How climate change adaptation and mitigation strategies can threaten or enhance the biodiversity of production forests: Insights from Sweden

Full text

How climatechange adaptation and mitigation strategies can threaten or
enhance the biodiversity of production forests: Insights from Sweden
A. Felton
a,
⁎, L. Gustafsson
b
, J.-M. Roberge
c
,T.Ranius
b
,J.Hjältén
c
,J.Rudolphi
c
,M.Lindbladh
a
,J.Weslien
d
,L.Rist
e
,
J. Brunet
a
,A.M.Felton
a
a
Southern Swedish Forest Research Centre, Swedish University of Agricultural Sciences —SLU, Box 49, 230 53 Alnarp, Sweden
b
Department of Ecology, Swedish University of Agricultural Sciences —SLU, Box 7044, 750 07, Uppsala, Sweden
c
Department of Wildlife, Fish, and Environmental Studies, Swedish University of Agricultural Sciences —SLU, 901 83 Umeå, Sweden
d
Skogforsk, Uppsala Science Park, 751 83 Uppsala, Sweden
e
Department of Forest Ecology and Management, Swedish University of Agricultural Sciences —SLU, Skogsmarksgränd, 901 83 Umeå, Sweden
abstractarticle info
Article history:
Received 27 June 2015
Received in revised form 18 November 2015
Accepted 26 November 2015
Available online xxxx
Keywords:
Biological conservation
Ecosystem services
Global warming
Picea abies
Planted forest
Sustainable forest management
Anthropogenic climate changeis altering the management of production forests. These changesare motivated by
the need to adapt to the uncertainties and risks of climate change, and by the need to enlist their carbon storage
and sequestration capacity as part ofglobal mitigation efforts. These changes do however raise concerns regard-
ing the potential implications for forest biodiversity. Here we evaluate these concerns by assessing the biodiver-
sity implications of climate change adaptation and mitigation strategies (CCAMS) being implemented in the
production forestsof Sweden. We do so by identifying biodiversity goals aimed specifically at closingthe existing
gap between the habitat requirements of forest-dependent species, and the conditions provided by production
forests, in terms of tree species composition, forest structures, and spatio-temporal forest patterns. We then
use the existingliterature to determinewhether and by which pathway each CCAMS is likely to bridge or extend
this gap. Our results indicate that CCAMS will often come into direct or partial conflict with Swedish biodiversity
goals in production forests. Furthermore, some CCAMS which are inconsistent with biodiversity goals, such as
logging residue removal, are being implemented more extensively than those which were most consistent
with biodiversity goals. We nevertheless challenge the necessity of setting the preservation of forest biodiversity
against climate changemitigation and adaptation. We clarify how CCAMS withnegative biodiversity implications
may still be implemented without adverse outcomes, if coupled with conservation interventions, or combined
with other CCAMS deemed complementary in habitat provision.
© 2015 Elsevier B.V. All rights reserved.
1. Introduction
Humanity depends on a diverse array of ecosystem services provid-
ed by forests, including the provision of wood fibers, the regulation of
water resources, and the creation of environments conducive to recrea-
tion and wellbeing (MEA, 2005). Over the coming century these and
other forest-derived ecosystem services will need to be sustained
to meet the increasing demands of a growing human population
(Costanza et al., 2014). This task is now made harder by the impacts of
anthropogenic climate change (Staudt et al., 2013). In some regions
anthropogenic climate change is already affecting the delivery of these
services by altering environmental conditions and increasing the fre-
quency and extent of disturbances to production forests (Allen et al.,
2010; Seidl et al., 2014). Concurrently, production forests are increas-
ingly being enlisted as part of mitigation efforts, due to their capacity
to store and sequester atmospheric carbon (Canadell and Raupach,
2008; Lindner and Karjalainen, 2007). Both of these developments are
motivating substantial changes in how forests are being managed and
how the ecosystem services they provide are prioritized (Driscoll
et al., 2012; Lindenmayer et al., 2012a).
As production forests constitute onethird of forest area globally, and
more than half of the forest area in Europe (FAO, 2010; Pawson et al.,
2013), altering the management of even a small proportion of these for-
ests as part of climate change adaptation and mitigation efforts is raising
concerns regarding the potential consequences for forest biodiversity
(Driscoll et al., 2012; Lindenmayer et al., 2012a). Of specificconcernis
that climate change adaptation and mitigation strategies (CCAMS) will
work against the accumulated benefits of recent biodiversity conserva-
tion efforts targeted towards improving habitat availability in produc-
tion forests. These strategies could influence the tree species grown,
the disturbance regimes employed, and the structural attributes
favored. Hence, adaptation and mitigation strategies may be counter-
productive in relation to biodiversity goals, and may even compromise
Biological Conservation 194 (2016) 11–20
⁎Corresponding author.
E-mail address:adam.felton@slu.se (A. Felton).
http://dx.doi.org/10.1016/j.biocon.2015.11.030
0006-3207/© 2015 Elsevier B.V. All rights reserved.
Contents lists available at ScienceDirect
Biological Conservation
journal homepage: www.elsevier.com/locate/bioc

the biodiversity foundations from which many ecosystem services
derive (Cardinale et al., 2012; Schröter et al., 2014; Thompson et al.,
2011).
In this study, we evaluate how climate change mitigation and adap-
tation strategies will impact on habitat availability within Sweden's
production forests. In Sweden, 70% of the land area is covered by forests,
which extend from the temperate zone to the boreal. Sweden is the
world's 6th largest producer of industrial round wood (FAOSTAT,
2013), and relies extensively on the rotational even-aged management
of production forests, approximately half of which are owned by small-
scale private forest owners (SFA, 2014). Norway spruce (Picea abies)is
the most common tree species by volume on productive forest lands
in Sweden, though its dominance is supplanted by Scots pine (Pinus
sylvestris) in the north of the country (SFA, 2014). The production
focus on these two tree species has greatly simplified the tree species
composition of production forest areas, largely to the detriment of
broadleaf tree species. The widespread simplification of Sweden's
forests has likewiseraised concerns regardingthe resilience of these for-
ests to the biotic and abiotic disturbance events projected to increase
over the coming century (Blennow, 2008; Felton et al., 2010a; Jönsson
et al., 2009).These concerns becameparticularly acuteafter a hurricane
(‘Gudrun’) hit southern Sweden in January of 2005, damaging 75 million
cubic meters of primarily spruce-dominated forest, at a cost of USD 2
billion to the forest sector (SCCV, 2007; Svensson et al., 2011).
With respect to the conservation of forest biodiversity, Sweden's
formally protected forests are limited to 3.6% of the productive forest
area, with an additional 4.8% of productive forest lands protected by
voluntary agreements (Johansson et al., 2013a; SFA, 2014). As such,
more than 90% of productive forest lands (volume increment of
≥1m
3
ha
−1
yr
−1
) in Sweden falls outside of these types of protective
frameworks. The large-scale transformation of natural forest cover has
contributed to widespread population declines for many forest depen-
dent species (Berg et al., 1994; Gustafsson and Perhans, 2010;
Gärdenfors, 2010) and nearly 2000 forest-associated species occurring
on the Swedish Red List (Gärdenfors, 2010). Therefore, over the last
two decades specific efforts have been made to increase quantities of
dead wood (Fridman and Walheim, 2000; Schroeder et al., 2011) and
the number of green trees retained within stands at final felling
(Johansson et al., 2013a). However, as these interventions have only
raised deadwood levels in production forests to an average of
8m
3
ha
−1
(N10 cm dbh; Kruys et al., 2013), and require only a limited
number of living trees be retained per hectare at final felling, some
question their adequacy relative to the demands of declining species
(Johansson et al., 2013a). Nevertheless these efforts have increased
the structural diversity of young production forests in relation to pre-
ceding decades (Kruys et al., 2013) andit is likely that these and related
actions will help reduce at least some of the negative effects of forestry
on biodiversity (Gustafsson et al., 2010; Johansson et al., 2013a).
With respect to anthropogenic climate change, current projections
indicate increases in mean annual temperatures in Sweden varying
from 2 to 7 °C by 2071–2100(Kjellström et al., 2014), therebyexceeding
global average projections for the same time period (IPCC 2014).
Projected temperature changes are accompanied by expected increases
in precipitation levels in Sweden of 0–40%, with large variation between
years and decades. Although there is no clear indication regarding
changes to storm frequency or intensity, the associated milder winters
with reduced soil freezing may produce forest conditions more vulner-
able to storm-fellings (Lindner et al., 2014). In response to these and
other adaptation requirements, and as part of global climate change
mitigation efforts, Sweden is already altering production forest man-
agement by, for example, extracting logging residue for bio-energy
production, and encouraging the diversification of production forest
alternatives as a risk-spreading strategy (European Commission, 2007;
Ulmanen et al., 2012).
Within this context of climate change, production forestry, and bio-
diversity conservation we assess the compatibility of a rangeof climate
change adaptation and mitigation strategies (hereafter CCAMS) being
enlisted in Sweden, with this nation's goals for improving habitat for
forest biodiversity. To do so we identified seven CCAMS either already
being implemented in Swedish production forests, or actively consid-
ered or advocated by government institutions and other major stake-
holders. Using the available scientific literature and government
reports we address themotivation for each CCAMS usage, as well as cur-
rent obstacles and incentives for uptake. We then describe each CCAMS
in terms of their resultant implications for three major determinants of
habitat availability in production forests: tree species composition,
forest structures, and spatial/temporal forest patterns. We anchor this
assessment to a baseline reference forest condition provided by
Norway spruce dominated production forests, and to biodiversity
goals provided by government legislation recommendations from the
Swedish Forest Agency and non-governmental forest certification agen-
cies. In the process we develop an internationally relevant framework
for conducting such assessments, provide insights regarding how
CCAMS can readily come into conflict with biodiversity goals, and iden-
tify opportunities for achieving more synergistic outcomes between
CCAMS and biodiversity goals in production forests.
2. Materials and methods
2.1. Selection and assessment of CCAMS
We define CCAMS as changes to the management of production
forests specifically motivated by the need to mitigate climate change,
or to adapt production forests to climate change associated impacts,
risks, and uncertainties (e.g. pestsand pathogen outbreaks, temperature
stress, andwind throw risk; see ‘Reference condition’below). We do not
assess changes to production forest management motivated primarily
by economic or production goals (e.g. increased fertilization), despite
the potential associated carbon sequestration benefits (but see
Schulze et al., 2012). We do not review all possible CCAMS, but identi-
fied seven prominent and distinct examples that are either already
being implemented in Sweden, or are cited as potential strategies for
application in Sweden by government agencies. The CCAMS we assess
include (1) logging residue removal; (2) conversion to short-rotation
hybrid broadleaves; (3) conversion to introduced conifers; (4) conver-
sion to broadleaf tree species; (5) continuous cover forestry; (6) short-
ened rotation times; and (7) prolonged rotation times. We elaborate
on the motivation for the implementation of each CCAMS, as well as
current obstacles and incentives for uptake, before describing the impli-
cations of each CCAMS for tree species composition, forest structures,
and spatio-temporal forest patterns (see Section 2.3 below).
2.2. Reference condition
We use a baseline reference condition representative of standard
production forest conditions in Sweden to anchor our assessment of
the biodiversity implications of CCAMS. This baseline consists of
rotationally cut even-aged stands, dominated by Norway spruce, due
to it being; 1) the most common production tree species in Sweden
by basal area (SFA, 2014), 2) vulnerable to climate change associated
abiotic and biotic disturbance (e.g. wind-damage, spruce bark beetle
(Ips typographus), root rot) (Jönsson et al., 2012; Müller et al., 2014;
Valinger and Fridman, 2011), and 3) a primary target for conversion
to a variety of other forest types under risk-spreading and adaptation
scenarios (Eriksson, 2007; Löf et al., 2012; SCCV, 2007; Skogsstyrelsen,
2011; Swartling et al., 2012; Vulturius and Swartling, 2013). Norway
spruce also exhibits the highest mean annual increment of any native
production tree species in Sweden, indicating the carbon sequestration
contribution of this stand type (SFA, 2014). Norway spruce production
forests are primarily managed as rotationally harvested even-aged
planted stands, which are felled after approximately 60–120 years de-
pending on site conditions. Notably, although spruce-dominated stands
12 A. Felton et al. / Biological Conservation 194 (2016) 11–20

commonly occur throughout Sweden, this production forest type is
dominant only in the south of the country.
2.3. Biodiversity goals
We identified biodiversity goals from Sweden's current legislation
which were relevant to production forests, and either of three key attri-
butes of importance to biodiversity, namely tree species composition,
forest structures, or the spatial and temporal scales of natural distur-
bances (Lindenmayer and Franklin, 2002). Biodiversity goals were
sourced from the Swedish Forestry Act (1993),Sweden'sofficial envi-
ronmental objectives (Swedish Government, 2001), as recommenda-
tions from governmental bodies such as the Swedish Forest Agency
(see Fig. 1), and the criteria and indicators of certification agencies.
The Swedish Forestry Act (1993) states that on forest land (public and
private) biodiversity preservation is to be granted importance equal to
that of timber production. Sweden's broader environmental objecti ves
state that ecosystems and viable populations of native species are to
be safeguarded (Swedish Government, 2001). In addition to official
policies and legislation, we include conservation considerations re-
quired by forest certification agencies. Over 70% of productive forest
land in Sweden is certified according to FSC (Forest Stewardship
Council) and/or PEFC (Program for the Endorsement of Forest Certifica-
tion) standards. The environmental standards of these two agencies are
specifically aimed at providing habitats otherwise lost in production
forests (Gustafsson and Perhans, 2010; Johansson et al., 2013a). The en-
vironmental goals of both the Swedish government and certification
agencies are consistent in terms of promoting an increase in broadleaf
and native tree species. In our conceptual framework (Fig. 1) we refer
to these as “goals for tree species composition”. Current Swedish gov-
ernmental regulations also state that older trees and the availability of
different categories of dead wood in production forestry stands should
increase (referred to as “goals for forest structures”). At broader land-
scape scales, the environmental goals of the Swedish government and
certification agencies recognize the need to increase the proportion of
older forest, limit the area of landscapes clear-cut at a given point in
time, and promote disturbance regimes which are more similar to the
spatial and temporal scales of natural disturbances (referred to as
“goals for spatio-temporal forest patterns”)(seeFig. 1).
2.4. CCAMS consistency or inconsistency with biodiversity goals
Using the published literature, we evaluated whether these CCAMS
are likely to improve or deteriorate forest conditions relative to both
the baseline forest state (i.e. Norway spruce production forests) and
the desired state for forest biodiversity, as codified by governmental
and non-governmental environmental goals. This was done in reference
to goals for tree species composition, forest structures, and spatio-
temporal forest patterns (Fig. 1). For each CCAMS, we also noted the
current extent and rate of increase in implementation (where possible)
using published data, in combination with identified production, finan-
cial, or policy-related obstacles and incentives to uptake. Due to the
extensive information made available by governmental authorities in
Sweden (e.g. Swedish Forest Agency), for this assessment we assumed
that an absence of data for a particular CCAMS is indicative of limited
usage.
Fig. 1. The conceptual framework. For each CCAMS we illustrate the relationship between the reference forest state, Sweden's stated environmental goals, and the resultant potential of
CCAMS to bridgeor extend the intervening gap. The center area symbolizes thebaseline forest state considered, i.e.spruce-dominatedforest under even-aged management. The divided
outer ringdepicts environmentalgoals in terms of three primarycategories of potential adjustmentsto production forests: treespecies composition, forest structures, andspatial/temporal
forest patterns. The outermost text indicates the specific intentions and source of the environmental goals, including Sweden's Forestry Act (FA), Swedish government'senvironmental
goals (SEG), Sweden'sForest Agency (SFA) recommendations, and the standards of the twoprimary forest certificationagencies (FSC, PEFC). When referring to the Forest Act,the relevant
chapter andsection are provided (e.g. 7:8). The spacebetween the centerareas and the outerring indicates the gapbetween the reference foreststate and these environmentalgoals. The
arrows indicate for each of the three categories of environmental goal whether the CCAMS will improve forest conditions relative to the baselineforest state (i.e. arrow directed towards
the respective environmental goal), or deteriorate these conditions (i.e. arrow directed away from this respective environmental goal). In cases where the direction of change is likely
neutral, the spaceis left un-arrowed, whereas in cases where the direction ofchange is indicated,but there are important caveats or uncertainties, the arrows are merely outlined. In
some cases both arrow directions are used with respect to a single environmental goal, to indicate both positive and negative implications of theCCAMS.
13A. Felton et al. / Biological Conservation 194 (2016) 11–20

3. Results
3.1. Logging residue removal
The European Union has agreed on joint mitigation goals for all
member states, with one of the principle objectives being to increase
the proportion of renewable energy by 20% by 2020, compared with
the 1990 base year (European Commission, 2013); a target which
Sweden has already surpassed and plans to exceed (www.
naturvardsverket.se/en/). There is a large potential for using forest bio-
mass as a renewable feedstock forbioenergy production globally (Millar
et al., 2007; Pawson et al., 2013; Spittlehouse and Stewart, 2003),
though the net climate change mitigation benefits may be challenged
or require caveats in some circumstances (see Schulze et al., 2012;
Zanchi et al., 2012). In Sweden this practice is used to supplement the
use of more carbon intensive energy sources for industry and the
heating sector (Ericsson et al., 2004). The biomass is obtained by
extracting logging residues (i.e. branches, tops, and stumps) after final
felling and thinning (Dahlberg et al., 2011). Extraction of branches and
tops has increased during recent years from a planned affected area
of 27,000 ha yr
−1
in 2001, to 117,636 ha yr
−1
in 2013, and currently
is taking place in approximately half of all clear fellings (SFA, 2014).
Stump harvesting is taking place in approximately 1–2% of all clear
fellings (SFA, 2014), and despite limitations put in place by the FSC,
the planned affected area has increased from 597 ha yr
−1
in 2006 to
2336 ha yr
−1
in 2013. The current goal is to further increase the propor-
tion of harvested stands in which logging residues are extracted
(Anonymous, 2009).
Large-scale harvest of logging residues in managed forest landscapes
significantly decreases habitat availability for many dead wood associat-
ed species (Geijer et al., 2014; Ranius et al., 2014; Johansson et al., in
press). Also red-listed species may be affected, since several of them
are using stumps (Jonsell and Hansson, 2011) and the branches and
tops of deciduous trees on clearcuts. Moreover, it exacerbates the extent
of unnatural disturbance caused during clear felling, due to the need to
use heavy machinery when extracting residues, in particular stumps,
because it may thereby adversely affect soil and ground arthropod
fauna, and field layer vegetation (Andersson, 2012; Bouget et al.,
2012). For these reasons logging residue removal is inconsistent with
goals to increase dead wood availability in production forests, as well
as goals to improve natural disturbance emulation (Fig. 2).
3.2. Short-rotation hybrid broadleaf trees
An additional source of wood fibers for bioenergy production is fast
growing broadleaf tree species (Millar et al., 2007; Pawson et al., 2013;
Rytter and Stener, 2011; Spittlehouse and Stewart, 2003). These species
occur in short-rotation stands that are harvested less than 30 years after
establishment (Tullus et al., 2012). In addition to their contribution to
climate change mitigation, their use is also consistent with risk spread-
ing initiatives (Eriksson, 2007; Fransila et al., 2005; Swedish Forest
Agency, 2003) and efforts to reduce stand vulnerability to wind damage.
After the storm Gudrun in 2005, governmentsupport for the diversifica-
tion and increased use of broadleaf dominated production forest con-
tributed to the establishment on 1.5% of storm-felled production forest
lands of hybrid aspen (Populus ×wettsteinii Hämet-Ahti), and 0.3% of
these areas with other poplar species (Wallstedt, 2013).
Whereas the use of hybrid aspen is consistent with the goal to in-
crease the proportion of broadleaf trees, it requires caveats with respect
to the goal to favor native tree species. Despite being phylogenetically
proximate to European aspen (Cervera et al., 2005; MacKenzie, 2010;
Tullus et al., 2012), hybrid aspen is nevertheless a cross between the
European Populus tremula and the American P.tremuloides.Inaddition,
there are concerns regarding the extent of introgression occurring be-
tween hybrid aspen and wild populations of European aspen (Felton
et al., 2013; Koivuranta et al., 2012), whichmay increase due to climatic
change (Koivuranta et al., 2012). The implications of hybrid aspen
stands for forest structures are likewise complex. If, for example, the
nutrient-rich and high pH bark of European aspen is retained in hybrid
aspen, then green tree retention programs and associated dead wood
creation will likely benefit the many organisms associated with
European aspen (Kouki et al., 2004). This is not known as yet. Hybrid
aspen's intensive rate of growth also enables stands to reach heights
in 25 years that are comparable to Norway spruce during normal rota-
tion periods of 60–70 years on similar sites (Johansson et al., 2013b).
Nevertheless, growth rates cannot be assumed to fully compensate for
age-related structural goals, as the development of microhabitats is
also dependent on tree age (Ranius et al. 2009). Furthermore, species
require sufficiently stable substrates through time to establish popula-
tions (Gjerde et al., 2012; Marmor et al., 2011). For these reasons we
find that hybrid aspen stands have aspects that are both potentially con-
sistent and inconsistent with biodiversity goals addressing forest struc-
tures (Fig. 2). We also caveat thisconclusion (i.e. indicate uncertainty or
Fig. 2. The degree of consistency between the seven CCAMS considered or implemented in Sweden, with stated biodiversity goals (see fig. 1). CCAMS are placed solely in relation to a
horizontalgradient (i.e. there is noy-axis) indicating from left to right their relative consistency with biodiversity goals (decreasing), asdescribed inthe conceptual framework section
and accompanying text for each CCAMS.
14 A. Felton et al. / Biological Conservation 194 (2016) 11–20

context-dependent aspects in figure outcomes), because recent evi-
dence indicates that the exclusion of ungulates via fencing, which is rec-
ommended in these stands, is, allowing for a more diverse and
structural complex understory than is achieved in unfenced spruce
monocultures of the same age, with associated apparent benefits for
some forest-dependent taxa (Lindbladh et al., 2014). The short rotation
times of hybrid aspen stands are however clearly inconsistent with
spatio-temporal goals to increase the proportion of older forest and
emulate natural disturbance regimes (Fig. 2).
3.3. Introduced conifers
Introduced tree species provide a means of adapting production
forests to climate change, as the global pool of tree species can thus be
sourced for those species with a reduced vulnerability to climate-
associated risks (Bolte et al., 2009; Kolström et al., 2011; Millar et al.,
2007; Spittlehouse and Stewart, 2003). Compared to many other EU
countries, only a low proportion (~ 1.5% by volume) of Sweden's
production forests consist of introduced tree species (SFA, 2014). The
potential for increasing their use was raised by the Swedish Commission
on Climate and Vulnerability as a risk spreading response to the uncer-
tainties associated with climatic change (SCCV, 2007). After the storm
Gudrun, diversification efforts were supported by government aid
for the establishment of introduced conifers (Wallstedt, 2013). As a
result, about 5% of the storm felled area was planted with hybrid larch
(Larix eurolepis/L. marschlinsii), as well as 1.4% with Sitka spruce
(Picea sitchensis) and 0.1% with Douglas fir(Pseudotsuga menziesii)
(Wallstedt, 2013). The extent of uptake is however limited by national
legislation (Skogsvårdsförordningen 1993:1096), by which the Forest
Agency has also suggested regulating the spatial extent (≤25% of pro-
duction unit area), and distribution (limits proximity to national parks
and mountainous areas) of introduced tree species (Norén and
Ringagård, 2009). Furthermore, theFSC limits the allowable usage of in-
troduced tree species in newly planted stands to no more than 5% of
productive forest area per estate (FSC, 2010).
The use of introduced conifer species is inconsistent with goals for
tree species composition, as it contradicts the prioritized use of native
tree species (Fig. 2). With respect to forest structures, their use may
severely curtail many of the intended benefits of within stand conserva-
tion actions such as high stump creation, as well as dead wood and
green tree retention (Felton et al., 2013). This is because introduced
tree species often fail to overlap with native tree species in the types
of resources or secondary plant compounds produced (Gossner and
Ammer, 2006; Lieutier, 2006), and are therefore depauperate in some
of the faunas they support (Gossner et al., 2007; Kennedy and
Southwood, 1984; Peterken, 2001). However, there are indications
that the relatively sparse and deciduous nature of hybrid larch canopies
can favor a relatively rich understorey plant community (Barbier et al.,
2008; Skogsstyrelsen, 2009). We therefore include a caveat with respect
to the adverse implications from the use of introduced tree species in
relation to forest structures (Fig. 2). With respect to goals for spatio-
temporal forest patterns, species like hybrid larch are likely to be har-
vested over a shorter rotation than Norway spruce, whereas Douglas
fir is often managed over longer rotation periods (Rytter et al., 2013).
For these reasons the direction of impact on spatio-temporal forest pat-
terns is dependent on the introduced conifer planted, and therefore
caveated (Fig. 2).
3.4. Conversion to native broadleaved tree species
The diversification of production forests is a recommended means of
spreading the risk posed by climate change (Bolte et al., 2009; Kolström
et al., 2011; Millar et al., 2007; Ogden and Innes, 2007; Pawson et al.,
2013). In Sweden, government reports have also raised or advocated
risk spreading strategies for addressing climate change, including the
diversification of production forests at stand and landscape scales
(Eriksson, 2007; Fransila et al., 2005; Swedish Forest Agency, 2003).
As part of these efforts, emphasis has been placed on increasing the
use of broadleaf tree species (Swartling et al., 2012) to reduce vulnera-
bility to wind felling (Wallstedt, 2013) and drought damage (SCCV,
2007). A related emphasis has been placed on the increased use of mix-
tures (Swedish Forest Agency, 2014; Valinger and Fridman, 2011;
Vulturius and Swartling, 2013), specifically those combining birch
(Betula spp.) and Norway spruce (Felton et al., 2010b; SCCV, 2007).
After the storm Gudrun, close to 20 million Euro was specifically allocat-
ed to the increased planting of native broadleaf stands or broadleaf-
dominated mixtures; of these funds only 10% was used by landowners.
This low level of uptake was due in-part to browsing related concerns
and the associated costs and complications of protective measures
(Ulmanen et al., 2012; Wallstedt, 2013). The net result was the estab-
lishment of 1% of storm-felled area with birch, and 1% with oak
(Quercus spp.), with an additional 1% of felled area established with
European beech (Fagus sylvatica), wild cherry (Prunus avium), small-
leafed lime (Tilia cordata)orNorwaymaple(Acer platanoides).
Increased use of native broadleaf tree species is clearly consistent
with the goals for tree species composition of the Swedish Forest
Agency, PEFC, and FSC (Gustafsson et al., 2010; Gustafsson and
Perhans, 2010; Johansson et al., 2013a). However, rotation lengths for
birch can be shorter than those used when growing Norway spruce —
depending on production aims (Dahlberg et al., 2006; Hynynen et al.,
2010); whereas rotation lengths for oak may be as long as 150 years
(Drössler et al., 2012). As such we acknowledge that rotation lengths
for broadleaf stands may be either consistent or inconsistent with
goals for the occurrence of older trees, but caveat the inconsistency
due to the benefits of increasing broadleaf dead wood in production for-
ests (Fig. 2). Furthermore, this caveat is also necessary due to the varia-
tion among tree species in the age at which microhabitats are formed
(Uliczka and Angelstam, 1999). Whether changes to spatio-temporal
forest patterns will be consistent with current goals will likewise vary
due to associated distinctions in the rotation lengths used for different
broadleaf species.
3.5. Continuous cover forestry
Continuous cover forestry (CCF) involves uneven-aged forest man-
agement with selective felling to maintain forest cover (Kuuluvainen
et al., 2012). Norway spruce stands managed using CCF are expected
to have lower risks of some types of damage projected to increase
under climate change, such as pest outbreaks and wind throw (O'Hara
and Ramage, 2013; Vulturius and Swartling, 2013), and are thus
discussed as a climate change adaptation option internationally
(Kolström et al., 2011), and by Swedish government authorities
(Swartling et al., 2012; Swedish Forest Agency, 2014; Vulturius and
Swartling, 2013). Whereas stands actively managed using CCF are
very rare in Sweden, it has been estimated that about one million hect-
ares of productive forest may be suitable for its application (Dahlberg,
2011). The adoption of CCFis however hampered by concerns regarding
production losses, harvesting costs, and the lack of expertise and
infrastructure necessary for its adoption (Cedergren, 2008). Further-
more, recent assessments suggest a potentially heightened risk in
uneven-aged Norway spruce stands of Heterobasidion root rot (Piri
and Valkonen, 2013).
Relative to clear-cutting approaches, the application of CCF in
Norway spruce-dominated stands is expected to better maintain late
successional forest conditions and associated micro-climates through-
out the management cycle, specifically in relation to the continued
availability of relatively mature trees and coarse woody debris
(Atlegrim and Sjöberg, 2004; Lähde et al., 2002). The specifics will how-
ever depend on how conservation measures (e.g. retention trees) are
applied (Gustafsson et al., 2012). Likewise, the use of CCF would limit
the area of clear-cuts, more evenly distribute the proportion of older
forests in the landscape at finer spatial grains, and is thus considered
15A. Felton et al. / Biological Conservation 194 (2016) 11–20

more consistent with the spatial scale of natural disturbance regimes in
this region, and thus spatio-temporal goals, than clear-cutting (see
Kuuluvainen and Aakala, 2011; Kuuluvainen et al., 2012). However, it
should be noted that many naturally regenerating broadleaftree species
in Fennoscandian production forests are pioneer species, which would
potentially experience poor regeneration in Norway spruce standsman-
aged usingCCF. CCF could thereby result in afurther increase in Norway
spruce abundance (Widenfalk and Weslien, 2009), and thus detract
from tree species composition goals. This concern is however likely to
depend, for example, on the spatial grain of disturbance employed,
and we therefore caveat this concern (see Fig. 2).
3.6. Shortened rotation times
Norway spruce's susceptibility to climate-associated storm damage,
bark beetle outbreaks, and drought are noted in governmental assess-
ments (SCCV, 2007). A range of international studies (Bolte et al.,
2009; Kolström et al., 2011; Millar et al., 2007; Ogden and Innes,
2007; Pawson et al., 2013; Spittlehouse and Stewart, 2003) and govern-
mental reports lend support to shortening rotation times as a means to
mitigate such risks (Swartling et al., 2012; Vulturius and Swartling,
2013). Furthermore, post-Gudrun assessments of stand vulnerability
to wind throw indicate that storm damage increases with the height
of the stand (Valinger and Fridman, 2011), and thus with longer rota-
tion lengths. More general assessments have also linked increased rota-
tion length to stand susceptibility to both abiotic and biotic disturbance
events (Jactel et al., 2012). Although uptake is difficult to estimate, there
are indications that in response to perceived risks, many forest owners
are harvesting earlier (Vulturius and Swartling, 2013). Furthermore,
the leading forest owner's association in southern Sweden has recently
provided forest management plans to over 50,000 owners indicating
that in some cases rotation lengths could be shortened by approximate-
ly 10–15 years in Norway spruce dominated stands, specifically to
reduce climate associated risks (Södra, 2012).
Shortened rotation times will reduce the availability of older trees.
Furthermore, the availability of coarse woody debris will decrease
with reduced rotation lengths (Ranius et al., 2003; Weslien et al.,
2009). For these reasons shortened rotation times are inconsistent
with goals relating to forest structures (Fig. 2). Likewise, shortened rota-
tion times are inconsistent with spatio-temporal goals focused on in-
creasing the proportion of older forest and to better emulate natural
disturbance regimes (Kuuluvainen and Aakala, 2011). Reduced rotation
lengths would likewise increase the proportion of forest area in a clear-
cut state. There are no clear repercussionsfor tree species composition.
3.7. Increased rotation times
Rotation times may also be lengthened to increase carbon storage
within production forests (Jandl et al., 2007; Pawson et al., 2013).
This possibility has been raised by Swedish forest management organi-
zations (Skogssällskapet, 2009), ENGOs (SNF and WWF, 2011)and
government funded projects (Kaipainen et al., 2004). The application
of prolonging rotation times in Norway spruce stands is however
hampered by concerns regarding a drop in annual harvest volumes,
discounting effects (Skogssällskapet, 2009), and potential production
losses from increased storm damage (Valinger and Fridman, 2011)or
bark beetle outbreak (Fettig et al., 2007; Schlyter et al., 2006).
The implications of prolonged rotation times arethe reverse of those
indicated for shortened rotations (see above). Lengthened rotation
times will increase the availability of older trees, as well as the availabil-
ity of coarse woody debris (Ranius et al., 2003; Weslien et al., 2009). For
these reasons, lengthened rotation times are consistent with goals for
forest structures (Fig. 2). Likewise,lengthened rotation timesare consis-
tent with spatio-temporal goals to increase the proportion of older for-
est, due to their relative consistency with natural disturbance regimes
(Kuuluvainen and Aakala, 2011), and the reduced proportion of land
in a clear-cut state at any one time. No clear repercussions for tree spe-
cies composition occur.
4. Discussion
The CCAMS we evaluated are regional variants of those being con-
sidered and implemented in production forests worldwide, including
altered rotation lengths, short-rotation bioenergy stands, the increased
use of introduced tree species, continuous cover forestry, logging resi-
due extraction, and risk-spreading approaches (Bolte et al., 2009;
Kolström et al., 2011; Millar et al., 2007; Ogden and Innes, 2007;
Pawson et al., 2013). We found that these CCAMS often cameinto direct
or partial conflict with national biodiversity goals and initiatives to
restore habitat availability in production forests. These conflicts arose
irrespective of whether CCAMS were primarily associated with mitiga-
tion or adaptation aims, and despite the fact that the reference condi-
tion, planted spruce forest, set a relatively low threshold for achieving
biodiversity benefits. Furthermore, the widespread implementation
of CCAMS which are inconsistent with Sweden's biodiversity goals
appears to be taking place relatively unimpeded, as demonstrated by
the extensive adoption of logging-residue removal (Geijer et al.,
2014). Our findings therefore support the contention that efforts to
address anthropogenic climatic change involving production forests,
can readily come into conflict with efforts to address biodiversity loss
(Lindenmayer et al., 2012a).
Our findings also reveal that several of the CCAMS evaluated were
synergisti c with Sweden's biodi versity goals, and t hus provide an appar-
ent means of meeting the challenge of both anthropogenic climate
change and biodiversity conservation (Pawson et al., 2013; Thompson
et al., 2011). For example, the conversion of Norway spruce stands to
broadleaf-dominated mixtures is consistent with all three categories
of the biodiversity goals assessed. In addition, this strategy provides a
potential means of reducing the risk of wind damage (Wallstedt,
2013), and pest and pathogen outbreaks (Jactel and Brockerhoff,
2007), while concurrently increasing adaptive capacity by providing
managers with alternative directions for stand development (Lindner
et al., 2010). Nevertheless, we found no indication that this or other
CCAMS identified as being consistent with achieving biodiversity
goals, are being adopted beyond a limited extent. Identified causal fac-
tors underlying their limited uptake varied depending on the CCAMS
considered, but included increased risk of browsing damages, produc-
tion and economic losses, increased harvesting costs, and increased
risks of storm damage (Cedergren, 2008; Skogssällskapet, 2009;
Ulmanen et al., 2012; Wallstedt, 2013).Wesuggestthatsomeofthese
obstacles could be overcome using targeted interventions to reduce
the concern (e.g. actions to limit browsing damage), or through the
use of policy incentives favoring the CCAMS' uptake (e.g. subsidies)
(Puettmann et al., 2015). In either regard, we donot see present circum-
stances asfavorable to the widespread useof those CCAMS identified as
being most consistent with biodiversity goals.
Instead, we suggest that those CCAMS which are inconsistent with
biodiversity goalswill continue to be used as long as forest owners per-
ceive them to reduce risks, increase production capacity, or associated
economic returns. For example, the extraction of logging residues for
bioenergy production provides an additional source of income (Ranius
et al., 2014), which probably aids its widespread adoption. Economic
and production incentives can also be enlisted to support the expanded
use of introduced treespecies (Nilsson et al., 2011), and the shortening
of rotation times in Norway spruce stands (Fries et al., 2015; Södra,
2012). Such incentives could likewise motivate the simultaneous
adoption within a single stand of several CCAMS deemed inconsistent
with biodiversity goals (e.g. both shortened rotation times and logging
residue extraction). Consistently, several previous assessments have
concluded that production goals continue to take precedence over envi-
ronmental goals in Sweden (see Ulmanen et al., 2012), despite the
Swedish Forestry Act (1993) stating that the goals of timber production
16 A. Felton et al. / Biological Conservation 194 (2016) 11–20

and biodiversity preservation are to be granted equal importance in
Swedish forest lands. Likewise, among Swedish private forest owners,
and in particular among their advisors, biodiversity is consistently placed
below production in terms of management priorities (Kindstrand et al.,
2008). In either regard, the implementation of CCAMS which are incon-
sistent with biodiversity goals, but which provide financial or production
benefits, appears likely to continue under present circumstances.
It could be argued that the local or regional-scale biodiversity costs
of CCAMS are justifiable losses, due to the necessity of enlisting produc-
tion forests in global mitigation efforts, and the extensive impact that
anthropogenic climate change will itself have on biodiversity if emis-
sions are not reduced. This argument has in fact been put forward in
Sweden, where some forest industry representatives argued that a
focus on biodiversity consideration in production forests could impair
the contribution of production forests to climate change mitigation
efforts (see Ulmanen et al., 2012). Notably, this point is not limited to
mitigation strategies per se, as adaptation strategies in production for-
ests are also essential if the contribution of production forests to carbon
sequestration and storage is to be continued (Kolström et al., 2011;
Lindner et al., 2010). Anthropogenic climatic change is projected to
greatly exacerbate global extinction rates (Foden et al., 2008; Thomas
et al., 2004), due to resultant changes to habitat (Julliard et al., 2004;
Warren et al., 2001), species' physiological limitations (Boyles et al.,
2011; Oswald and Arnold, 2012), altered species interactions (Van der
Putten et al., 2010), ocean acidity (Kroeker et al., 2013; Pandolfiet al.,
2011), increased risk of disease (LaPointe et al., 2005), and synergistic
interactive effects (Brook et al., 2008; Driscoll et al., 2012). Thus, the
suggestion that climate change mitigation efforts take priority over
local conservation concerns is at least superficially persuasive.
However, the underlying suggestion that forest managers must
choose between theneed for climate change adaptation and mitigation,
versus the necessity of biodiversity maintenance in production forests,
overlooks the importance of biodiversity in the resilience, regulation
and function of production forest ecosystems (Rist et al., 2014;
Thompson et al., 2011). Furthermore, as our results demonstrate, an
either/or approach to forest management is often a false dichotomy.
First, we have identified several climate change adaptation and mitiga-
tion strategies, including continuous cover forestry, conversion to native
broadleaf tree species, and increased rotation times, which are largely
consistent with biodiversity goals and improving habitat availability in
production forests. Second, we suggest that CCAMS deemed incompat-
ible with biodiversity goals may nevertheless be implemented without
incurring biodiversity losses, if conservation interventions are increased
or otherwise modified to ensure adequate compensation. A recentstudy
demonstrates, for example, that habitat loss for a range of dead-wood
dependent taxa caused by logging residue extraction, can be compen-
sated for by the creation of additional high stumps in stands (Ranius
et al., 2014). Likewise, green tree retention practices could be modified
to increase the availability of older larger trees and dead wood in short-
rotation stands (Nilsson et al., in press). Third, as our findings demon-
strate, CCAMS vary widely in their resultant consequences for tree
species composition, forest structures, and spatio-temporal forest
patterns. We therefore suggest that adverse biodiversity impacts could
also be avoided by combining at landscape scales the use of those
CCAMS complementary in terms of habitat provision. This concept is
similar to that advocated as part of differentiated land-use approaches,
in which distinct forest land-use categories are combined at landscape
scales to achieve conservation and economic goals (Côté et al., 2010).
Because this approach can involve the diversification of production
forests, it also provides the potential benefit of increasing the adaptive
capacity of the production forest system overall (Lindner et al., 2010).
In many cases this approach will however also require a shift from
stand-level to landscape-level management, which brings with it a
number of potential obstacles to implementation (Pawson et al.,
2013) especially in regions, like southern Sweden, dominated by a
large number of small-scale forest owners (Gustafsson et al., 2015;
McDermott et al., 2010). In addition, we caution that some CCAMS
may cause ecological problems of sufficient severity (pest and pathogen
outbreaks, invasiveness) to negate the effectiveness of such compensa-
tory approaches. In cases where the use of CCAMS is associated with
severe or likely irreversible ecological risks (see Felton et al., 2013),
the only adequate measure may be to avoid or limit the use of such
CCAMS altogether.
We suggest that the identified conflict between some of the CCAMS
considered and a region's biodiversity goals is unlikely to be isolated
to Sweden. This expectation stems from the similarities in the habitat
and resource requirements of forest-associated biodiversity irrespective
of the forest biome considered. These similarities are reflected in
the consistent relevance for forest biodiversity of old large trees
(Lindenmayer et al., 2012b; Lindenmayer et al., 2014), tree cavities
(Remm and Lõhmus, 2011), coarse woody debris (Stokland et al.,
2012), natural disturbanceregimes (Attiwill,1994), and habitat connec-
tivity (Lindenmayer and Fischer, 2006). Thus, the implementation of
CCAMS which we expect to diminish such attributes in Sweden's
production forests (e.g. shortened rotation times, logging-residue
removal, the use of introduced conifers) can likewise be expected to
come into conflict with the habitat requirements of many forest-
dependent species globally.
Nevertheless, there are forest systems for which our outcomes will
be less relevant. For example, our baseline reference condition (i.e.
rotationally felled even-aged Norway spruce dominated stands), and
most of the CCAMS assessed, involve relatively highlevels of anthropo-
genic input (see Rist et al., 2014). This limits the relevance of some of
our findings in forest systems with small silvicultural interventions.
We also caution that, even just within the Swedish context, geographi-
cal gradients in climate, natural disturbance dynamics, tree species
composition, and historical and current land-use (Gustafsson et al.,
2015), will alter the extent to which CCAMS affect habitat availability.
The more extensive biogeographical differences found in other forested
regions of theworld will likewise dictate the biodiversityimplications of
these CCAMS. We also suggest that consideration is given to the biodi-
versity costs of retaining production forests which are prone to increas-
ingly severe disturbance regimes under anthropogenic climate change
(Lindner et al., 2014). Relatedly, the retention of current forest compo-
sition can also be inconsistent with the habitat requirements of future
species pools (Felton et al., 2014). Though these are all important con-
siderations, they do not override the general principles conveyed by
our framework: if, in relation to baseline forest conditions, and identi-
fied environmental goals, the implementation of a particular CCAMS
results in the loss of natural tree species composition or diversity (e.g.
conversion to introduced conifers), the removal of important habitat
structures (e.g. logging residue removal), or causes further departure
from natural disturbance regimes (e.g. shortened rotation times), habi-
tat degradation for native taxa can be expected.
Our conceptual framework thereby provides an effective pathway
for identifying both synergies and tradeoffs among CCAMS and forest
biodiversity goals. We also see its potential for more widespread appli-
cation as a policy development tool, whenever changing management
practices have the capacity to adversely affect environmental condi-
tions. The necessary conditions for applying this framework include;
1) a baseline reference forest condition from which CCAMS can be
contrasted; 2) codified biodiversity goals providing regionally relevant
targets for increasing biodiversity in production forests; and 3) expert
evaluations of the biodiversity implications of CCAMS in relation
to this baseline and identified goals. A key assumption underlying our
conceptual framework (Fig. 1) is that the identified biodiversity goals
are relevant to closing the gap between thehabitats provided in produc-
tion forests and those required by forest dependent species. If so, then
our approach should provide an a priori means of evaluating the biodi-
versity implications of CCAMS. We suggest that once such an evaluation
is completed, the resultant insights can be added to the broader suite of
considerations necessary for determining the respective costs and
17A. Felton et al. / Biological Conservation 194 (2016) 11–20

benefits of implementing a particular CCAMS (Felton et al., in press;
Roberge et al., in press).
5. Conclusions
We found that climate change adaptation and mitigation strategies
can readily come into direct or partial conflict with biodiversity goals
and related efforts to restore habitat within production forests. The
widespread implementation of these CCAMS risks reversing some of
the important steps taken by societies to combat biodiversity loss in
production forests. Nevertheless, we have also identified feasible path-
ways for achieving both goals, for example by i) prioritizing the use of
those CCAMS consistent with biodiversity goals, ii) by adjusting conser-
vation interventions to compensate for the use of CCAMS not consistent
with biodiversity goals, or iii) by combining at landscape scales the use
of those CCAMS complementary in terms of habitat provision. We sug-
gest that targeted regulations or policy incentives will be needed to
ensure that such pathways are taken, and that those CCAMS which
take us a step backwards in terms of biodiversity conservation are
limited in implementation, adequately compensated for, or both.
Acknowledgements
We sincerely thank three anonymous reviewers for their construc-
tive comments and suggestions. AF, JMR, TR, and LR were funded in
part by Future Forests (FOR 2008/019), a multi-disciplinary research
program supported by the Foundation for Strategic Environmental Re-
search (MISTRA). JMR acknowledges funding from the Kempe Founda-
tion (SMK-1339).
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