Download Bengtsson, J., Nilsson, SG, Franc, A., and Menozzi, P. (2000).

Bengtsson, J., Nilsson, S. G., Franc, A., and Menozzi, P. (2000). Biodiversity, disturbances,
ecosystem function and management of European forests. Forest Ecology and Management 132:
39-50.
Keywords: 78Eur/biodiversity/disturbance/ecosystem/ecosystem function/forest/forest
dynamics/forestry/herbivores/management
Abstract: We review the effects of human impact on biodiversity of European forests in the light of
recent views on disturbances and succession in ecosystems, and discuss recent ideas on how
biodiversity affects ecosystem functions such as productivity and ecosystem stability. With this as
a background we discuss how to better manage European forests for both production and
biodiversity. We argue that the next generation of forestry practices need to understand and
mimic natural disturbance dynamics much better than the present ones. Of particular importance
is the fact that most species in European forests have evolved in forests that were to a large
extent influenced by large grazers, first by megaherbivores and later, in historic times, by
domestic animals. We highlight several areas where new knowledge and management tools are
urgently needed: (i) How do species survive and adapt to the natural disturbance regimes in
different regions and forest types? (ii) How can new and imaginative forest management
practices be devised that take natural disturbance regimes into account? (iii) How does forest
biodiversity affect ecosystem function and stability in a changing world, in particular in the light of
predicted climate changes? (iv) How are ecological processes at different levels and scales
related to diversity, and how do different management practices affect biodiversity? (v) How can
efficient agroforestry methods be developed to preserve biodiversity? (vi) What is the role of
humans and human behaviour for sustainable management of ecosystems?
Forest Ecology and Management 132 (2000) 39±50
Biodiversity, disturbances, ecosystem function and
management of European forests
Jan Bengtssona,*, Sven G. Nilssonb, Alain Francc, Paolo Menozzid
a
Department of Ecology and Environmental Research, SLU, Box 7072, S-750 07 Uppsala, Sweden
b
Ecology Building, Lund University, S-223 62 Lund, Sweden
c
ENGREF MAI, 19 Avenue du Maine, F-75 732 Paris Cedex 15, France
d
Department of Environmental Sciences, University of Parma, Parma, Italy
Abstract
We review the effects of human impact on biodiversity of European forests in the light of recent views on disturbances and
succession in ecosystems, and discuss recent ideas on how biodiversity affects ecosystem functions such as productivity and
ecosystem stability. With this as a background we discuss how to better manage European forests for both production and
biodiversity. We argue that the next generation of forestry practices need to understand and mimic natural disturbance
dynamics much better than the present ones. Of particular importance is the fact that most species in European forests have
evolved in forests that were to a large extent in¯uenced by large grazers, ®rst by megaherbivores and later, in historic times, by
domestic animals. We highlight several areas where new knowledge and management tools are urgently needed: (i) How do
species survive and adapt to the natural disturbance regimes in different regions and forest types? (ii) How can new and
imaginative forest management practices be devised that take natural disturbance regimes into account? (iii) How does
forest biodiversity affect ecosystem function and stability in a changing world, in particular in the light of predicted climate
changes? (iv) How are ecological processes at different levels and scales related to diversity, and how do different management
practices affect biodiversity? (v) How can ef®cient agroforestry methods be developed to preserve biodiversity? (vi) What
is the role of humans and human behaviour for sustainable management of ecosystems? # 2000 Elsevier Science B.V. All
rights reserved.
Keywords: Biodiversity; Disturbances; European forests; Natural forest dynamics; Ecosystem function; Forest management; Succession;
Renewal cycle; Grazers; Megaherbivores; Stability and insurance hypothesis
1. Introduction
The natural vegetation in most of Europe is forest,
from the Mediterranean through the deciduous forests
*
Corresponding author. Present address: Department of Ecology
and Crop Production Science, SLU, Box 7043, S-750 07 Uppsala,
Sweden. Fax: ‡46-18-672890.
E-mail address: [email protected] (J. Bengtsson)
of Central and Western Europe to the boreal forests in
Fennoscandia. Many of these forests are now long
gone due to human activities, and some European
forest types belong to the most endangered ecosystems in the world. For example, it has been estimated
that only 0.2% of the Central European deciduous
forests remains in a relatively natural state (Hannah et
al., 1995), making it and the species living therein as
threatened as more recognised endangered ecosystems
0378-1127/00/$ ± see front matter # 2000 Elsevier Science B.V. All rights reserved.
PII: S 0 3 7 8 - 1 1 2 7 ( 0 0 ) 0 0 3 7 8 - 9
40
J. Bengtsson et al. / Forest Ecology and Management 132 (2000) 39±50
such as tropical rainforest. A long history of human
land-use changes has created the European forests that
we know today (e.g. Perlin, 1988). The hunting of
large herbivores and predators, the grazing of domestic herbivores, the clearing of forests for agriculture,
®rewood and industrial purposes, extraction of litter
for fertilising purposes, and more recently the planting
of monocultures and exotic species are examples of
such human impacts. The Mediterranean forests have
been altered by human activities for a very long time
and practically no pristine forest appears to be left in
the region. The generally high human population
densities have led to intensive cultivation practices
like coppice with short rotations. Along with overgrazing and frequent ®res this has resulted in severely
degraded forests (often given a speci®c local name)
becoming common (Di Castri et al., 1981). This
situation makes the remaining pristine or close to
pristine forests even more valuable (Stanners and
Bourdeau, 1995).
Traditionally, ecosystems left without human intervention and disturbances were supposed to develop
towards stable `climax' systems (e.g. Clements, 1936;
Odum, 1969). This static view of nature has now given
way for a more dynamic one in which change and
disturbance are seen as natural features of ecosystems
(e.g. White, 1979; Pickett and White, 1985; Botkin,
1990; Holling, 1992; Holling et al., 1995; Ellenberg,
1996). Most species in European forests have evolved
under and adapted to past natural disturbance regimes.
Therefore, an understanding of former natural disturbance dynamics and how they relate to human disturbances and management practices is essential to
preserve and manage biodiversity as well as ecosystem functions in the present human-dominated European forests (Nilsson and Ericson, 1997).
During the last decade, studies and discussions of
biological diversity have emerged from the obscurity
of basic scienti®c discourse to become one of the most
pressing and relevant subjects in environmental policy.
This volte-face has been met with mixed feelings both
in the scienti®c community and in forestry. While
many ecologists are happy that ecological issues
become part of the political agenda, they also have
problems in coming out of academia to become part of
the real world of resource management, economy and
politics. In forestry, many foresters have had problems
accepting that biodiversity, which previously had no
economic value, should affect how they manage forests. However, market pressure requiring forestry to
manage biodiversity quickly made forestry pay attention to the preservation of diversity.
The concern over declining biological diversity has
several roots, and only some of them are part of what is
normally regarded as science. One of the main ecological arguments for preserving biodiversity is that
the losses of diversity may impair the life-supporting
processes that humans need, i.e. ecosystem functions
such as primary productivity, carbon storage, water
retention and provision of clean water. Diversity may
also entail ecological stability. Thus, maintaining
diversity may be necessary for long-term sustainability. These are old unsolved questions in ecology, and
the recent debates have not resolved them (see below).
A major force behind the Biodiversity Convention in
Rio 1992 was that many developing countries wanted
some degree of control and gains from the largely
unexplored biological diversity in the tropics Ð an
economic rather than an ecological argument for
preserving biodiversity. Other reasons to preserve
biodiversity are ethical, and the pleasure many people
experience from a high diversity in their environment.
In the following, we offer a personal perspective on
some of the issues that we believe are of major
importance for preserving and enhancing biodiversity
in European forests. The paper is not intended as a
review of the extensive literature on the subject (see
e.g. Hansson, 1997; NiemelaÈ, 1997). Rather, it should
be seen as part of an ongoing discussion, and many of
the issues we raise are subjects of current debate and
disagreement. Since biodiversity conservation priorities tend to focus on species diversity, this will be the
focus of our paper, but we acknowledge that in the
long term the conservation of diversity also involves
conservation of genetic diversity.
2. Humans and biodiversity in European forests
Ð a historical perspective
We do not know exactly what the European forests
looked like before humans became important agents
of landscape change (Rackham, 1998). However, large
megaherbivores such as the forest elephant and rhinoceros were probably important in in¯uencing forest
structure in many areas. Most trees have adapted and
J. Bengtsson et al. / Forest Ecology and Management 132 (2000) 39±50
evolved in response to unselective and fairly intense
grazing pressure by megaherbivores (Owen-Smith,
1987). It is likely that grazing resulted in a forest
with more openings and gaps, a higher abundance of
large trees and less saplings than the present `virgin
forests'. Otherwise it is dif®cult to understand why a
majority of forest species, e.g. among insects and
lichens, prefer or require a half-open forest (e.g.
Andersson and Appelkvist, 1990; Nilsson and Ericson, 1997).1 Some pasture woodlands, like the New
Forest in England and the Dehesha woodland in Spain,
are probably more similar to the forests where most
forest organisms evolved than dense ungrazed stands.
Fire has also been an important disturbance factor in
many forests, especially in the taiga forests of the
boreal zone (Zackrisson, 1977) and in the Mediterranean, but not in Central Europe.
In addition to short-term disturbances, long-term
changes such as glacial cycles have in¯uenced European forests. The refuge areas in south-eastern,
southern and south-western Europe support the highest diversity and contain the species with most
restricted continental ranges. This is clearly born
out by some well-known organisms such as longhorn
beetles, Cerambycidae (Bense, 1995). In addition, as
predicted by population genetic theory (Hewit, 1996),
high genetic diversity has been found in the refugia
from which recolonization waves were started. Good
examples of this are oak species in which a higher
diversity of cpDNA haplotypes was found in southern
refugia in Italy, the Iberian peninsula and the Balkans
(Dumolin-Lapegue et al., 1997), and beech which
within Italy shows a higher allozyme diversity in
the glacial refuge area of Sicily (Leonardi and
Menozzi, 1995). These refuge areas are essential for
the preservation of the biodiversity of the forests of
Europe.
At least during the last 4000±5000 years humans
have been important agents affecting landscape composition and dynamics and biological diversity in
Europe. Many large predators and herbivores have
gone regionally extinct since Roman times. As these
1
One of the reviewers questioned our emphasis on the
importance of megaherbivores. We regard the evidence for
substantial megaherbivore impacts in many pre-historical forests
as convincing, but acknowledge that the issue is still partly
unsolved.
41
large animals disappeared domestic animals took the
place of the herbivores. The domestic grazers have
affected most forests in Europe for centuries to millennia. Relative to the megaherbivores, smaller grazers seem to be more selective in diet than the extinct
ones, resulting in a changed species composition of the
forests, although several key structures of the original
forests, such as small open glades and large old trees,
were retained (Peterken, 1996; Nilsson, 1997). The
cover of forest has been ¯uctuating as human populations have gone down and up, mainly in conjunction
with plagues and with industrialism and trade, respectively. The last century has seen the most rapid landscape changes, as technological advances have made
agriculture capable to utilise more land, and the
remaining marginal areas have gradually been converted to managed forests. During this process, many
features of old forest and old agricultural landscapes
such as dead wood, old trees, ®re and natural grazed
areas have declined or been lost over large areas. This
has had large negative consequences for biodiversity.
We know that a high proportion of red-listed (threatened) species in Scandinavia, Britain and Central
Europe are dependent on these structures and processes (e.g. Berg et al., 1994, 1995), but such knowledge is not available for all Europe. In the
Mediterranean, applications of the standard proposed
for red-listing species are dif®cult because of the lack
of information on the rarity and fragility of local
species. This has led to the red-lists containing species
that are rare in central European forests but relatively
common in the Mediterranean region, and the overlooking of taxa absent at more northern latitudes but
threatened in the Mediterranean.
The large-scale changes in landscape structure have
in¯uenced diversity both positively and negatively.
The opening up of landscapes by human activities has
probably led to increased diversity during prehistoric
and historical times, although the changes in this
century have mainly resulted in diversity losses. For
example, many species adapted to open grazed forests
have disappeared. However, there are exceptions.
Many bird species, primarily those adapted to scrub
habitats, have increased temporarily as marginal agricultural land and grazed forests have been abandoned
and are undergoing succession towards high forest.
Also, many exotic species have become naturalised. In
the short term diversity has often increased at the local
42
J. Bengtsson et al. / Forest Ecology and Management 132 (2000) 39±50
and national (regional) levels, while over evolutionary
time it may result in a lower global diversity.
Unfortunately, neither ecologists nor foresters have
had the tools or the knowledge to examine the consequences of large-scale landscape changes on diversity, as most research has been conducted on relatively
small scales, e.g. experimental plots. Research from
small scale experiments may not be relevant to predict
large-scale dynamics of ecosystems. For example,
among soil fauna, responses of groups such as springtails and mites which have population dynamics on a
smaller spatial scale than the normal size of plots (ca.
102 m2), can probably be extrapolated to larger areas.
This may not be the case for more mobile animal
groups such as Dipterans, where responses at the
spatial scale of the plot re¯ects individual movements
and egg-laying preferences rather than population
dynamics (e.g. Bengtsson et al., 1997b). Recently
developed research areas such as landscape ecology,
metapopulation biology, and analyses of large-scale
and long-term patterns in communities may yield
better insights into these questions in the future,
and are imperative for a better management of forest
resources with respect to biodiversity (e.g. Siitonen
and Martikainen, 1994; Kareiva and Wennergren,
1995; Angelstam, 1997).
3. Changing views on stability and change in
natural ecosystems
For a long time, ecologists and especially people in
conservation organisations viewed nature as rather
static. Without human intervention and disturbances,
ecosystems were supposed to develop towards `climax' systems which were stable and in which biomass
and nutrients had accumulated (e.g. Clements, 1936;
Odum, 1969). Although there are some examples from
boreal forests that this may actually happen, the static
view of nature has given way for a more dynamic one
in which change and disturbance are regarded as
integral features of ecosystems (e.g. White, 1979;
Botkin, 1990; Holling, 1992; Holling et al., 1995).
The dynamic view of nature has its roots in discussions in vegetation ecology in the beginning of the
20th century. It is useful to contrast two opposing
views. The ®rst one, pioneered by (Clements, 1916,
1936), proposed that the different components of
vegetation communities were almost like parts of an
organism. The equilibrium point was called climax,
and it was mainly the result of climatic factors. Any
discrepancy from the supposed climax was viewed as
a disturbance. This view was rather static, as climaxes
theoretically might last for a very long time if no
disturbance did occur. The other view was that of
Gleason (1926), who proposed that individuals within
the vegetation were adapting independently to similar
conditions, and the vegetation cover should be more or
less seen as a collection of non-interacting plants that
happened to be adapted to the same site conditions.
This view is not very far from a standard Darwinian
view of communities, where selection at the individual
level (and not the species or community level) is
regarded to be the rule. The problem evolved with
Tansley's seminal paper on the ecosystem (Tansley,
1935). After a while, most researchers acknowledged
that vegetation is never locally stable, with disturbances and succession as the major triggers of change.
Succession is thus of importance for understanding
the dynamics of natural ecosystems. Other important
processes are disturbances and factors affecting how
organisms survive in temporally changing mosaic
landscapes, e.g. dispersal. Holling (e.g. Holling,
1992; Holling et al., 1995) has formulated the modern
dynamic ideas on succession and disturbance in a
simple conceptual cyclic model of ecosystems
(Fig. 1). Several things are important here: According
to this view, ecosystems alternate between periods of
weak interactions between organisms (phases 1,
exploitation, and 4, reorganisation) and periods when
species interact strongly (phases 2, conservation, and
3, release). On a small spatial scale, the system is everchanging, but on a larger scale the shifting mosaic of
patches is more stable. Different processes are important to maintain the system during the transitions
between the phases: succession during the slow transition from phases 1±2, rapid disturbances such as ®re
or pests in the transition from 2 to 3, and dispersal
during 4 to 1. These ideas were linked to the biodiversity issue by Peterson et al. (1998). Biological
diversity is especially important during the succession
and reorganisation phases, because a high regional
diversity means that there will be an appropriate pool
of species to continue the cycle. Some disturbances
can carry the ecosystem into different stability
domains (Fig. 1), essentially ¯ipping the present
J. Bengtsson et al. / Forest Ecology and Management 132 (2000) 39±50
43
Fig. 1. Succession and re-organization of ecosystems as viewed by recent theory on change in ecosystems. Arrows close to each other indicate
rapid changes, far from each other slow changes. The thick arrow to the left indicates the possible exit from the cycle, and that the change to a
new system (degraded or more productive) is most likely at the stage between re-organization and exploitation. Diversity in forested
landscapes is needed for the transitions between all the four stages, which is why managed forests must be managed so that a sufficient number
of species is available for the cycle to continue (modified after Holling et al., 1995).
system to a new one, and the likelihood for this
increases as species disappear regionally.
4. Natural forest dynamics
Different natural forests are characterised by different types of natural dynamics, because the scale of
disturbances vary much between forest types. Taiga
forests in northern Europe were often disturbed by
large-scale ®res (Zackrisson, 1977), while nemoral
deciduous forests, the natural vegetation of most of
central and western Europe, are mostly disturbed by
small-scaled windthrows (Falinski, 1986). In the wide
transition zone both large-scale and small-scale disturbances occur in different types of forests (Nilsson,
1997). These different types of disturbances have
major implications for conservation of diversity and
for sustainable forestry (Nilsson and Ericson, 1997).
Recent studies in boreal forest (Wardle et al., 1997)
indicates that ®re frequency can have important effects
on the forest ecosystem, in¯uencing vegetation
dynamics, diversity patterns and ecosystem processes.
Fire is also a particularly relevant ecological factor for
Mediterranean forests. It has been estimated that 0.5±
1% of Mediterranean forests are affected by ®re every
year (Di Castri et al., 1981).
It is reasonable to assume that organisms in natural
forests are adapted to the characteristic disturbance
regimes of these forests. In addition, there are forest
species with adaptations that happen to make them
favoured by the human-induced landscape changes in
historical times Ð for example, open glades created
by small-scaled forestry and domestic herbivores.
As discussed above, it is likely that over evolutionary
time large herbivores have been an important factor
in forests. Therefore, the continuation of some
traditional use of self-regeneration forests may be
important to maintain their biodiversity, because
it to some degree mimics the megaherbivore grazing
that the species were originally adapted to. The
megaherbivores contributed to several structures
that nowadays are becoming rare and need to be
retained in managed forests. An example may be that
they created wounds on large branches and trunks
resulting in barkfree wood on living trees. Many
insects are dependent on such trees with sun exposure
(Nilsson and Ericson, 1997). Also, standing dead
44
J. Bengtsson et al. / Forest Ecology and Management 132 (2000) 39±50
or dying trees were much more common than they are
now.
5. Forest management practices and natural
forest dynamics
Since most forestry practices involve disturbances
to forest ecosystems, sometimes severe ones, forestry
will inevitably have effects on both diversity and on
ecosystem function. The preservation of biodiversity
in managed forests requires a consideration of how
forestry practices such as clearcutting or selective
cutting correspond to natural disturbances and natural
forest dynamics (e.g. Haila et al., 1994; Angelstam,
1997; NiemelaÈ, 1997). It is sometimes argued by the
forest sector that, for example, clearcutting is almost
equivalent to ®re or windthrows and thus not very
`unnatural'. This is misleading, if not completely
wrong (Essen et al., 1997). The similarities between
®re or windthrows and clearcutting are super®cial.
Many features that are present after ®re and other
natural disturbances are lost when clearcutting. An
important task for a sustainable forest management
will be to approach natural forest dynamics (NiemelaÈ,
1997). Comparisons with the few large natural forest
areas in Europe, such as eastern Poland or northeastern Russia, have led to important insights, and
will continue to do so.
Such comparisons have revealed that several features and structures of natural forests decline or disappear in managed forests. Old-growth forest features
with old and large trees, both standing, living, dying
and laying are especially prone to disappear (e.g.
Peterken, 1996). Some such structures are obvious,
such as old logs, while some are often overlooked. For
example, in old-growth forests strong winds and
megaherbivores often broke branches leading to
wounds on large branches and trunks of living trees.
Recent studies indicate that several species are dependent such hollow trees (e.g. Martin, 1989; Nilsson and
Baranowski, 1997). Other features of most natural
forests often lacking in managed forests is burned
wood, open glades created by grazers, deciduous trees
Ð in particular old ones Ð and a mixture of tree
species and of tree ages.
In many areas, the continuation of some traditional
use of self-regeneration forests, such as grazing and
selective cutting, may be important in maintaining
their diversity, although it does not automatically
follow that all traditional practices favour biodiversity.
Activities similar to the effects of natural disturbances
and grazing by large herbivores could (and should) be
simulated by forestry, while other traditional practices
such as litter-raking may be destructive. It should also
be pointed out that in Mediterranean forests overgrazing is still a problem at many places. This use
of forests is different from the northern ones which
suffer from excessive spatial homogeneity. In these
Mediterranean cases grazing should be reduced to
allow a recovery of degraded forests already threatened by other disturbances such as ®res and water
stress.
Not all species are equally affected by forestry.
Some species, primarily generalists, are often little
or even positively affected. Many of these species
disperse easily and are less sensitive to fragmentation
as long as their habitat requirements are ful®lled. Such
species are often found in the formerly glaciated areas
in Russia, Finland and Scandinavia. Other species are
severely affected by fragmentation and habitat loss,
for example specialist species requiring old-growth
forest or old trees. These are often poor dispersers and
very sensitive to forestry disturbances, fragmentation
and isolation of habitat patches. Among such species
is a rich ¯ora and fauna connected to ancient trees of
nemoral deciduous species (e.g. Rose, 1976; Speight,
1989). The most extreme of these are a number of
insect species found only in small refuges in humandominated landscapes in SW and SE Europe (e.g.
Dajoz, 1966). For such species as well as for oldgrowth forest species with particular habitat requirements it is imperative to identify and preserve forest
areas with long continuity over time (Peterken, 1996;
Nilsson and Ericson, 1997). Thus, many species may
require species- or group-speci®c management practices to be preserved in today's forests. Some species
may need large areas of long continuity, while for
others only small unmanaged patches or even single
trees in managed forests may be required. Simpli®ed
forestry practices that are good for some species may
not be appropriate for others. In order to protect
diversity in the long perspective, the spatio-temporal
dynamics of both species and disturbance regimes at
several scales in forest landscapes need to be understood.
J. Bengtsson et al. / Forest Ecology and Management 132 (2000) 39±50
6. Effects of biodiversity on ecosystem function
The projected decline in biodiversity has concerned
ecologists for several reasons: species diversity may
have direct positive effects on important ecosystem
processes such as productivity and nutrient cycling
(e.g. Tilman et al., 1996). A high diversity may also
entail ecosystem stability, for example, resistance to
perturbations, resilience after disturbances and more
stable and reliable ecosystem functions over time (e.g.
Peterson et al., 1998; Naeem, 1998). The major longterm importance of diversity may be as a source of
species capable of performing desired ecosystem
functions if the present species should disappear,
for example because of climate change (the `insurance' hypothesis; Folke et al., 1996).
6.1. Diversity and ecosystem stability
The idea that diversity is related to ecosystem
stability has a long history in ecology (see, e.g. Pimm,
1991). Until the early 1970s, this relationship was
usually regarded as positive, the argument being that
more diverse and complex ecosystems have more
possible pathways for energy and nutrient ¯ows than
species-poor ones. Note that this argument is essentially one about stability in ecosystem processes, not
in population numbers or species composition. May
(1973) argued the opposite. Based on the stability
properties of multi-species model communities with
strong species interactions, he argued that diverse
communities were less stable than simple ones. May's
argument was about species abundances, not ecosystem processes. After hot discussions, the tide seems to
have turned once again. Polis (1994) put this succinctly when he wrote ``. . . recent analyses of food
webs pound yet another nail (hopefully the ®nal one)
into the cof®n of May's paradox. . . . I am attracted to
the position that complexity promotes stability and
conclusions from dynamic models are simply wrong.''
Thus, more diverse systems may be more resistant to
perturbations and more resilient than species-poor
ones, as suggested for plant communities by Tilman
and Downing (1994), although this study has been
heavily criticised. Nonetheless, although there are
some exceptions, recent research has largely supported this view (McNaughton, 1993; McGrady-Steed
et al., 1997; Naeem and Li, 1997; see below), restating
45
the `ecological intuition' of the times before the 1970s
that diverse ecosystems are more stable.
6.2. Diversity and ecosystem processes
Although many ecologists agree that diversity often
is important for the rate of ecosystem processes such
as productivity or decomposition rate, this has until
recently mainly been based on informed guesses and
common sense arguments. The few experiments on
these issues are ambiguous. For example, Tilman et al.
(1996) showed positive effects of diversity on productivity, while e.g. Rusch and Oesterheld (1997) claimed
the opposite. Sometimes, species diversity per se is not
be as important as the presence of several functional
groups (Hooper and Vitousek, 1997; Tilman et al.,
1997; Laakso and SetaÈlaÈ, 1999). It is unclear if these
results can be extrapolated from the relatively speciespoor experimental systems to natural species-rich
ecosystems.
However, there are some general patterns worth
noticing. Often, it seems that there are advantages
with more diverse ecosystems, at least compared to the
monocultures found in forestry and agriculture. For
example, species mixtures may provide a higher yield
than monocultures (e.g. spruce and birch, Burkhart
and Tham, 1992) and several experiments in grasslands show that average productivity increases as
species richness increases from 1 to ca. 10 species.
However, often it is possible to ®nd some single
species that under some circumstances do as well as
mixtures, as witnessed by the widespread use of
monocultures in forestry and agriculture. Other interesting observations are that species mixtures appear to
have lower levels of pest damage and higher resistance
to invaders and weeds (McGrady-Steed et al., 1997; S.
Naeem in Bengtsson et al., 1997a). In forests, a lower
frequency of insect damage and stem rot have been
found in mixed forests compared to monocultures (Su
et al., 1996; Gerlach et al., 1997).
In agroecosystems, populations with a low genetic
diversity (monocultures) have sometimes Ð but not
always Ð been associated with higher levels of
damage from pests (e.g. Power and Kareiva, 1990).
Yields of monocultures are also more susceptible to
environmental vagaries. Such effects may on a longer
time-scale be important also in forestry and deserve
more attention.
46
J. Bengtsson et al. / Forest Ecology and Management 132 (2000) 39±50
However, the direct positive effects of diversity on
ecosystem functions such as productivity may only be
relevant at the low end of a diversity gradient, i.e.
when we go from monocultures to few-species mixtures, and sometimes not even there (see e.g. Bengtsson et al., 1997a). A more generally relevant and
important advantage of retaining diverse ecosystems
is the so-called `insurance hypothesis', which suggests
that high diversity leads to more reliable ecosystem
functions over time when environmental conditions
vary (Naeem, 1998), e.g. because of climate change.
This hypothesis is extremely dif®cult to test in the
®eld, but there is some support from studies of
zooplankton in lakes (Frost et al., 1995). Recently,
two laboratory studies using unicellular protists
have supported the insurance hypothesis (McGradySteed et al., 1997; Naeem and Li, 1997), although
these still concern the low end of the diversity
gradient. The insurance hypothesis may thus provide
one of the most convincing arguments for preserving
biodiversity, as suggested by e.g. Bengtsson et al.
(1997a).
If the goal is to maintain and create sustainable
forest ecosystems, the most crucial role for diversity is
as a supplier of species performing functions and
conferring stability (Bengtsson, 1998). Declines in
diversity means losses of species including key species, which in turn may have a number of undesirable
consequences: Changes in food webs, loss of specialist functions, and loss of ecosystem engineers, keystone species and other species or functional groups.
Present ecological theory is not ready to address the
effects of these changes on ecosystem function,
although recent theoretical advances may swiftly lead
to applications in forestry. For example, Bengtsson's
research has focused on the impact of intensive harvesting on soil food webs, and how this affects carbon
and nitrogen cycling in theory and practice (Zheng
et al., 1997, 1999; Bengtsson et al., 1997b). Extensions of such analyses incorporating diversity changes
may be directly relevant to forest management.
Another example is the suggestion from food web
theory that consumers sometimes can speed up nutrient cycling and thus be positive for ecosystem productivity (Loreau, 1995). If these types of analyses
proceed to incorporate changes in diversity as well as
changes in species identities they may become useful
tools in forest management.
Developments in the analysis of the role of diversity
in species-rich and complex communities may shed
further light on the importance of biodiversity for
ecosystem function. It has been suggested that focus
on the number of species, being taxonomic entities,
may not be the most ef®cient way of examining
this problem. Instead, it could be more fruitful to
concentrate on the distribution of ecological traits
in a community, for example the distributions of
resource acquisition ef®ciencies or tolerances to
environmental stress across species. Of course, to
some degree these distributions are likely to be related
to species diversity. More importantly, different
distributions probably re¯ect differences between
communities in ecosystem function or species complementarity. Focus on trait distributions rather than
species numbers is potentially more relevant for
understanding ecosystem function and its relation
to biodiversity.
7. Implications for forest management
Although expanded nature reserve systems will be
essential in managing forests for sustainability at the
landscape level, reserves alone are not suf®cient to
preserve the biodiversity and dynamics of European
forests. It will always be necessary to manage most
forest land for both production and biodiversity, even
with an expanded reserve system. Although some
minor part of the forest area could possibly be used
for very intensive production, it is naõÈve to argue that
splitting forests into reserves and high-intensity forestry areas will preserve even a fraction of the biodiversity of European forests. The area of existing and
potential reserves is simply too small. For example, in
Scandinavia forest reserves constitute only 1±3% of
the productive forest area (Essen et al., 1997), and
there are much too few large stands >100 years old that
could become part of larger reserve systems. Therefore, we strongly argue that the only way to achieve
sustainability of European forests will require that the
major part of forest land is managed by methods that
are acceptable from the point of view of the environment and of biodiversity, as well as timber yield (e.g.
NiemelaÈ, 1997). This means that there will be a strong
need and incentive for new forestry methods and
innovative technology.
J. Bengtsson et al. / Forest Ecology and Management 132 (2000) 39±50
To retain biodiversity during forestry operations, it
is ®rst necessary to identify stands with a long continuity. Such forests are the dispersal sources to the
surrounding forests for many slowly dispersing organisms, e.g. many snails, insects and lichens. These
refuge forests make up a decreasing and very small
proportion of the forests in Europe, but they are
essential for restoring biodiversity in the surrounding
landscape. In most cases they should be left unmanaged, while in others non-native species invading
from planted forests must be removed. In all cases
suitable habitat should be present in the near surroundings for maintaining viable populations in the long
run. Furthermore, old trees and coarse woody debris
must be allowed to develop in managed forest. Such
trees may also be useful for forestry as wind breaks
along edges to open habitats.
The next generation of forestry practices need to
more thoroughly understand and mimic natural disturbance processes and dynamics than the present
ones. For example, present methods of large-scale
clearcutting need to be reconsidered in the light of
recent advances in succession theory and conservation
biology. The recognition of the importance of large
herbivores for diversity in many forests requires attention during forestry operations. Recent studies showing the problems with monocultures in forestry imply
that we have to reconsider the present dominating
forestry methods in Europe. This is in fact already
happening in, for example, Germany where mixedspecies forestry is often practised nowadays. We have
to develop practices and technics that are adapted to
more diverse forests. Locally, there have always been
such methods in Europe, but few have been scienti®cally evaluated and improved. The development of
new technology is unlikely to be a big problem in
forestry. Rather, the most crucial gaps in ecological
knowledge need to be identi®ed and made the subject
of research (see below), and the results communicated
to the forest sector.
However, some species, such as those dependent on
old-growth forest, will not be able to survive in
managed forests. For these species nature reserves
are necessary, and forestry will have to accept that
reserves or other set-aside areas are an integral part of
land management. In some parts of Europe, such as in
the high-diversity glacial refuge areas, large reserves
are especially important. Besides preserving rare and
47
threatened species, reserves may be useful because
they have more natural forest dynamics and thus act as
source areas for managed forests, increasing landscape resilience (Folke et al., 1996).
8. Identification of key areas for future research
1. How does what we know and may learn about
natural forest dynamics and ecosystem processes
affect management practices? An understanding
of natural forest dynamics is a prerequisite for
sustainable forest management. For example, most
natural forests in Europe are composed of several
or many tree species and with varying structure
and ages of trees, but forest management is too
often focused on even-aged monocultures. How
can this be changed?
Firstly, it is crucial to understand the natural
disturbance regimes in different regions and forest
types, and how species survive and adapt to these
regimes. Secondly, imaginative and innovative new
forest management practices that take these regimes
into account need to be devised. For instance, recent
research on mycorrhizal fungi (KaÊreÂn, 1997) suggests that with appropriate management practices,
more species survive the harvesting phases and the
mycorrhizal communities in middle-aged stands
become more similar to old-growth stands. These
species may be required for efficient ecosystem
function (although this is not yet established). If
forestry practices can be changed to incorporate
knowledge on these issues, there will be a great
potential for a more sustainable forest management
for both production and biodiversity.
2. How does biodiversity affect ecosystem function
and stability in a changing world, in particular in
the light of predicted changes in climate? This is a
key issue that involves (i) establishing which are
the present key species and functional groups, and
(ii) an understanding of the possible roles of rare
species in ecosystems. The issues must be
addressed in long-term experiments Ð the present
funding on a 3-year basis does not provide the
incentives or facilities for such long-term research.
In particular, a concerted action to establish field
tests of the insurance hypothesis in forests would
be appropriate.
48
J. Bengtsson et al. / Forest Ecology and Management 132 (2000) 39±50
3. How are ecological processes at different levels
and scales related to diversity? The classical
questions in ecology concerning the mechanisms
behind different levels of diversity have not yet
received satisfactory answers. In this sense,
different management practices can be regarded
as ongoing ecological experiments that are useful
to elucidate diversity patterns and suggest mechanisms that could be examined in more detail.
This potential is clearly under-utilised. However, a
major problem with this approach is that proper
controls in extensive natural forests must be
established. In view of this it is tragic that the
largest area of Central European mixed deciduous
forest in Bialowieza in Poland is still diminishing
in area.
4. There is also a clear need to develop efficient
agroforestry methods. As argued above, pasture
woodlands may simulate structures and processes
that have been important when megaherbivores
dominated European forests, when most present
forest organisms acquired their adaptations.
Therefore, we suggest that production of wood,
meat and other products in a half-open forest
could be the most sustainable use of forests in
many parts of Europe. By tradition, forestry and
agricultural research have been different branches
and this has clearly hindered research efforts in
agroforestry. Organisational changes are necessary
to rectify this.
5. An issue that recently has been discussed is the
role of humans and human behaviour for sustainable management of ecosystems. Previous views
that economy rules over ecological and environmental considerations have now been challenged.
It has been shown that environmental costs have
seldomly been taken into account, but if they were
this would change the rationality of many
decisions in, e.g., forestry. Interactions between
ecologists, foresters, economist and social scientists may be of great value, in particular as it
becoming important to estimate the economic and
societal values of ecosystem functions and biodiversity (e.g. Perrings et al., 1995; Folke et al.,
1996; Bengtsson et al., 1997a).
To examine questions such as these, it is clear that we
need to use a variety of approaches from studies of
population dynamics of different key species, mechanisms for diversity patterns, and studies of ecosystem
processes affecting forest functioning for humans.
Also, the previous small-scale studies in forestry need
to be complemented with studies at the landscape
level, something that is already done in several European countries (e.g. Angelstam, 1997). However, an
appropriate understanding of biodiversity in forest
ecosystems, and how to use biodiversity for the sustainable use of forest resources, requires linkages
between all these approaches.
Acknowledgements
Ê ke Berg and two reviewers for comWe thank A
ments on the manuscript. JB thanks Carl Folke for
discussing several of the ideas that ®nally made their
way into the paper. The idea that trait distributions
might be more fruitful than species numbers was
discussed at a TERI workshop at Silwood Park,
UK, in 1997. The travels during the formation of
the paper were funded by the EU-FAIR programme
EFERN S6. Our research has been funded by, among
others, NFR-Sweden (JB), EU (through TERI/Globis;
JB), NUTEK-Sweden (JB), and MISTRA/SUFORSweden (SGN).
References
Andersson, L., Appelkvist, T., 1990. Megaherbivores of the Ice
Age formed the nemoral and boreo-nemoral ecosystems
(Istidens stora vaÈxtaÈtare utformade de nemorala och boreonemorala ekosystemen). Svensk Botanisk Tidskrift 84, pp. 355±
368 (in Swedish with English summary).
Angelstam, P., 1997. Landscape analysis as a tool for the scientific
management of biodiversity. Ecol. Bull. 46, 140±170.
Bengtsson, J., 1998. Which species? What kind of diversity?
Which ecosystem function? Some problems in studies of
relations between biodiversity and ecosystem function. Appl.
Soil Ecol. 10, 191±199.
Bengtsson, J., Jones, T.H., SetaÈlaÈ, H., 1997a. The value of
biodiversity. Trends Ecol. Evol. 12, 334±336.
Bengtsson, J., Persson, T., Lundkvist, H., 1997b. Long-term effects
of logging residue addition and removal on macroarthropods
and enchytraeids. J. Appl. Ecol. 34, 1014±1022.
Bense, U., 1995. Longhorn beetles. Margraf Verlag, Weikersheim,
Germany.
Ê ., EhnstroÈm, B., Gustafsson, L., HallingbaÈck, T., Jonsell,
Berg, A
M., Weslien, J., 1994. Threatened plant, animal and fungus
J. Bengtsson et al. / Forest Ecology and Management 132 (2000) 39±50
species in Swedish forests: distribution and habitat associations.
Conserv. Biol. 8, 718±731.
Ê ., EhnstroÈm, B., Gustafsson, L., HallingbaÈck, T., Jonsell,
Berg, A
M., Weslien, J., 1995. Threat levels and threat to red-listed
species in Swedish forests. Conserv. Biol. 9, 1629±1633.
Botkin, D.B., 1990. Discordant harmonies. Oxford University
Press, Oxford, UK.
Ê ., 1992. Predictions from growth and yield
Burkhart, H.E., Tham, A
models of the performance of mixed-species stands. In:
Cannell, M.G.R., Malcolm, D.C., Robertson, P.A. (Eds.), The
Ecology of Mixed-species Stands of Trees. Blackwell, Oxford,
pp. 21±34.
Clements, F.E., 1916. Plant succession. Publication No. 242.
Carnegie Institution, Washington DC.
Clements, F.E., 1936. Nature and structure of the climax. J. Ecol.
24, 252±284.
Di Castri, F., Godall, D.W., Specht, R.L., 1981. Mediterranean type
shrublands. Elsevier, Amsterdam.
Dajoz, R., 1966. EÂcologie et biologie des coleÂopteÁres xylophages
de la heÃtraie. Vie et Milieu 17, 1±763.
Dumolin-Lapegue, S., Demesure, B., Fineschi, S., Le Corre, V.,
Petit, R.J., 1997. Phylogeographic structure of white oaks
throughout the European continent. Genetics 146, 1475±1487.
Ellenberg, H., 1996. Die Vegetation Mitteleuropas mit den Alpen in
È kologisches, historisches und dynamischer Sicht. Ulmer
O
Verlag, Stuttgart, (5 Aufl.).
Essen, P.-A., EhnstroÈm, B., Ericson, L., SjoÈberg, K., 1997. Boreal
forests. Ecol. Bull. 46, 16±47.
Falinski, J.B., 1986. Vegetation dynamics in temperate lowland
primeval forest. Ecological studies in Bialowieza forest.
Geobotany 8, 1±537.
Folke, C., Holling, C.S., Perrings, C., 1996. Biological diversity,
ecosystems and the human scale. Ecol. Appl. 6, 1018±1024.
Frost, T.M., Carpenter, S.R., Ives, A.R., Kratz, T.K., 1995. Species
compensation and complementarity in ecosystem function. In:
Jones, C.G., Lawton, J.H. (Eds.), Linking Species and
Ecosystems. Chapman and Hall, New York, pp. 224±239.
Gerlach, J.P., Reich, P.B., Puettmann, K., Baker, T., 1997. Species,
diversity, and density affect tree seedling mortality from
Armillaria root rot. Can. J. For. Res. 27, 1509±1512.
Gleason, H.A., 1926. The individualistic concept of the plant
association. Bull. Torrey Bot. Club. 53, 7±26.
Haila, Y., Hanski, I.K., NiemaÈlaÈ, J., Punttila, P., Raivio, S., Tukia,
H., 1994. Forestry and boreal fauna: matching management
with natural forest dynamics. Ann. Zool. Fennici 31, 187±202.
Hannah, L., Carr, J.L., Lankerani, A., 1995. Human disturbance
and natural habitat: a biome level analysis of a global data set.
Biodiv. Conserv. 4, 128±155.
Hansson, L. (Ed.), 1997. Boreal ecosystems and landscapes:
structures, processes and conservation of biodiversity. Ecol.
Bull. 46.
Hewit, G.M., 1996. Some genetic consequences of ice ages and
their role in divergence and speciation. Biol. J. Linnean Soc.
58, 247±276.
Holling, C.S., 1992. Cross-scale morphology, geometry and
dynamics of ecosystems. Ecol. Monogr. 62, 447±502.
Holling, C.S., Schindler, D.W., Walker, B.W., Roughgarden, J.,
49
1995. Biodiversity in the functioning of ecosystems: an
ecological synthesis. In: Perrings, C., MaÈler, K.-G., Folke, C.,
Holling, C.S., Jansson, B.-O. (Eds.), Biodiversity Loss:
Economic and Ecological Issues. Cambridge University Press,
pp. 44±83.
Hooper, D.U., Vitousek, P.M., 1997. The effects of plant
composition and diversity on ecosystem processes. Science
277, 1302±1305.
Kareiva, P., Wennergren, U., 1995. Connecting landscape patterns
to ecosystem and population processes. Nature 373, 299±302.
KaÊreÂn, O., 1997. Effects of air pollution and forest regeneration
methods on the community structure of ectomycorrhizal fungi.
Ph.D. thesis. Acta Univ. Agricult. Sueciae, Silvestria, 33.
Laakso, J., SetaÈlaÈ, H., 1999. Primary productivity and functional
redundancy in below-ground food webs. Oikos, in press.
Leonardi, S., Menozzi, P., 1995. Genetic variability of Fagus
sylvatica L. in Italy: the role of postglacial recolonization.
Heredity 75, 35±44.
Loreau, M., 1995. Consumers as maximizers of matter and energy
flow in ecosystems. Am. Nat. 145, 22±42.
Martin, O., 1989. Click beetles (Coleoptera: Elateridae) from old
deciduous forests in Denmark. Entomol. Medd. 57, pp. 1±507
(in Danish with English summary).
May, R.M., 1973. Stability and complexity in model ecosystems.
Princeton University Press, Princeton, NJ.
McGrady-Steed, J., Harris, P.M., Morin, P.J., 1997. Biodiversity
regulates ecosystem predictability. Nature 390, 162±165.
McNaughton, S.J., 1993. Biodiversity and function of grazing
ecosystems. In: Schultze, E.-D., Mooney, H.A. (Eds.), Biodiversity and Ecosystem Function. Springer, pp. 361±383.
Naeem, S., Li, S.B., 1997. Biodiversity enhances ecosystem
reliability. Nature 390, 507±509.
Naeem, S., 1998. Species redundancy and ecosystem reliability.
Conserv. Biol. 12, 39±45.
NiemelaÈ, J., 1997. Invertebrates and boreal forest management.
Conserv. Biol. 11, 601±610.
Nilsson, S.G., Ericson, L., 1997. Conservation of plant and animal
populations in theory and practice. Ecol. Bull. 46, 117±139.
Nilsson, S.G., 1997. Forests in the temperate-boreal transition:
natural and man-made features. Ecol. Bull. 46, 61±71.
Nilsson, S.G., Baranowski, R., 1997. Habitat predictability and the
occurrence of wood beetles in old-growth beech forests.
Ecography 20, 491±498.
Odum, E.P., 1969. The strategy of ecosystem development. Science
164, 262±270.
Owen-Smith, N., 1987. Pleistocene extinctions: the pivotal role of
megaherbivores. Paleobiology 13, 351±362.
Perlin, J., 1988. A Forest Journey. Norton, NY.
Perrings, C., MaÈler, K.-G., Folke, C., Holling, C.S., Jansson, B.-O.
(Eds.), 1995. Biodiversity Loss: Economic and Ecological
Issues. Cambridge University Press.
Peterken, G.F., 1996. Natural Woodland. Cambridge University
Press.
Peterson, G., Allen, C.R., Holling, C.S., 1998. Ecological
resilience, biodiversity and scale. Ecosystems 1, 6±18.
Pickett, S.T.A., White, P.S. (Eds.), 1985. The Ecology of Natural
Disturbances and Patch Dynamics. Academic press, New York.
50
J. Bengtsson et al. / Forest Ecology and Management 132 (2000) 39±50
Pimm, S.L., 1991. The Balance of Nature? Chicago University
Press.
Polis, G.A., 1994. Food webs, trophic cascades and community
structure. Aust. J. Ecol. 19, 121±136.
Power, A.G., Kareiva, P., 1990. Herbivorous insects in agroecosystems. In: Carroll, C.R., Vandermeer, J.H., Rosset, P.M.
(Eds.), Agroecology. McGraw Hill, New York, pp. 301±327.
Rackham, O., 1998. Savanna in Europe. In: Kirby, K.J., Watkins,
C. (Eds.), The Ecological History of European Forests.
Cambridge University Press, Cambridge, UK, pp. 1±24.
Rose, F., 1976. Lichenological indicators of age and environmental
continuity in woodlands. In: Brown, D.H., Hawksworth, D.L.,
Bailey, R.H. (Eds.), Lichenology. Progress and Problems.
Academic Press, London, pp. 279±307.
Rusch, G.M., Oesterheld, M., 1997. Relationship between
productivity, and species and functional group diversity in
grazed and non-grazed Pampas grassland. Oikos 76, 519±526.
Siitonen, J., Martikainen, P., 1994. Occurrence of rare and
threatened insects living on decaying Populus tremula: a
comparison between Finnish and Russian Karelia. Scand. J.
For. Res. 9, 185±191.
Speight, M.C.D., 1989. Saproxylic invertebrates and their conservation. Nature and Environment Series, No. 42, Strasbourg.
Stanners, D., Bourdeau, P. (Eds.), 1995. Europe's Environments.
European Environment Agency, Copenhagen.
Su, Q., MacLean, D.A., Needham, T.D., 1996. The influence of
hardwood content on balsam fir defoliation by spruce budworm. Can. J. For. Res. 26, 1620±1628.
Tansley, A.G., 1935. The use and abuse of vegetational concepts
and terms. Ecology 16, 284±307.
Tilman, D., Downing, J.A., 1994. Biodiversity and stability in
grasslands. Nature 367, 363±365.
Tilman, D., Wedin, D., Knops, J., 1996. Productivity and
sustainability influenced by biodiversity in grassland ecosystems. Nature 379, 718±720.
Tilman, D., Knops, J., Wedin, D., Reich, P., Ritchie, M., Sieman,
E., 1997. The influence of functional diversity and composition
on ecosystem processes. Science 277, 1300±1302.
Wardle, D.A., Zackrisson, O., HoÈrnberg, G., Gallet, C., 1997. The
influence of island area on ecosystem properties. Science 277,
1296±1299.
White, P.S., 1979. Pattern, process and natural disturbance in
vegetation. Bot. Rev. 45, 229±299.
Zackrisson, O., 1977. Influence of forest fires on the North Swedish
boreal forest. Oikos 29, 22±32.
Ê gren, G.I., 1997. Soil food webs and
Zheng, D.W., Bengtsson, J., A
ecosystem processes: decomposition in donor-control and
Lotka±Volterra systems. Am. Nat. 149, 125±148.
Ê gren, G.I., Bengtsson, J., 1999. How do soil
Zheng, D.W., A
organisms affect total organic nitrogen storage in soils and
substrate nitrogen to carbon ratio? A Theoretical Analysis
Oikos 86, 430±442.