists in high-latitude areas, the southern, more productive forests are under management of varying intensities (Burton et al., 2010). Human impact and forest management has been most intensive and long-lasting in Scandinavia, whereas many forest areas in Canada and Russia have been exploited for the first time only recently. In general, the boreal countries have long-standing and revered traditions in forestry education and institutions, along with timber resources management and exploitation methods. Forestry and the forest industry have played, and continue to play, an important role in the economy of the countries, which produce a large share of the lumber, pulp, and paper in the global export market.
The boreal forests of Scandinavia present an interesting case for ecosystem management because they encompass a wide range of forests, from those with a long history of intensive utilization in the southern regions to near-pristine forests in northern and/or high-altitude regions (Keto-Tokoi & Kuuluvainen, 2014). Recent comparative studies of ecosystem structure and dynamics between managed and unmanaged forests, as well as the long history of faunistic and taxonomic studies in countries such as Sweden and Finland, have provided a unique opportunity to analyze the long-term impacts of human utilization on forest habitats and biota (Esseen, Ehnström, Ericson, & Sjöberg, 1997; Siitonen, 2001; Lilja & Kuuluvainen, 2005). Analyzing this knowledge is valuable because the human impact on forests is predicted to intensify throughout the circumboreal zone (Gauthier, Bernier, Kuuluvainen, Shvidenko, & Schepaschenko, 2015). This, together with the anticipated rapid warming of the climate at high latitudes (IPCC, 2013), potentially threaten the health of forest ecosystems across the vast boreal biome (Bradshaw, Warkentin, & Sodhi, 2009; Gauthier et al., 2015). Knowledge and experience gained from dealing with Fennoscandian forests are valuable in developing functional and more sustainable management approaches that can help both forests and humans cope with the global changes to come (Kuuluvainen & Siitonen, 2013).
Forest use in Fennoscandia has taken varied forms at different times and places during the course of history. In southern regions, the permanent conversion of forests on fertile grounds to agricultural land dates back to medieval times (Keto-Tokoi & Kuuluvainen, 2014). From 1600 to the late 1800s, shifting cultivation and the production of extracts, such as charcoal, potash, and tar, were important and affected forests over large areas. As a result of these activities involving use of fire, forest fires increased significantly from previous levels. Forests were also widely used as pastures and a source of fodder for cattle, as well as a source of firewood. In forests outside agricultural use, timber was traditionally extracted by selective logging for domestic use, and to some extent as lumber for sawmills.
Starting in the early 1900s, Scandinavian forests became a subject for organized industrial-scale forestry, with the purpose of supplying raw materials to the rapidly expanding forest industries. Previously common forest fires were practically eradicated, as human activities in the forest decreased and fire prevention methods were developed. Following examples from German forestry, the enforced intensive management system was compartmentwise, even-aged management based on precommercial and commercial thinning and final clear-cut harvesting. Forest regeneration is carried out mostly by tree planting, but also by using seed trees. This high-input management regime is aimed at efficiently producing timber and preferably increasing sustainable yield. This goal has been achieved, but the enforced management methods also have caused major changes in forest ecosystem characteristics and triggered a decline in forest biodiversity (Siitonen, 2001; Auvinen et al., 2007). Changes detrimental to biodiversity include an increase in early-successional, plantation-type forests and a decrease in the amount of late-successional forests, both of which contributed to an overall decline in deadwood amounts and forest structural variability (Raunio, Schulman, & Kontula, 2008; Kuuluvainen, 2009).
The magnitude of change brought up by intensive forestry is exemplified in Finland, where a recent study evaluated 70% of forested habitat types on mineral soil (constituting 49% of the country’s forest area) as being threatened, mostly because of their low amounts of deadwood and simplified structure (Raunio et al., 2008). Moreover, such a comprehensive modification of forested landscapes by modern forestry has taken place very rapidly over most of boreal Scandinavia during the last 70 years (Keto-Tokoi & Kuuluvainen, 2014). Because of the slow response times of northern ecosystems, it is likely that we are only beginning to see the cumulative ecological effects of such grand-level alteration of forest ecosystem structure and dynamics. Such delays in ecological responses are due to long-lasting legacies from more natural forest stages in the past (e.g., slowly decaying pools of ground deadwood; Lilja & Kuuluvainen, 2005) and delayed population responses to habitat degradation and loss. This phenomenon is known as the extinction debt (Hanski, 2000; Hanski & Ovaskainen, 2000).
This, combined with the anticipated rapid climate warming at high latitudes, means that we may be facing an accelerating rate of loss of biodiversity and boreal carbon pools in the future (Moen et al., 2014; Gauthier et al., 2015). The possibility exists that at some point, the system crosses critical transition thresholds, resulting in ecosystem state shifts, after which sustainable management is no longer possible (Scheffer, Hirota, Holmgren, Van Nes, & Chapin, 2012; Gauthier et al., 2015). Such ecosystem state shifts from a closed forest to low-productivity open woodland have already been documented in some parts of the boreal (Jasinski & Payette, 2005; Girard, Payette, & Gagnon, 2008). In brief, the ongoing loss of habitat and species diversity is highly undesirable considering the anticipated changes in future environmental, economic, and social conditions brought about by global change.
In addition to ecological impacts posed by environmental change, changes in societal views and values are taking place regarding how and for what purpose forests should be managed. Thus, in addition to ecological threats, new social pressures on forestry force managers to accommodate a wider range of environmental, economic, and sociocultural values and goals than before. This calls for ecosystem-based management approaches incorporating adaptive systems perspectives (Messier, Puettmann, & Coates, 2013).
What Is Forest Ecosystem Management?
Under pressures of global changes and changing societal views, boreal forestry is forced to adopt new management approaches to sustain the ecosystem conditions and services that societies value and have come to depend on (Kuuluvainen, 2009; Moen et al., 2014; Puettmann, 2014; Gauthier et al., 2015). However, defining forest ecosystem management is not an easy task. Here, we choose the definition of Gauthier et al. (2009, p. 26), who argue that ecosystem management is “[a] management approach that aims to maintain healthy and resilient forest ecosystems by focusing on reduction of differences between managed and natural landscapes to ensure long-term maintenance of ecosystem functions and thereby retain the social and economic benefits they provide to society.”
Here, the term forest health can be defined following Brandt, Flannigan, Maynard, and Thompson (2013, p. 221), as the maintenance of desirable ecosystem function and processes, biodiversity, resistance to biotic and abiotic disturbances, and ability to renew after these disturbances. The three important interrelated indicators of ecosystem health are biodiversity, resilience, and adaptive capacity. Biodiversity denotes the variation of ecosystem structures and processes, and species (including genetic variation) in space and time. Resilience denotes the resistance to and ability of ecosystems to recover from disturbances. Ability to renew can be understood as the adaptive capacity, which refers to the long-term ability of ecosystems to evolve and adjust to changing environmental conditions, either through short-term changes in structure and species composition and functioning (acclimation) or through long-term changes in the genetic pool of the organisms (adaptation). The adaptive capacity aspect is becoming an increasingly important feature of sustainability, as human utilization is modifying global environmental conditions and forest ecosystems at an unprecedented rate. We are likely to enter an era of rapid global change with possibly drastic consequences on environmental conditions, especially in the northern latitudes (Gauthier et al., 2015). In addition to commodity production, ecosystem managegement focuses on the maintenance of biodiversity because it directly links to a range of ecosystem services and allows a variety of responses to changing environmental conditions, therefore increasing the capacity of the forest ecosystem to adjust and take advantage of new conditions (Kuuluvainen & Siitonen, 2013; Filotas et al., 2014).
Natural Disturbance Emulation and Ecosystem Management
According to the definition of ecosystem management by Gauthier et al. (2009), a reduction of the differences between managed and natural landscapes is seen as a central means of maintaining ecosystem health and associated properties. To achieve this, restoration and maintenance of natural forest structures by emulating natural disturbances have become a mainstream approach in forest ecosystem management (Attiwill, 1994; Bergeron, Leduc, Harvey, & Gauthier, 2002; Gauthier et al., 2009; Kuuluvainen & Grenfell, 2012). The underlying idea is that by adjusting forest management, particularly timber harvesting, to the disturbance patterns observed in natural disturbances, the majority of the natural habitat structures, processes, and species of the ecosystem could be safeguarded (Fig. 1).
Whether this is true naturally depends on how well, and to what extent, the ecological functions and habitat structures and their natural range of variability (NRV) can be emulated at different scales in forest management.
The idea of using the natural variability of forests and their developmental patterns as a reference for sustainable forest management is an old one and was formulated in Europe during the 19th century (Johann, 2006). Although the goals of management and understanding on what is a “natural” forest have changed over time, the idea of working parallel with nature’s own dynamics has been persistent (Leibundgut, 1978, 1981). During past decades, our understanding of the structure and dynamics of natural forest ecosystems has grown exponentially (Kuuluvainen & Aakala, 2011). There has been a growing understanding of the importance and prevalence of natural disturbances in forest ecosystems, including the boreal forest (Kuuluvainen, 1994, 2009; Kneeshaw, Bergeron, & Kuuluvainen, 2011). Following the development of ecological theory and research on forest disturbances, the idea of using natural disturbances as the basis of ecosystem management has gained ground (Attiwill, 1994; Angelstam, 1998; Bergeron et al,. 2002; Seymour et al., 2002; Kuuluvainen & Grenfell, 2012). This development is based on the fundamental recognition that natural disturbances are of pivotal importance for the maintenance of biodiversity and key ecological processes in forest ecosystems, and not merely a harmful factor (Fig. 1; also see Attiwill, 1994).
The natural disturbance emulation approach has been criticized as inappropriate in the future because it is based on past conditions and ecosystem dynamics and therefore is not applicable to “novel ecosystems” that will appear in changed conditions. It can be argued, on the other hand, that although future ecological conditions and ecosystems probably differ from past ones, the basic features of ecosystem functions, as well as their responses and dynamics, will remain (Keane et al., 2006; Kuuluvainen & Grenfell, 2012). Another argument is that because the existing biota has evolved under past environmental conditions, ecosystem management has to pay due respect to those past conditions in some way. These arguments support the view that accumulated ecological knowledge is valuable and can be used to develop the ecosystem management of forests also in future changing conditions (Keane et al., 2006; Kuuluvainen, 2009).
Defining Reference Conditions for Forest Ecosystem Management
It is generally accepted that knowledge concerning the variability of past states and conditions of unmanaged ecosystems, referred to as their natural/historical range of variability (NRV/HRV), provides an important reference for developing methods to evaluate the success of forest restoration and ecosystem management (Keane et al., 2006; Gauthier et al., 2009). Indeed, resent research in the remaining wilderness areas of the boreal zone has provided important new insights into the structure, dynamics, and biodiversity of the boreal forest under negligible human influence (Kneeshaw et al., 2011). New knowledge concerning forest disturbance ecology is obviously of pivotal importance when developing ecosystem management based on natural disturbance emulation.
The conventional view of boreal forest dynamics held that stand-replacing fires constitute the most important and dominating natural disturbance factor across the boreal biome, and that regeneration after such events results in the development of even-aged stands (e.g., Sirén, 1955). However, the accumulated body of research results indicates that such a simplified view is not supported by accumulating research evidence, indicating that historical disturbance regimes in boreal forests have displayed substantial variability in disturbance frequency, type, and severity (Shorohova, Kuuluvainen, Kangur, & Jogiste, 2009; Kneeshaw et al., 2011; Kuuluvainen & Aakala, 2011).
In Fennoscandia, for example, fire regime reconstructions, which were based on dendrochronological studies (fire scar samples and dating) and paleoecological data (sediment samples from peat layers and lake bottoms), suggest that prior to significant human influence, fire cycles were considerably longer than during the 17th to 19th centuries, when humans were highly active in the forest. Based on the available evidence, it can be estimated roughly that the historical “natural” range of fire cycles varied between 100 and 500 or more years (Niklasson & Granström, 2000; Pitkänen, Huttunen, Jugner, & Tolonen, 2002; Pitkänen, Huttunen, Tolonen, & Jugner, 2003; Wallenius, Kauhanen, Herva, & Pennanen, 2010). This is because the moist to mesic spruce Picea abies–dominated forests rarely ignite, and thus have intrinsically long fire rotations (Ohlson et al., 2011). Fires are more common but less severe on drier pine Pinus sylvestris–dominated sites due to lower fuel availability. In addition, the high fire resistance of large, thick-barked Pinus trees, together with common fire skips, typically result in only partial or patchy tree mortality. This leads to the development of structurally diverse forests with several age cohorts of trees, with some trees surviving fires and some regenerating after them (Kuuluvainen, 2002; Wallenius et al., 2010).
Nonpyrogenic forests also occur, which have experienced no fire since the last glaciation (Zackrisson, Nilsson, Steijlen, & Hörnberg, 1995; Pitkänen et al., 2003). When fire cycles extend to hundreds of years, or fire is absent altogether, forest dynamics is driven by nonpyrogenic disturbance agents such as wind, fungi, and insects, and landscapes become characterized by the dominance of old forest driven by patch- or gap-scale dynamics (Kuuluvainen, 1994; McCarthy, 2001).
Although methodologically difficult, it is conceptually important to make a distinction between historical human-affected fire regimes and their range of variability (HRVs) and historical “natural” fire regimes (NRVs), with negligible human impact, under which the boreal biota has developed over past millennia (Josefsson, Hörnberg, & Östlund, 2009). This is exemplified by the contrast between the longer natural fire cycles compared with the much shorter fire cycles due to mostly anthropogenic fires in the 17th to 20th centuries. The fact that most fire history studies temporally overlap with the period when humans significantly increased fires can easily lead to overestimation of the intrinsic occurrence and ecological role of fire in forests (Kuuluvainen, 2009).
Fire regime characteristics vary geographically. In Fennoscandia, for example, factors affecting fire occurrence and its variability include the length of the fire season in the south-north gradient and the density of lightning strikes (Granström, 1993; Larjavaara, Kuuluvainen, & Rita, 2005). In general, lightning ignitions increase from north to south and from coastal to inland areas (Granström, 1993; Larjavaara et al., 2005). At the landscape level, important factors affecting fire behavior include topography and the often fine-scale patchiness of the landscapes, with an abundance of paludified forest and natural fire breaks (peatbogs, lakes, and waterways). As a consequence, sections in a landscape may burn rarely, escape fire altogether, or burn as low-intensity surface fires (Lampainen, Kuuluvainen, Wallenius, Karjalainen, & Vanha-Majamaa, 2004; Wallenius, Kuuluvainen, & Vanha-Majamaa, 2004). The commonness of fire-resistant tree species, such as Scots pine (Pinus sylvestris), makes total stand-replacement uncommon over large areas. All these factors and conditions explain the dominance of old forest in naturally dynamic landscapes (Kuuluvainen, 2009).
Although obviously intrinsically more rare than previously assumed, fire is an important ecosystem driver in the boreal forest because of its long-lasting influence on ecosystem properties (e.g., Aakala, Kuuluvainen, Wallenius, & Kauhanen, 2009). Especially high severity fires, occurring in a stochastic manner, impose long-term legacy effects on ecosystems in terms of tree age structures and deadwood dynamics (Gromtsev, 2002). On moist fertile sites, deciduous tree succession takes place after high-severity fires. Fire, along with deciduous litter, release soil nutrient resources and activate their cycling (Sirén, 1955). Deadwood pools and stand dynamics also can be dictated by past fire events. For example, an old-growth spruce forest examined by Aakala et al. (2009) proved to consist of a single tree cohort initiated after a severe fire 317 years prior to the study. This demonstrates the slow development and extremely long-lasting legacy effects of severe fire events in these northern forest ecosystems.
The commonness of partial disturbances, as well as fire cycles, which are longer than the lifespans of tree species, indicate that disturbance agents other than fire play an important role in driving forest dynamics in natural conditions (Kuuluvainen & Aakala, 2011). In the absence of severe disturbances, trees die due to competition- and senescence-related attacks of fungi and insects, or due to the forces of wind or heavy snow loads (Lännenpää et al., 2008). These are fine-scale, bottom-up processes associated with neighborhood interactions and leading to gap and patch formation (Kuuluvainen, 1994; Kuuluvainen & Aakala, 2011). The resulting structural heterogeneity is a key ecosystem property and could be called self-generated complexity (Harris, 2007).
Types of Forest Dynamics and Landscape Age Structures
The high variability of natural disturbance characteristics and successional pathways in forests, as reviewed previously, makes the definition and categorization of forest dynamics challenging. However, for management purposes, it is important to describe—even only roughly—the most important stand dynamic types (Kuuluvainen, 2009). Accordingly, the following simplified classification of forest disturbance–succession cycles and the resulting structural dynamics has been suggested (Fig. 2; also see Angelstam & Kuuluvainen, 2004; Shorohova, Kuuluvainen, Kangur, & Jogiste, 2009):
• Class 1: Even-aged forest dynamics, driven by stand-replacing disturbances
• Class 2: Cohort forest dynamics, driven by partial disturbances
• Class 3: Gap dynamics, driven by tree mortality at a fine scale
Although strongly simplifying, such a classification is useful because it helps to translate the daunting variability found in natural forests into reference types of forest dynamics.
Forest age distribution at the landscape level is formed by the prevailing types of forest dynamics, which depend on disturbance factors and their interactions with trees possessing varying life history traits (Shorohova et al., 2011; Rogers, Soja, Goulden, & Randerson, 2015). As discussed previously, the available scientific evidence suggests that, contrary to some earlier beliefs, stand-replacing disturbances (Class 1 dynamics) do not always play a dominant role in the dynamics of pristine boreal forest. This appears to be especially the case in Western Eurasia (Shorohova et al., 2009; Kuuluvainen, 2009; Kuuluvainen & Aakala, 2011), where cohort dynamics and gap dynamics (Class 2 and Class 3 dynamics) would be most prevalent (Fig. 2; also see Kuuluvainen & Aakala, 2011). This conclusion is supported by historical studies from Fennoscandia and recent empirical studies carried out in unmanaged landscapes in eastern boreal Fennoscandia (Kuuluvainen, 2009; Kuuluvainen & Aakala, 2011). Based on empirical studies of forest disturbance ecology, stand dynamics, and simulation modeling, Pennanen (2002) concluded that uneven-aged stands containing at least some trees living up to their biological age of 200–300 years would be a characteristic feature of unmanaged forests. These characteristics define the NRV of forest and ecosystem reference conditions for ecosystem management in Fennoscandian boreal forests.
Operationalizing Ecological Knowledge in Forest Ecosystem Management
Defining ecological reference conditions, in terms of forest structure, dynamics, and biodiversity, is indispensable for developing methods and goals for ecosystem management. However, the knowledge has to be translated into strategies, tactical models, and operational principles before they can be applied effectively in practice. In this section, a number of general principles, strategies, and more practical models are presented, through which ecosystem management can be realized.
General Strategic Principles
1. Focus on habitats rather than on species. In most cases, a decline in biodiversity is due to habitat loss or degradation. Because of this, maintaining natural habitats or their essential characteristics is the most important principle of ecosystem management (Gauthier et al., 2009, also see “Natural Disturbance Emulation and Ecosystem Management”). In conservation, this has also been called the coarse filter strategy, which aims to maintain the natural features of the habitat mosaic and, consequently, viable populations of all native species and ecosystem functions, not just populations of some large or “sexy” species. The aim of this holistic approach is to take into account the entire species community, including any poorly known and unknown species (Kuuluvainen & Siitonen, 2013). Species-by-species planning is a contrasting approach that is unsuitable as a comprehensive strategy for ecosystem management because the habitat requirements are known for only a small number of species, and acting for the benefit of one species can easily harm another.
2. Plan and act at multiple scales. Biodiversity and ecological processes are represented at different hierarchical levels in the forest ecosystem, ranging from fungi-decaying fragments of deadwood to large predators with regionwide habitat ranges. Incorporating multiple scales into ecosystem management calls for hierarchical planning, which means thinking and acting at the same time at various levels of the forest ecosystem. The microhabitat, stand, and landscape scales are important to pay attention to simultaneously. Microhabitats are small objects, such as twigs, snags, logs, stumps, or big and hollow trees, which provide habitat and shelter for small species. Important stand-level habitat characteristics to consider are tree species composition, tree size and age distribution, the spatial pattern of trees, and the amount and quality of deadwood. Important landscape-level attributes are distribution of stand ages or forest successional stages, amount of protected forest, and habitat fragmentation and connectivity. Across-scale thinking may be particularly challenging in forestry because forestry operations are often planned and executed at the forest stand or compartment level.
Landscape fragmentation is a common problem related to forest management. It means that a previously more continuous forest habitat becomes divided into smaller parts. Fragmentation as a result of management usually occurs parallel to habitat loss. Apart from the negative effects of habitat loss, fragmentation of the remaining habitat increases the area affected by edge effects and decreases the interior forest habitat and the connectivity between the remaining habitat patches. Ameliorating the adverse effects of fragmentation is a challenge in forest ecosystem management, but its effects can be ameliorated by the spatial arrangement of unmanaged areas, by using appropriate silvicultural tools (e.g., continuous cover silviculture), and by thoughtful planning of harvesting cuttings and methods in space and time.
3. Pay special attention to aquatic habitats. Special attention is needed for aquatic and riparian habitats because they are hot spots of biodiversity (e.g., Nummi & Kuuluvainen, 2014). Such aquatic habitats include brooks, rivers, beaver ponds, lakes, springs, vernal pools, paludified areas, and peat bogs. These habitats are characterized by high plant species diversity, which hosts high overall organism diversity, and they have low resilience to severe disturbances because they usually escape fires and storms and are instead characterized by small-scale disturbances and the continuity of specific conditions. As such, they act as ecological corridors and refugia for species in the disturbance-prone landscape, helping species to survive, move and colonize postdisturbance sites. The question of whether these riparian habitats should be left outside management depends on the possibility of maintaining their ecological characteristics and important functions in the landscape.
4. Manage processes rather than structures. Forest ecosystems are constantly changing due to their autogenic dynamics, allogenic disturbances, and variation in environmental conditions. This emphasizes that for attaining the goals of ecosystem management (i.e., maintenance of biodiversity, ecosystem resilience, and adaptive capacity), managers should pay special attention to the long-term spatial and temporal processes that replenish natural ecosystem structures in the landscape. Designing restoration and sustainable management as inspired by natural disturbances provides a tentative framework that emphasizes long-term ecosystem dynamics (Bergeron, Leduc, Harvey, & Gauthier, 2002; Kuuluvainen & Grenfell, 2012).
An example of ignoring the long-term perspective and nature’s constant change and connectedness is provided by the once common “nature-as-museum” thinking. It was thought that protecting small patches of natural forest was enough to maintain representative samples of “original nature.” However, such a strategy was doomed to fail because processes driving ecosystem change occur on a larger spatial scale (e.g., climate change at global scale, disturbance and species meta-population dynamics at landscape and regional scales). Besides being doomed to ecological deprivation due to isolation, a single fire or storm event can wipe out small habitat patches and much of their species. Planning habitat networks, based on ecological understanding of local disturbances and species population dynamics, is fundamental for successful ecosystem management.
5. Consider managed and protected areas in concert. The traditional approach to protecting nature in managed landscapes has been to establish specific set-aside areas, which can be small (such as the so-called key biotopes) or larger forest patches. The success of this approach depends on the area, quality, representativeness, and connectedness of the protected areas, as well as on the quality of the surrounding managed forest matrix. If the protected areas sufficiently and representatively cover all ecosystem types, they probably can ensure critical ecosystem properties, such as the connectedness and variability of disturbances (and hence biodiversity).
However, in many boreal regions, the cover of protected areas is very small and geographically unevenly distributed, which means that they are often isolated, do not represent the natural variability of ecosystem types, or both (Potapov et al., 2008). For example, the protected areas in Finland and Sweden are primarily located in the low-productivity, northern high-altitude forests with low levels of native biodiversity compared to more productive, low-altitude, or southern areas. Thus, the representativeness of the reserve network is typically far from adequate. This problem is particularly acute in the southern regions, where natural biodiversity would naturally be highest (Auvinen et al., 2007; Angelstam et al., 2011).
Another problem is the lack of a guarantee that species and habitats will be found in the future where they occur today. This is, first, because habitats may be too small and too affected by harmful edge effects to maintain their ecological characteristics in the long run. Small protected areas also may be the last isolated refuges left for species remaining as small relict populations, which are doomed to local extinction (i.e., the extinction debt). This is because of the stochastic source-sink dynamics of populations between habitat patches. If the habitat network is not sufficiently connected for the species in question, meta-population dynamics may fail, resulting in local or regional species extinction. Second, habitats may be destroyed by unexpected severe disturbance events, especially if the protected areas are small. Third, habitats also may be degraded or even undergo state shifts due to lack of natural disturbance processes or environmental change. The anticipated rapid warming at high-latitude ecosystems would mean that a dramatic transformation of ecosystem and habitat characteristics would occur by the end of the 21st century (Gauthier et al., 2015).
To foster the adaptive capacity of landscapes and protect their biodiversity, it is necessary to concurrently consider landscapes consisting of protected areas and managed forests. Even when a relatively large proportion of forest is under protection, the majority of forest area and its species populations typically will be under some kind of forest management. Therefore, the way that the managed forest is treated plays a key role in ecosystem management. Managed forest fundamentally contributes to the amount, configuration, and connectivity of available habitats in the landscape. Important considerations include how to treat the managed forest in the vicinity of protected and riparian areas to reduce harmful edge effects, and to improve the connectivity of important habitat.
Strategies for Forest Ecosystem Management
Coarse- and Fine-Filter Strategy
The coarse- and fine-filter strategy distinguishes two complementary approaches for protecting biodiversity. The coarse-filter strategy emphasizes the importance of maintaining natural ecosystem structures (habitats) at the landscape scale (Fig. 1; also see Hunter, 1993). This requires knowledge of the historical range of variability (HRV) of ecosystem properties based on the understanding of past disturbance regimes (Keane et al., 2006). The assumption is that restoring or maintaining coarse-scale landscape conditions within a range to which the biota has adapted will most likely conserve biodiversity and important ecosystem functions, and hence maintain sustainable ecosystems. The coarse-filter approach does not necessarily consider reserves; rather, it emphasizes ecological processes and the dynamic distribution of habitats across the landscape over time. The coarse-filter strategy has been promoted, as it may avoid the pitfalls of narrow species-by-species conservation approaches, which are ultimately restricted by lack of knowledge of ecological habitat preferences for the majority of species.
A complementary strategy to the coarse filter is the fine-filter approach, which focuses on individual species or fine-scale elements of biodiversity, which cannot be accounted sufficiently by the coarse-filter approach. The fine-filter approach consists of specific and tailored conservation strategies for focal species, which are considered to be at particular risk under the coarse-filter approach. Examples of fine-filter strategies include providing nesting boxes for cavity-nesting species (e.g., the flying squirrel) when hollow trees are not sufficiently available in the landscape.
Natural Disturbance Emulation (NDE)
Natural disturbance emulation (NDE) is an approach that actually implements the coarse-filter strategy in forest ecosystem management (Fig. 1; also see Angelstam, 1998; Bergeron et al., 2002; Kuuluvainen, 2002). According to this approach, management actions are planned and executed to emulate natural disturbances and their outcomes (Fig. 3).
This is carried out at multiple scales, from deadwood microsites to stand structures and landscape patterns, to create a natural-like habitat mosaic pattern. The natural disturbance approach is an appealing approach, especially when protection areas are scarce.
Applying the natural disturbance model in management requires adequate knowledge of the historical disturbance regime and forest structure at the landscape scale—that is, the natural range of variability (NRV) (Kuuluvainen, 2002). However, in many areas, such a knowledge basis is not available because of pervasive human impact on forests and the small size of the remaining protected fragments, which prohibit studies of natural forest dynamics at the landscape level. For example, stand structures and species composition in small protected areas in southern Fennoscandia were found to differ considerably from naturally dynamic larger landscapes in Russian Karelia (Lilja & Kuuluvainen, 2005; Nordén, Penttilä, Siitonen, Tomppo, & Ovaskainen, 2013).
Despite such difficulties, it is usually possible to define the main features of the historical range of forest dynamics and variability for a specific area, using information from multiple sources such as (a) retrospective analyses based on historical materials, (b) historical reconstructions based on biological archives using dendrochronological and palaeoecological methods, (c) studies carried out in geographically close and ecologically similar more natural forests, and (d) integrative modeling (e.g., Pennanen, 2002). In Fennoscandia, for example, northern European Russia provides a special opportunity to reference landscape studies, as large areas of primeval forests still exist there (Potapov et al., 2008).
The history of human impact and modification of forest habitats vary greatly in different parts of the boreal zone. In many regions in northern Northwest Russia and western Canada, primeval forest habitat is still abundant and probably above the habitat threshold of most species. On the other hand, in southern Fennoscandia, where only a few percent of unmanaged forest is left, the level of primeval forest habitat is certainly well below the threshold of demanding species confined to natural forests (Hanski, 2000). In such situations, the only way to apply ecosystem-based management is to carry out habitat restoration (e.g., by implementing natural disturbance emulation) (Kuuluvainen, 2009; Nordén et al., 2013).
Restoration is an important component of ecosystem management in human-modified forests. Forest restoration has the goal of rehabilitating those structures and attributes that are found in natural forest ecosystems. Following this definition, the natural disturbance emulation in managed forests can be regarded as restoration. Restoration is needed when the ecosystem is degraded due to intensive and long-lasting human utilization. Restoration actions often aim at quickly creating critically important habitats, such as deadwood, or initiating important long-term processes such as vegetation succession and paludification (Kuuluvainen et al., 2002; Halme et al., 2013).
Tactical Models for Forest Ecosystem Management
Tactical models are needed to implement holistic strategic approaches, such as the coarse-filter strategy and NDE, into real-world forest management. Only a few such tactical conceptual models have been proposed for the boreal forest. These models are tactical because they make concrete, practical proposals on how to act to attain specific management goals. Examples of such tactical models are provided by (a) the patch-corridor-matrix model (Forman, 1995), (b) the ASIO-model, and (c) the multi-cohort model, all of which are discussed next.
The patch-corridor-matrix model is based on the conceptualization of landscapes as mosaics of spatially separate habitat patches (the patch-mosaic paradigm; Wiens, 1997). Three different types of mosaic elements are distinguished: (a) habitat patches, which are suitable for the species being considered and clearly separable from their surroundings; (b) corridors and stepping stones, which help and allow organisms’ movements between habitat patches; and (c) matrix forest, which is often assumed to be an unsuitable (or at least a low-quality) habitat for the organisms of interest (Forman, 1995). The background of this model lies in the theory of island biogeography (MacArthur & Wilson, 1967) and the metapopulation theory (Hanski, 1999). The model has been widely applied as a theoretical approach and tool in “forest landscape ecological planning” in Finland and Sweden, both on public and private forestland.
The management approach is straightforward: (a) valuable habitat patches and (b) corridors and stepping stones are either set aside or managed in a way that maintains their key ecological characteristics. Normal forest management is applied in (c) the remaining matrix area. If the set-aside patches are habitats of low natural disturbance probability, such as paludified depressions and riparian areas, this could be seen to represent some form of natural disturbance emulation. However, the model’s representation of landscape structure is extremely simplified and static, and therefore it does not emulate well the complexity of structure and dynamics of unmanaged forest ecosystems (Filotas et al., 2014).
The ASIO-model was proposed in Sweden in the 1990s to enhance biodiversity conservation challenges in intensively managed forests (Rülcker, Angelstam, & Rosenberg, 1994). The model was inspired by research on the pivotal ecological role of fire and other disturbances in the northern boreal forests (Angelstam, 1998). It is founded on the basis of research indicating that the probability of fire occurrence in boreal Fennoscandia is related to site fertility and the soil moisture gradient (Zackrisson, 1977; Wallenius et al., 2004), so that fire frequency varied from low, in moist herb-rich forests with spruce and deciduous trees, to high, in dry sandy heaths dominated by Scots pines. The acronym of the model comes from the four perceived fire regimes along this site type gradient: namely, fire being Absent or Seldom, Infrequently and Often present. The management actions are based on these simplified assumptions of fire ecology. The four classes of the ASIO-model can be related to various management and cutting methods. To be compatible with the principles of ecosystem management, these cutting methods should be congruous with the assumed frequency and effects of natural fire on forest structure. In practice, herb-rich sites with highest biodiversity and conservation value are set-asides, while cuttings and the management of variable types and intensities are possible in the remaining portion of the landscape (Angelstam, 1998).
The ASIO-model resembles the patch-corridor-matrix model in that it sets aside vulnerable areas of highest biological value. How well the management practices in the rest of the landscape are able to emulate natural forest structures depends on the knowledge level concerning historical disturbance regimes along with the level of ambition. Different restoration and silvicultural tools are available to emulate the structures of fire-prone forests (Kuuluvainen, 2009). However, the assumption made by the ASIO-model—namely, that fire frequencies are constant and dictated solely by site properties—is not realistic. This is because forest fires are a stochastic phenomenon, and factors such as landscape pattern also play an important role in the fire regime (Hellberg, Niklasson, & Granström, 2004).
In practice, the ASIO-model has been used more as a theoretical reference, with limited ambition to use the model as a true restoration or ecosystem management tool. Instead, conventional low-retention cutting is used across most of the forest area (i.e., on areas belonging to the S-, I-, and O-classes), with a loose justification for “emulating” high-severity fire dynamics. In such a case, the outcome is basically the same as with the corridor-patch-matrix model (see “Patch-Corridor-Matrix Model”). However, this poorly emulates the natural fire effects on forests, which we know to be highly variable (e.g., Kuuluvainen, 2009).
The multi-cohort model, implementing natural disturbance emulation, was suggested by Bergeron, Harvey, Leduc, and Gauthier (1999) and Bergeron et al. (2002). It is grounded on the pioneering research of forest fire ecology carried out in Quebec, in eastern Canada (Bergeron et al., 2010). In disturbance emulation, the model explicitly considers multiple spatial scales from stand structure to landscape pattern. Intensive fire of variable periodicity is the dominant disturbance agent to be considered. The fundamental idea is that after a stand-replacing fire, forest succession moves through distinct phases, called structural cohorts, which are under the potential influence of further disturbance events. The natural proportions, or the range of variability, of these various structural cohorts can be derived from a historical analysis of the past spatiotemporal distribution of fire in the landscape. For example, under conditions where stand-replacing disturbances are common, the landscape is dominated by young forest, while old forest prevails when disturbances are infrequent and partial.
Based on the analysis of historical landscape forest age structure and disturbance regime, different silvicultural methods are applied at the stand level to maintain the desired proportions and spatial arrangement of different forest “structural cohorts” in the landscape (Bergeron et al., 2002). Stands are subjected to different cutting treatments depending on which “structural cohort” they should emulate. Younger stands can be treated with variable density thinning in order to emulate natural self-thinning processes. Older stands, representing late-successional stages, can be subjected to partial or gap cutting to emulate small-scale disturbances and associated stand structures. Finally, clearfelling with appropriate retention is applied to emulate severe stand-replacing disturbances and to regulate the proportion of young forest age classes.
It is assumed that through this approach, it is possible to maintain forest stand and landscape structures within their NRV. In addition to the variability of management methods, the importance of set-aside areas for conservation are highlighted. An important distinction from the ASIO-model is that fires were originally assumed to occur randomly across the landscape in the multi-cohort model (Bergeron et al., 2002), whereas fire frequency was assumed to be dependent on site type in the ASIO-model. Although neither of these assumptions may be strictly realistic, they probably reflect the fact that high-severity fires have historically been more common in northern North America, while mixed-severity fires have prevailed in northern Eurasia (Kneeshaw et al., 2011; Rogers et al., 2015).
Hierarchical Implementation of Ecosystem Management
The different tactical models of ecosystem management reviewed in the previous sections can also be seen as complementary approaches to hierarchical ecosystem management (Kuuluvainen & Grenfell, 2012). The patch-corridor-matrix model emphasizes the importance of network connectivity of permanent unmanaged conservation areas. Models of natural disturbance emulation, the ASIO-model and multi-cohort model, provide the principles of how, by implementing natural disturbance emulation in the managed forest, it is possible to maintain desired forest structures at stand scale and habitat mosaic at the landscape scale. Such a hierarchical landscape design is essential because it creates synergies between the different approaches and methods applied in forest ecosystem management.
A functional conservation area network is an essential component of forest ecosystem management at the landscape scale. Set-aside areas are needed to sustain specialized and vulnerable species that are strictly confined to natural forest habitat. Restoration can be applied to those parts of protected areas that are degraded by past human utilization. On the other hand, natural disturbance emulation can be implemented in the managed forest matrix to keep the forest within its NRV (i.e., coarse-filter strategy; Kuuluvainen & Grenfell, 2012). For example, it may be useful to create buffer zones in the vicinity of conservation areas by applying continuous cover silviculture with a strong restoration component. Such a restoration-oriented silviculture would aim to create ecologically important natural-like forest structures with higher deadwood levels (e.g., 30–60 m3/ha) than in ordinary managed forests, while still harvesting most of the timber and ensuring rapid regeneration (Vanha-Majamaa et al., 2007, Koivula et al., 2014). Such approaches, with a strong restoration emphasis, could enhance the functioning of protected areas and thus improve the efficiency of the overall strategy for sustainable forest management.
In the hierarchical approach, partial and local-scale disturbances are emulated by practicing continuous cover silviculture to maintain fine-scale structural heterogeneity (Kuuluvainen, 2009). At the landscape scale, clearcutting with retention is carried out to emulate severe disturbances (to the extent they would occur naturally), while a sufficient amount of retention trees is left at the cutting sites. Methods like variable density thinning can be used to enhance structural heterogeneity in young stands to imitate the variability of stand successional pathways characteristics of natural forest development.
A Motivation and Way Forward in Forest Ecosystem Management
Global change, which consists of the combined effects of direct human activity and climate change, poses a serious threat to boreal forest health and the provisioning of vital ecosystem services that societies depend on both locally and globally (Bradshaw et al., 2009; Burton et al., 2010; Gauthier et al., 2015). The threat of major ecosystem changes, including increased disturbances due to a warming climate and the increasing possibility of ecosystem state shifts (Scheffer et al., 2012), calls for ecosystem-based precautionary approaches in boreal forest management. This requires a shift from wood supply–focused management toward management approaches that may not be optimal for one asset like timber production, but which are functional under a range of possible scenarios of future environments. Such no-regret management scenarios must pay special attention to restoring and maintaining ecosystem components and properties vital to system health, such as biodiversity, resiliency, and adaptive capacity (Gauthier et al., 2015). The aim should be to lower the risk of highly undesirable outcomes, which could arise from simplified boreal forest ecosystems unable to adapt to future conditions (Gauthier et al., 2015). Such outcomes could be an increased severity and amount of disturbances, which could result in ecosystem state shifts under some climate warming scenarios, so that large tracts of closed canopy forest would become open canopy woodland, shrub land, or even steppe (Scheffer et al., 2012).
Especially in countries like Sweden and Finland, managed boreal forests are heavily homogenized as a result of long-term utilization and intensive management for timber production. Considering the anticipated rapid climatic change in high-latitude regions, a shift from mainly timber-oriented management approaches to ones provisioning and maintaining a wider set of ecosystem goods and services is desirable (Burton et al., 2010). To accomplish this, management practices need to conform to the general principles of ecosystem approach or ecosystem management (CBD, 2004), which emphasize the maintenance of intrinsic structures, functions, dynamics, and biodiversity of ecosystems under management (Gauthier et al., 2009). This is an ambitious goal, requiring the development and application of new management approaches (Bergeron et al., 2002; Kuuluvainen & Grenfell, 2012; Messier et al., 2013).
The accumulated ecological knowledge from unmanaged wild forests indicates that for restoring and sustaining crucial ecosystem properties, more attention must be paid to maintaining variability in forest structures and tree species compositions at different spatial scales and across all temporal (or successional) stages of forest development (Bergeron et al., 2002, Bengtsson et al., 2003). This is because the ecological processes that regulate the availability of ecosystem services occur at characteristic spatial scales ranging from microscale (e.g., habitats for small immobile creatures), to stand scale (e.g., timber production), to watershed or landscape scale (e.g., visual quality, water purification, carbon sequestration, or species migration). The maintenance of ecological processes interacting across scales is important for ecological resilience and, subsequently, for the long-term provision of the goods and services that ecosystems provide for local communities and the global market (Bengtsson et al., 2003; Drever, Peterson, Messier, Bergeron, & Flannigan, 2006; Gauthier et al., 2015).
Favoring live and dead tree retention is one avenue that has gained acceptance and is used in boreal forestry (Gustafsson et al., 2012). However, this approach in low-retention systems may not be enough to reproduce an important range of habitat feature variability of late-successional forests at stand and landscape scales (Burton et al., 1999). A more diverse set of management methods applied over a range of scales is therefore required for ecosystem management (Gauthier et al., 2009; Kuuluvainen & Grenfell, 2012; Halme et al., 2013). For example, a greater use of irregular shelterwood systems could be employed in forests naturally subjected to intermediate-severity partial disturbances (Seymour, 2005; Raymond, Bedard, Roy, Larouche, & Tremblay, 2009).
When forest structures are strongly altered due to management (e.g., Fennoscandian boreal forests), ecosystem management could involve deliberate restorative practices that aim at rehabilitating ecosystem complexity and species diversity across multiple scales (Drever et al., 2006; Burton et al., 2010). Forest harvesting, regeneration, and stand-tending activities can be developed to maintain or restore ecosystem complexity (Messier et al., 2013). Such a management and restoration strategy can be realized through diversifying silvicultural and management practices to create and maintain a wider range of forest structures inspired by knowledge gained from unmanaged forest ecosystems (Angelstam, 1998; Kuuluvainen, 2009; Gauthier et al., 2009; Shorohova et al., 2009; Raymond et al., 2009). Instead of the currently dominant use of even-aged management, a variety of multi-cohort and uneven-aged management practices should be implemented. A common counterargument to this point is that applying systems other than even-aged management ones is not possible because of their low economic profitability. However, recent studies have shown that uneven-aged management can be competitive with even-aged management from the economic point of view (Tahvonen, 2009; Tahvonen, Pukkala, Laiho, Lähde, & Niinimäki, 2010), while simultaneously fulfilling ecological and social sustainability goals (Kuuluvainen et al., 2012; Pukkala, 2016). This is encouraging and indicates that the grand goal of forest ecosystem management—the reconciliation of economic profit and maintenance of ecosystem health and diversity, and hence various ecosystem services—is indeed possible in boreal forestry.
Boreal forestry is confronted with increasing challenges posed by global change and demands to combine wood supply with the provisioning of a wide range of other ecosystem goods and services for society. Experiences in boreal countries like Sweden and Finland show that intensive forest management for timber has been successful in providing raw materials for industries, but has concurrently led to a simplification of forest ecosystem structure and decline in biodiversity (e.g., Auvinen et al., 2007; Raunio et al., 2008). To a large extent, this has been attributed to the monotonous use of even-aged plantation-type management and other forms of low-retention harvesting systems with short rotations, which create forest structures that are fundamentally different than those found in unmanaged forests (Kuuluvainen, 2009). From the boreal ecosystem point of view, such intensive resource utilization and modification, together with ongoing climate change, implement conditions and disturbance regime characteristics, which the ecosystem has never experienced and to which elements of the biota therefore have difficulty adjusting (Nordén et al., 2013). The result is a decline in biodiversity, which may put forest health at risk (Gauthier et al., 2015). Implementing ecosystem management should be a high priority in this situation; it is needed to safeguard biodiversity, resilience, and adaptive capacity and lower the risk of highly undesirable outcomes, such as ecosystem state shifts. Diversifying forest management and silviculture to restore and maintain tree species and forest structural diversity at multiple scales, as inspired by unmanaged forests, is a promising avenue toward a new and more balanced ecosystem management strategy for the boreal forest.
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