2010, VOL. 86, No3 — THE FORESTRY CHRONICLE
Negative or positive effects of plantation and intensive forestry
on biodiversity: A matter of scale and perspective
by Henrik Hartmann1, Gaëtan Daoust2, Brigitte Bigué3and Christian Messier4
Terrestrial biodiversity is closely linked to forest ecosystems but anthropogenic reductions in forest cover and changes in
forest structure and composition jeopardize their biodiversity. Several forest species are threatened because of reduced
habitat quality and fragmentation or even habitat loss as a result of forest management activities. In response to this threat,
integrated forest management (IFM) was developed in the early 1990s and has been applied over large spatial scales ever
since. While IFM seeks to satisfy both human resource demands and ecosystem integrity, the whole forest matrix is
affected and this may also have negative impacts on biodiversity. The concept of forest zoning (e.g., Triad) avoids these
issues by physically separating land uses from each other. The zoning approach has been developed in the same period as
IFM, but there are still very few examples of large-scale applications. This may be because its distinctiveness from IFM
may not always seem clear and because forest zoning is not easily implemented. Here we explain these differences and
show that IFM and the zoning approach are indeed different management paradigms. We advocate the use of high-yield
plantations within the zoning paradigm as a means for biodiversity conservation and review the literature (with an
emphasis on the northern hemisphere and on plantation forestry within a land-zoning approach) on impacts of forest
management activities on biodiversity. Furthermore, we give advice on issues that require consideration when implement-
ing forest zoning at both the stand and the landscape levels. We recommend several small changes in design and manage-
ment of forest plantations as a means to significantly increase their biodiversity value. We conclude that while forest zon-
ing seems an adequate strategy for the Canadian forestry sector, a shift in paradigm must carry over to policy-makers and
legislation if this approach is to succeed.
Key words: biodiversity, landbase zoning, forest management, intensive silviculture, plantation forests
La biodiversité terrestre est étroitement liée aux écosystèmes forestiers, mais la réduction du couvert forestier par des acti-
vités humaines et des changements dans la structure et la composition forestière constituent une menace pour la biodi-
versité. Plusieurs espèces forestières sont menacées par la dégradation, la fragmentation ou même la perte totale de leur
habitat occasionnées par l’aménagement forestier. Pour y remédier, l’aménagement intégré des ressources forestières
(AIRF) a été développé et appliqué à grande échelle depuis le début des années 1990. Quoique l’AIRF vise à la fois à satis-
faire la demande de produits et services forestiers et à protéger l’intégrité des écosystèmes forestiers, cette approche
engendre une pénétration presque totale de la matrice forestière et pose ainsi des risques pour la biodiversité. Par contre,
le zonage forestier (ex. Triade) vise à réduire la pénétration du territoire forestier par une séparation spatiale des objectifs
d’aménagement. Quoique développé dans la même période que l’AIRF, l’application du zonage à grande échelle est très
récente. Ceci pourrait être dû au fait que le zonage est souvent considéré comme un outil de l’AIRF et non comme une
stratégie différente d’aménagement. Aussi, toutes les étapes de la planification du zonage sont complexes et sa mise ne
œuvre est ainsi difficile. Dans cette revue de littérature, nous expliquons les différences entre l’AIRF et le zonage et nous
montrons que les deux approches sont en fait des paradigmes d’aménagement distincts. Nous promouvons l’utilisation de
plantations à haut rendement dans le cadre du zonage comme un moyen de conservation de la biodiversité. Nous
revoyons la littérature (en mettant l’emphase sur l’hémisphère nord et sur la foresterie de plantation dans l’approche de
zonage) sur l’impact des activités d’aménagement forestier sur la biodiversité. De plus, nous donnons des conseils sur la
mise en œuvre du zonage forestier à la fois à l’échelle du peuplement et du paysage ainsi que des recommandations pour
le design et la gestion de plantations afin d’augmenter leur valeur pour la biodiversité. Nous concluons que, bien que le
zonage forestier semble une stratégie adéquate pour le secteur forestier canadien, un changement de paradigme doit s’éta-
blir parmi les décideurs politiques et dans la législation pour que cette approche si prometteuse peut réussir.
Mot-clés: biodiversité, zonage forestier, aménagement forestier, sylviculture intensive, foresterie de plantations
1Max–Planck Institute for Biogeochemistry, Hans–Knöll Str. 10, 07745 Jena, Germany. Corresponding author. E–mail: hhart@bgc-jena.
2Canadian Forest Service, Laurentian Forestry Centre, 1055 Rue du P.E.P.S. Sainte–Foy (Québec) G1V 4C7. E–mail: Gaetan.Daoust@
3Réseau Ligniculture Québec, Pavillon Charles–Eugène–Marchand, Université Laval, Québec (Québec) G1K 7P4. E–mail: Brigitte.Bigue@
4Université du Québec à Montréal, succursale Centre–Ville, CP. 8888, Montréal (Québec) H3C 3P8. E–mail: email@example.com.
MAY/JUNE 2010, VOL. 86, NO. 3 — THE FORESTRY CHRONICLE 355
Introduction: Forest Management as a Threat to
Terrestrial biodiversity is closely linked to forest ecosystems
and forests are considered “essential to economic develop-
ment and the maintenance of all forms of life” (UNO 1992).
The world’s forests cover almost 30% of the land surface,
comprise more than 4 billion hectares and the majority of ter-
restrial species depend on or dwell in these forests (Gaston
2000, FAO 2007). However, human activities reduce the nat-
ural forest cover by 13 million hectares annually and despite
afforestation and reforestation efforts, the world’s forest cover
was reduced by 3% during the period 1990 to 2005 at an
annual rate of 0.2%; hence the world is losing 20000 hectares
of forest every day (FAO 2007).
In Canada the forest cover has been reduced by 77 million
hectares since European colonization began but the remaining
402 million hectares of forest and forested land include about
10% of the world’s forests and 30% of boreal forests and offer
habitats to about two-thirds of an estimated 140000 species in
Canada (Mosquin et al. 1995). However, this forest cover con-
tinues to shrink and to be modified due to human activities
including forestry. In 2007 alone, about 47700 hectares were
deforested in Canada (CFS 2009). The major cause of defor-
estation in Canada is agriculture and forestry partially coun-
teracts the loss of forest cover with annual afforestation efforts
(9400 hectares in 2005, CFS 2009). However, the resulting net
loss of forest cover may have an important impact on biodiver-
sity. In Quebec alone, more than 2000 plant, 95 fish, 19
amphibian, 15 reptile, 223 bird and 63 mammal species live in
approximately 50 million hectares of forest (MRNFQ 2008a).
About 0.92% of Quebec’s commercial forest is harvested annu-
ally and more than 50 woodland species have been declared
endangered or vulnerable (MRNFQ 2008a).
In the southern parts of Quebec, habitat destruction or
degradation from urbanization and agricultural activities are
responsible for this threat to biodiversity but in more remote
and forested regions, impacts from forest harvest (e.g., habitat
fragmentation, changes in stand age, density and composi-
tion, snag removal) have been identified as causal links to
threats to biodiversity (MRNQ 1996).
In past decades an integrated forest management approach
(IFM)—the management of the ensemble of all desirable for-
est resources at the stand scale—has been advocated in
response to these threats. More recently, forest zoning at the
regional level has been proposed as yet another potential
solution to human threats to forest biodiversity. However,
there still is much confusion about what the forest zoning
approach actually is, how it differs from IFM and what its
This literature review aims at elucidating these points. We
critically define the concept of biodiversity and show the dif-
ferences between IFM and the zoning approach. We then
review the literature on impacts of forest management on bio-
diversity at the stand and the landscape levels. Finally, we pro-
vide conceptual advice on implementing intensive and plan-
tation forestry within a zoning approach as a means to
conserve biodiversity within specific areas, to maintain it on a
larger scale or even restore biodiversity where depleted by for-
mer land uses.
What Exactly Is Biodiversity?
Biodiversity is “the variation of life at all levels of biological
organization” (Gaston and Spicer 2004) and, as such, a diffuse
concept that can be defined for organisms (e.g., individuals),
groups of individuals (e.g., species) or groups of species (e.g.,
functional groups) or even for biophysical components of the
environment (e.g., habitats) and their characteristics (e.g.,
complexity of a forest canopy). Furthermore, biodiversity can
be defined at different hierarchical (e.g., organismic, spatial)
scales and for diverse attributes such as composition, struc-
ture and function. Hence, there is no unique and equivocal
definition of what biodiversity actually means.
Measures of biodiversity further complexify the issue. Most
intuitively, biodiversity is often measured as richness, i.e., the
number of different expressions of an observable attribute (e.g.,
alleles, species identity) over the extent of interest (e.g., loci,
area). However, the number of individuals may not be uni-
formly distributed across species, and measures of abundance
can be used to quantify this distribution. Indices of biodiversity,
like the Shannon Index, include both richness and abundance
and are more informative (Gaston and Spicer 2004).
However, what one wants to measure depends on specific
objectives and constraints. In forest ecology, we often quantify
the richness, abundance or diversity indices of species or
functional groups (e.g., shrubs, herbs, trees). While there are
good reasons for the use of particular species because of their
ecological significance in key functions of ecosystems (i.e.,
species that offer habitat to other species or that are determi-
nants of the trophic chain), the choice of these species is often
questionable, comprises a host of pitfalls and may not be use-
ful in all circumstances (Bengtsson 1998; Simberloff 1998,
1999; Bifolchi and Lodé 2005). Hence, all of these measures
Henrik Hartmann Gaëtan Daoust Brigitte Bigué Christian Messier
2010, VOL. 86, No3 — THE FORESTRY CHRONICLE
will only give a partial picture of the biodiversity and this even
more so because for some species groups (e.g., soil animals
like mites) applying the species concept may be problematic
altogether (Bengtsson 1998).
Integrated Forest Management and Forest Zoning –
What’s the Difference?
Integrated forest management has been developed and
applied on large scales since about the early 1990s (Behan
1990, Born and Sonzogni 1995). It stipulates that the “protec-
tion of diversity must be incorporated into everything we do
every day on every acre, whether preserve or commodity
land” (Franklin 1989: 8). Integrated forest management seeks
to provide a sustained flow of timber and non-timber forest
products by setting up a framework of management priority
areas. Protected areas with no industrial resource extraction
are to provide essential habitats and wildlife refuge areas. The
managed landbase can be subdivided into zones depending
on management objectives, i.e., commercial use with conser-
vation of unique features (e.g., habitats), commercial use for
timber production and other forest resources and enhanced
management zones subjected to the protection of environ-
mental quality. However, integrated forest management
affects the whole forest matrix (even though some chunks of
forest may be excluded from actual harvest) and therefore
environmental impacts from forest operations will also affect
the whole forest matrix. For example, it has been recognized
that forest management contributes indirectly to environ-
mental damage through its road network via erosion, siltation
of watercourses and increases in hunting pressure on wildlife
populations (Forman and Alexander 1998).
The forest zoning approach, on the other hand, seeks to
avoid disturbances of the whole landbase. The concept of zon-
ing has a long history in urban development (e.g., separating
industrial and residential developments) and has also been
employed in natural settings for the protection of habitats
since the late 1800s (Grove 1992). The forest zoning approach
builds on designating permitted uses of land, not only land-
use priorities as within IFM, which are based on mapped
zones that separate one land use from another. The concept
was proposed in the early 1990s as a three-zone approach in
Maine, USA (Seymour and Hunter 1992) but it is only recently
that examples of large-scale application of this concept have
emerged (e.g., Messier et al. 2009). Although the zoning
approach seems at first glance quite similar to IFM, it is a con-
ceptually different paradigm. Because some of the manage-
ment objectives (e.g., biodiversity conservation) can only be
defined on broader scales that span across priority zones, com-
mercial timber extraction in IFM is often subject to harvest
constraints (e.g., minimum harvesting age, harvesting rates,
size, spatial and temporal distribution of cutblocks) and there-
fore residualized (i.e., effectively being treated as a non-priority
objective) even within enhanced management zones (Saha-
jananthan et al. 1998). The zoning approach avoids residualiz-
ing timber extraction (or any other land use) because land use
objectives are geographically separated.
Economic analysis has shown that the zoning approach is,
at the broad level, a more viable strategy for sustainable multi-
ple use of forests than the stand-scale approach of integrated
forest management (Vincent and Binkley 1993). By geograph-
ically separating competing management objectives, land zon-
ing allows for more intensive production and an increase in
cumulative returns for specialized practices on their respective
landbases instead of mediocre returns of generalized practices
across the entire landbase (Boyland et al. 2004). A comparison
of several simulated management strategies, including inte-
grated management and several zoning scenarios, showed that
forest zoning made it possible to produce more timber and
maintain more old growth in the landscape than the other
strategies examined (Coté et al. 2009).
Forest zoning directly contributes to the conservation of
biodiversity by reducing the overall length of the road net-
work associated with forest operations (Sahajananthan et al.
1998) and is thought to relieve harvesting pressure on natural
forests through increased fibre production on an intensively
managed commodity landbase of limited geographical extent.
As a consequence, extensive proportions of natural forests
could be taken out of the forest management scheme and set
aside for conservation purposes (Sedjo and Botkin 1997). The
remainder of forest stands would be managed under a less
intensive management approach by emulating natural distur-
bance regimes (Hunter 1999, Bergeron et al. 2002) and imple-
menting new approaches based on complexity and resilience
(Chapin et al 2009, Puettmann et al. 2009). In this zone, the
road network may be made less permanent or decommis-
sioned after harvest, since the objective is to let the forest reor-
ganize itself following harvest, further reducing the negative
impact of permanent roads. Expected gains in productivity
from the use of high-yield plantations will make it possible to
maintain an acceptable level of environmental constraints,
such as riparian or roadside vegetation buffers, not only
throughout the entire landbase but also within the intensive
management zone (Krcmar et al. 2003). Furthermore, inten-
sive (plantation) forestry will be an important element in sus-
taining international economic competitiveness (Park and
Wilson 2007) and may be unavoidable when biodiversity
conservation is at stake (Wagner et al. 2004).
Are Plantations Biological Deserts?
Forest plantations, especially exotic single-species plantations,
are thought to offer a less favourable habitat than natural
forests (Hunter 1999, Hartley 2002) and have a reputation for
being “biological deserts” (Allen et al. 1995, Dyck 1997).
However, objectively evaluating the value of plantations for
conservation purposes is not trivial and requires assessing
their contribution to biodiversity at higher spatial scales
(Brockerhoff et al. 2008). Hence, one should consider (i)
whether plantations reduce harvest pressure on natural
forests, (ii) what kind of land use or vegetation they replace,
(iii) what other potential alternative land uses there are to be
compared with, (iv) whether local species had enough time to
colonize and adapt to the new habitat and (v) are plantations
managed for production purposes only or with conservation
goals in mind (Brockerhoff et al. 2008).
However, these issues are rarely considered when planta-
tions are assessed for their biodiversity value and therefore
studies from around the world show that single-species plan-
tations are, at the stand scale, often less diverse than natural or
semi-natural forests with respect to birds (e.g., Baguette et al.
1994, Gjerde and Sætersdal 1997, Twedt et al. 1999), arthro-
pods (e.g., Samways et al. 1996, Fahy and Gormally 1998,
Anderson and Death 2000, Davis et al. 2000, Magura et al.
2000) or plants (e.g., Fahy and Gormally 1998, Humphrey et
al. 2002, Aubin et al 2008). Nevertheless, it has also been
MAY/JUNE 2010, VOL. 86, NO. 3 — THE FORESTRY CHRONICLE 357
shown that forest plantations can contribute to restoring
some of the floristic diversity on abandoned agricultural land
(Newmaster et al. 2006, Aubin et al. 2008) and some planta-
tions may have a surprisingly diverse understory (e.g., Allen
et al. 1995, Keenan et al. 1997, Oberhauser 1997). Further-
more, some plantations can have communities as diverse as
natural or secondary forests of birds (e.g., Clout and Gaze
1984, Brockie 1992, Kwok and Corlett 2000), fungi or inver-
tebrate species (e.g., Humphrey et al. 1999, 2000, 2002;
Ohsawa 2004). Moreover, plantations can have, in the absence
of management strategies aiming at eliminating naturally
occurring woody understory species, a “catalytic” effect by
facilitating the colonization of early and even late successional
tree species and other floristic elements from the surrounding
forest (Brockerhoff et al. 2008).
The influence of plantation type (conifer vs. broadleaf) on
biodiversity is ambiguous and results from mechanisms
driven by a species’ ecological footprint on light, water and
nutrient availability, as well as by physical effects and the pres-
ence of phytotoxic compounds in the litter (Aubin et al. 2008,
Barbier et al. 2008). In a conifer-dominated landscape, under-
story plant diversity was greater in pure broadleaf stands than
in mixed-species or pure coniferous stands (Oaten and
Larsen 2008). However, differences in woody species richness
between broad-leaved and conifer stands was dependent on
patch size and the greater diversity observed in small patches
of broad-leaved forests decreased with patch size and may be
due to different management regimes (Estevan et al. 2007).
Furthermore, some conifer plantations may have greater
plant species richness than broad-leaved secondary forests
Keeping in mind that biodiversity in plantations is depend-
ent on the local natural settings (e.g., proximity, type, and age
of neighbouring natural forest stands, see Smith et al. 2005), it
has been observed that pure stands (i.e., deciduous or conifer-
ous) could support, in some cases, a richer understory vegeta-
tion than mixed-species stands and that species richness was
generally greater in deciduous stands than in coniferous
stands (Barbier et al. 2008). Similarly, plant and carabid beetle
richness was greater in semi-natural deciduous forests than in
conifer plantations (Fahy and Gormally 1998). Understory
functional groups and environmental conditions of deciduous
plantations converged toward those of old naturally regener-
ated forests although species diversity remained low and
understory structure developed poorly compared with
unplanted stands (Aubin et al. 2008). Diversity also changed
throughout the development cycle of forest stands. In general,
species diversity was greater in the early stages and declined
during later stages of stand development and early stages host
mainly generalist species whereas later mature stands favoured
specialist species (Smith et al. 2005). During the stand devel-
opment cycle, understory vegetation composition changed
from annual to woody species (Eycott et al. 2006) and in
conifer stands most vascular plant species disappeared rapidly
during stand development; however, bryophyte species could
persist in these stands (Smith et al. 2005).
The negative impact on wildlife species of replacing natu-
ral forests with exotic plantations has been well documented
(e.g., Clout and Gaze 1984, Estades and Temple 1999, Linden-
mayer et al. 1999, Magura et al. 2000) although biodiversity
tends to increase with plantation age (Smith et al. 2005). On
the other hand, even exotic single-species plantations can
contribute to restoring animal diversity on degraded land
(Lugo 1997) and, in heavily modified landscapes, may offer
the only available habitat for local species (Pawson et al. 2008)
or even for endangered or threatened species (Barbaro et al.
2008). Moreover, exotic single-species plantations may some-
times represent the very last remaining habitat for locally
endemic species (Berndt et al. 2008).
When plantations replace other productive land uses, such
as agriculture, they are usually considered the “lesser evil”
even among environmentalists (Brockerhoff et al. 2008). In
support of this statement, Stephens and Wagner (2007)
showed that when prior land use was considered in the com-
parisons, plantations may be of greater diversity than the land
use they replace. The authors conclude that “increasing biodi-
versity is a desired outcome that can be achieved through
appropriate forest management that includes the use of plan-
tations” (Stephens and Wagner 2007: 312). In light of this
information, it can be assumed that the use of sensitive and
appropriate forest management practices during planning,
execution and maintenance of forest plantation can avoid the
creation of biological deserts and may even contribute to the
conservation, promotion and restoration of biodiversity.
Impact of Intensive Forest Management on Biodiversity
Forest management practices have the potential to affect bio-
diversity at many spatial scales (Wigley and Roberts 1997). At
the stand level, silvicultural treatments, such as site prepara-
tion, planting, mechanical or chemical vegetation manage-
ment and thinning, alter stand density and age, and tree
species composition as well as soil properties and understory
vegetation. The changes in stand properties may then affect
plant and animal habitat quality and biodiversity.
In general, silvicultural treatments associated with intensive
forest management cause a temporal increase in species diver-
sity and shifts in the relative abundance and species composi-
tion of overstory tree, understory herb, and shrub communities
(Rowland et al. 2005). Although some of the impacts from
intensive management practices, such as stand type conversion,
stand structural changes and especially age-class truncation,
may decrease habitat suitability for several vertebrate wildlife
species that prefer old-growth forests, these same changes may
also increase habitat availability for species favouring young to
mature coniferous-dominated forests (Thompson et al. 2003).
Site preparation practices such as scariﬁcation, ploughing,
tilling, crushing, and slash burning may have negative
impacts on species such as salamanders and invertebrates
associated with or entirely restricted to the forest floor
(deMaynadier and Hunter 1995). The greatest impact of these
treatments is on the availability of dead organic matter—cav-
ity trees, snags, coarse woody debris, and the organic horizon
of the forest floor (Freedman et al. 1993, 1996)—and mini-
mizing intensive site preparation will help maintain coarse
woody debris and thus biodiversity (Carey and Johnson
1995). Plant species diversity is not at all or somewhat posi-
tively affected by ploughing or tilling (Haeussler et al. 2002,
2004), but moderate treatments may be necessary to prevent
site colonization by ruderal plants (Soo et al. 2009) and other
early colonizers, many of which could be invaders.
Although exotic tree species are often preferred for refor-
estation or afforestation because of their superior yield and
tolerance to native forest pests, indigenous plant and animal
2010, VOL. 86, No3 — THE FORESTRY CHRONICLE
species are usually better adapted to native tree species and
can more readily access their resources (Harrison et al. 2000).
However, even exotic tree species may have beneficial effects
on biodiversity if planted on degraded sites or on specific soil
types (Chiarucci 1996, Lugo 1997). Furthermore, several
studies have shown that many local plant and animal species
can make use of the habitats created by exotic tree species
(Barlow et al. 2007).
Chemical release treatments were applied in Canada on a
large scale in the 1970s and 1980s but increased public resist-
ance has amounted to a complete ban on the use of herbicides
in Quebec’s forests (Fortier and Messier 2006). However,
chemical release treatments cause only—if any—short-term
effects on wildlife biodiversity (Lautenschlager 1993, Thomp-
son et al. 2003) which may in some cases even be positive
(Newton et al. 1989). Similarly, herbicide applications usually
do not cause changes in understory vegetation diversity and
reported changes are ephemeral (Rowland et al. 2005).
Mechanical treatments are even less prone to cause changes to
biodiversity (Fredericksen et al. 1991) because they do not
reinitiate the successional pathway of the ground vegetation,
unlike chemical treatments.
Fertilizer treatments may temporarily increase plant
species richness (Fehlen and Picard 1994), change vegetation
composition (Nohrstedt 1994) or the abundance of specific
plant forms (Olsson and Kellner 2006) but usually have no
lasting impact on plant diversity or richness (Demchik and
Sharpe 2001) or on the diversity of plant functional groups
(Bauhus et al. 2001).
Thinning increases light penetration in the forest stand
and thereby favours pioneer species (Moore and Allen 1999).
Increases in light and nutrient availability from thinning facil-
itates the development of dense herbaceous understory vege-
tation (Griffis et al. 2001) and improved vertical structure fol-
lowing thinning increases bird diversity and habitat use
(Bisson and Stutchbury 2000, King and DeGraaf 2000). How-
ever, potential changes in forest floor habitat condition from
thinning did not negatively influence the use of sensitive indi-
cator species (salamander) (Messere and Ducey 1998) and
thinning may increase the abundance of particular wildlife
species (Sullivan and Klenner 2000).
Clearcutting sets forest stands back to an earlier succes-
sional stage (Hansen et al. 1991, Christensen and Emborg
1996) and although there are some similarities between
clearcuts and forest fires, clearcutting can hardly fulfill all the
ecological functions of this type of natural disturbance
(McRae et al. 2001, Bergeron et al. 2002). Furthermore, the
use of heavy machinery can cause soil compaction, decrease
water infiltration and soil aeration and destroy soil structure
(McClurkin and Duffy 1973, Xu et al. 1999) and soil com-
paction may cause decreases in plant diversity, specifically of
grasses and other herbaceous plants in some instances
(Mellin 1995). However, by increasing radiation levels on the
forest floor, clearcutting also increases biomass production
and diversity of the plant community (Perison et al. 1997,
Yorks and Dabydeen 1999). Progressive removal of the
canopy layer (shelterwood cut) may enhance the diversity of
the plant community compared with complete, patch, or no
canopy removal (Beese and Bryant 1999). Also, there may be
a threshold of cutblock size beyond which clearcutting
impacts on biodiversity increase disproportionately to their
size (Pawson et al. 2008). Clearcutting impacts on wildlife are
ambiguous and as with any other form of management deci-
sion—including taking no action at all—clearcuts favour
some species while disadvantaging others (Bunnell et al.
1999, Bayne and Hobson 2000). Furthermore, impacts on
wildlife depend on the spatial extent of the silvicultural action
in relation to the living space of the studied species (Potvin et
al. 1999) but in general the loss of vertical structure and
removal of snags and coarse woody debris are important fac-
tors related to habitat loss of most vertebrate and invertebrate
species (Humphrey et al. 2003).
The landscape is a mosaic of different biophysical elements
(e.g., mountain ranges, forests, agricultural fields, human
infrastructures, streams and lakes) and the spatial distribu-
tion of critical resources within these elements determines the
availability of habitats and the distribution of species within
the landscape (Wiens et al. 1987, Debinski et al. 2001). Biodi-
versity itself is a dynamic concept that is defined at different
spatial and temporal scales. Within a landscape, successional
pathways alter the composition of animal and plant commu-
nities (Kimmins 1997, Bunnell et al. 1999) as well as their
structure and functional role (Westworth and Telfer 1993).
Hence, the role of a species or a functional group (e.g., pio-
neer species) within the processes of an ecosystem may be
important in an early successional phase but declines in later
phases (McLaren 1996, Kimmins 1999). The assessment and
management of biodiversity must take this dynamic into
account (Hansen et al. 1995, Gaines et al. 1999). However,
most studies on the impact of forest management on biodi-
versity are limited to small spatial (i.e., forest stands or plan-
tations) and temporal scales (often not more than several
years) and therefore do not include this temporal dynamic
(Bunnell and Huggard 1999, Crawley and Harral 2001).
While we recognize the difficulty of obtaining data on larger
spatial and temporal scales, we must stress the fact that short-
term data may possibly have contributed to the overall criti-
cism of plantation and intensive forestry. As has been shown
by Smith et al. (2005) and Aubin et al. (2008), there was a sub-
stantial gain in the biodiversity value during the development
cycle of plantations. Hence, long-term data may reveal that
plantations are generally of greater value as habitats than
While studies over longer time windows may help
improve the assessment of long-term impacts of intensive for-
est management on biodiversity (Rowland et al. 2005), spatial
constraints may be harder to surmount. Even within a single
stand, the assessment of all taxa is extremely difficult and
costly so it seems even more impossible to accurately measure
the biodiversity of a whole watershed or even a landscape
(Humphrey et al. 2003). Hence, any assessment of the biodi-
versity of either a natural forest or a plantation will necessar-
ily always be incomplete.
At the stand scale, many silviculture interventions result in
some increases in biodiversity compared to what was present
in the same stand before, and its negative public reputation
(Wagner et al. 1998) may not always be justified. However, the
key issue in conserving and promoting biodiversity with for-
est management lies within the coordination of silvicultural
activities at the landscape scale. In the past, forest managers
have defined their silvicultural approach at the stand scale
and thereby neglected the spatial and temporal variability of
MAY/JUNE 2010, VOL. 86, NO. 3 — THE FORESTRY CHRONICLE 359
forested landscapes (Kimmins 1995). It is the repetition of
similar and simplified stand-scale management practices over
a larger landscape that has created the problem with tradi-
tional forestry practices as discussed in a recent book by
Puettmann et al. (2009). In the boreal forest of Scandinavia,
the large-scale transformation of forest habitats into homoge-
nous and production-intensive systems, with an extensive use
of exotic tree species plantations, has posed a threat to biodi-
versity and has led to major public concerns about forest
management practices (Larsson and Danell 2001). These prob-
lems can be avoided when objectives are defined over larger
spatial scales as is the case with the forest zoning approach.
At the landscape scale
Forest zoning seems a particularly well-adapted strategy for
the Canadian forestry sector. More than 90% of Canadian
forests are publicly owned and therefore fulfill a major
requirement for successful application of a zoning approach:
ownership control over entire landscapes, cooperation
between agencies and flexibility in assigning land allocations
to any of the zones (Sarr and Puettmann 2008). Proportions
of landbase required for each zone are context-specific and
depend on overall management objectives. However, esti-
mates for intensive silviculture range from 10% (Sedjo and
Botkin 1997) to 30% (Sahajananthan et al. 1998). Messier et
al. (2003a) envision for Canada proportions of 14% for inten-
sive silviculture (including 4% of fibre farms), 74% for ecosys-
tem management and 12% for conservation, whereas a case
study covering 1.2 million hectares of forest land in coastal
British Columbia allocated proportions of 65% for timber
production, 25% for habitat (e.g., long rotations or modified
partial harvesting to retain habitat structures such as large
snags, large downed wood) and 10% for old-growth (biodi-
versity values) (Boyland et al. 2004). The first large-scale
application of the zoning approach was implemented in cen-
tral Quebec and divides the landscape into conservation areas
(11%), intensive forestry areas that includes fast-growing
plantations (20%) and an extensively managed areas (69%)
(Messier et al. 2009).
The challenge for forest managers will be to merge multiple
and competing objectives, such as timber production, recre-
ational values, habitat quality and biodiversity, over the land-
scape (Lindenmayer et al. 1999) while protecting the diversity
of the whole suite of ecosystems within conservation zones
covering only 10% to 14% of the landbase. This requires iden-
tifying critical structures and functions, species, communities
and ecosystems of the landscape and the conservation of the
most important (Lautenschlager 1995), although the criteria
for identifying these species, ecosystems and attributes are not
well defined (Sample 2003). However, the inverse approach of
allocating the richest sites to intensive silviculture first and
then establishing conservation sites may exacerbate any effect,
whether positive or negative, of intensive silviculture on biodi-
versity because of the positive relationship between site pro-
ductivity and animal and plant diversity (Thompson et al.
2003). A more appropriate approach would establish a core of
conservation sites that are representative of the ecosystem
diversity within the landscape and are large enough to sustain
landscape-scale processes first before deciding where to put
the intensive forestry areas (as has been done in the TRIAD
project of central Quebec: Messier et al. 2009). Also, areas cru-
cial for the visual quality of the landscape, exceptional wildlife
habitat, water resources and other desired values must be iden-
tified (Sahajananthan et al. 1998). Critical ecosystems such as
riparian zones should be protected (Sedjo and Botkin 1997)
and conservation zones should comprise high-value conserva-
tion forests (HVCF) (Jennings and Jarvie 2003). Within the
zone of active forest management, age-class truncating may
threaten species adapted to old-growth habitat (Thompson et
al. 2003) and a diverse configuration of silvicultural units at
the landscape level should be maintained to promote biodiver-
sity (Sargent 1992, Wigley and Roberts 1997, Smith et al. 2005,
Eycott et al. 2006).
Considering the potential impact of management prac-
tices on the composition and structure of forests, the impact
of forest management on biodiversity at the landscape level
may be assessed using the spatial distribution and proportion
of forest stands under different management regimes
(Hansen et al. 1991, Freedman et al. 1996) because the result-
ing stand characteristics can serve as indicators of habitat
quality at the landscape level (Watts et al. 2005). For example,
changes in forest edge length, core habitat proportion or iso-
lation of forest patches resulting from silvicultural treatments
may cause habitat fragmentation for several species (e.g., Rip-
ple et al. 1991, Mladenoff et al. 1993, Spies et al. 1994, Jules
1998). On the other hand, complex stand structure, the pres-
ence of big and old trees, snags and coarse woody debris gen-
erally increase the habitat quality of a stand (Humphrey et al.
2003, Humphrey 2005). The temporal variability of such indi-
cators can be modeled with long-term simulations of vegeta-
tion succession (Perry and Millington 2008) and simulation
models can give estimates of habitat availability and quality
over entire landscapes (Messier et al. 2003b, Marcot 2006).
Simulation models can also model direct effects of different
forest management strategies on vegetation characteristics
and habitat quality (Klenner et al. 2000, Shifley et al. 2008)
and their sustainability can be evaluated with dynamic-land-
scape metapopulation models (Wintle et al. 2005). Further-
more, prioritization tools can provide estimates of threats to
biodiversity resulting from alternate management strategies
and therefore can contribute to increasing transparency,
accountability and efficiency of public investment in biodi-
versity (Wintle 2008). Algorithms using biodiversity indica-
tors have been developed to facilitate landbase allocation
(Boyland et al. 2004) and conservation site selection (Walker
and Faith 1998). Such tools can be linked to optimization pro-
cedures (Haight et al. 2000) and assist in finding trade-off
solutions between different sets of forest values (e.g., biodiver-
sity vs. fibre production; Faith et al. 1996) and in analyzing the
feasibility of alternate zoning scenarios (Krcmar et al. 2003).
Within the zone of intensive forest management, the use of
herbicides could be permitted because increases in fibre pro-
duction are necessary for allocating a large proportion of pro-
duction forests to conservation and ecosystem management
(Fortier and Messier 2006). With an expected increase in the
global demand for wood, the use of herbicides may actually
become a crucial part of biodiversity conservation (Wagner et
al. 2004) although this is only true if forest conservation
objectives in exchange for increased fibre production are
guaranteed by rigorous laws (Paquette and Messier 2009).
Governmental commitments towards forest zoning (e.g.,
Government of Alberta 1997, MRNFQ 2008b) are a first and
important step in initiating the zoning approach but more
2010, VOL. 86, No3 — THE FORESTRY CHRONICLE
concrete measures for critically and objectively assessing and
verifying management achievements as well as promoting
cooperation between different agencies will be necessary to
ensure a successful application. The implementation of a zon-
ing approach within a 0.86 million hectare public forest man-
agement unit in central Quebec has been proven economi-
cally viable and socially acceptable and forest zoning could be
a good strategy for general application in the Canadian forest
management sector (Messier et al. 2009).
At the stand scale
Within stands at the landscape level and in areas where forest
zoning does not apply—mostly the southern belt of Canada
where urban development, agriculture and private forests
dominate—a stand-scale approach must be used for conserv-
ing and promoting biodiversity. Because plantations can
hardly fulfill all the ecological functions of the natural forest
(Christian et al. 1998), replacement of natural forests, where
they still exist, with plantations should be avoided. Similarly,
conversion of secondary or semi-natural forests to plantations
should also be avoided (Dyck 1997, Brockerhoff et al. 2008)
although this may be justified in specific circumstances, i.e.,
partial conversion of selected sites to high-yield plantations
within large forest tracts at close proximity to processing
industries. However, afforestation of abandoned agricultural
fields could contribute to maintaining or even increasing the
biodiversity value of the landscape by adding forest habitat to
the agricultural matrix and by mitigating edge effects and
increasing connectivity between patches of remnant natural
forest (Brockerhoff et al. 2008).
Minor adjustments in design and management are often
enough to improve the biodiversity value of plantations at the
stand level without compromising their productivity (Hartley
2002). However, the mechanisms that cause various species to
accept an environment as their habitat are complex and pre-
dicting the impact of changes in plantation management on
biodiversity is difficult. However, plantation design and man-
agement can add structural components and complexity to
plantation forests which, in turn, increase their potential to
host a larger community of plant and animal species (Allen et
al. 1996, Moore and Allen 1999). To do so, foresters can apply
several of the following measures (e.g., Kimmins 1995, Chris-
tensen and Emborg 1996, Bunnell et al. 1999, Thompson et al.
2003): (i) maintain snags and old trees during partial or
clearcuts; (ii) create or maintain coarse woody debris by leav-
ing large bole sections during harvesting or by girdling stand-
ing trees; (iii) maintain patches of natural forest in planta-
tions; (iv) create edges of natural, nonlinear shape and
maintain a balanced edge–area ratio in plantations; (v) pre-
serve wetland patches and open patches within plantations;
and (vi) vary spacing between rows of planted trees. Similarly,
Norton (1998) identifies four types of management action
that may improve the biodiversity value of plantations: (i)
retain indigenous plant and animal communities; (ii) estab-
lish a greater diversity of planted species; (iii) plant a diversity
of tree species along streams and roads; and (iv) modify silvi-
cultural practices within plantations by increasing, for exam-
ple, rotation length between thinnings or stand age at harvest.
Increasing ecosystem diversity at the landscape level can
be achieved by planting on a rotational basis, i.e., a spatial and
temporal mix of species and age classes (Norton 1998). Simi-
larly, early and heavy thinning of some plantations combined
with late or no thinning of other plantations will create a
mosaic of relatively open areas and dense thickets (Hartley
2002). Indigenous tree species often have characteristics quite
similar to exotic species and should be favoured during plan-
tation establishment (Sedjo and Botkin 1997) because several
animal species may depend entirely on indigenous species
(Hartley 2002). Mixed-species plantations are often more
productive than monocultures (Kelty 1992, Potvin and
Gotelli 2008) and more resistant to pests (Stiell and Berry
1985) and have ecological, operational and economic benefits
even if one species comprises 90% of a site (Hartley 2002).
Further benefits anchored within the functional significance
of tree species diversity (e.g., facilitation, niche differentiation,
redundancy) are being investigated on long-term research
sites (Scherer-Lorenzen et al. 2007). Either way, tree improve-
ment programs should maximize the genetic diversity of the
selected tree species in order to preserve the species’ capacity
to adapt to changing environmental conditions and to natural
disturbances (Lambeth and McCullough 1997, Rajora 1999).
During site preparation, severe treatments favouring nutrient
leaching and erosion should be avoided (Hartley 2002) as
should the scarification, piling and burning of slash and
coarse woody debris (Carey and Johnson 1995). Retention of
native vegetation within plantations creates “life boats” for
animal species and facilitates the recolonization of plant
species and can be achieved by conserving patches of native
vegetation during harvest operations as well as by leaving
untreated “skips” during herbicide applications (Hartley
Some Last Words
This literature review has shown that forest management dis-
turbs, through its impact on the abiotic and biotic compo-
nents of forest ecosystems, the established equilibrium
between native species and thereby modifies biodiversity.
Even though there is an ever-increasing body of literature on
biodiversity, there still is a need for further studies on non-
vascular plant, fungi, and soil organism responses as well as
for long-term studies that widen our understanding of the
long-term effects of forest management on biodiversity
(Humphrey et al. 2003, Rowland et al. 2005). However, as
Humphrey et al. (2003: 111) stated “The need for rapid dis-
semination of data, interpreted in the form of readily usable
management advice, means that researchers are unlikely to be
given such opportunities for long-term studies. Instead, there
will be an increasing requirement for interim advice, based on
available data coupled with ecological wisdom.” This litera-
ture review aims at providing the former. However, because of
the complexity of biodiversity in terms of its attributes, scales
and measures, we cannot go beyond general advice. Specific
advice requires the definition of biodiversity objectives at a
local and regional scale. Here, forest managers can make a
substantial contribution by providing the necessary ecologi-
We thank three anonymous reviewers for the constructive
comments that helped to improve the paper. Thanks to the
Réseau Ligniculture Québec for providing a grant to H. Hart-
mann to write this review.
MAY/JUNE 2010, VOL. 86, NO. 3 — THE FORESTRY CHRONICLE 361
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