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Ecological implications for assisted migration in Canadian forests



Forest ecosystems are already being impacted by climate change as natural migration rates are outpaced by rapidly changing climate conditions. Human-assisted migration has been proposed as a potential management option to maintain optimal health and productivity of Canada's forests; however, a better understanding of the ecological implications is needed to inform decision-making on assisted migration (AM). This paper examines the ecological constraints and consequences of AM, and discusses options for their mitigation at three scales: translocation over long distances (assisted long-distance migration), translocation just beyond the range limit (assisted range expansion), and translocation of genotypes within the existing range (assisted population migration). From an ecological perspective, we find that AM is a feasible management option for tree species and that constraints and consequences can be minimized through careful application of available knowledge and tools.
Ecological implications for assisted migration
in Canadian forests
Richard Winder1, Elizabeth A. Nelson2, and Tannis Beardmore3
Forest ecosystems are already being impacted by climate change as natural migration rates are
outpaced by rapidly changing climate conditions. Human-assisted migration has been proposed as a
potential management option to maintain optimal health and productivity of Canada’s forests; however,
a better understanding of the ecological implications is needed to inform decision-making on assisted
migration (AM). This paper examines the ecological constraints and consequences of AM, and
discusses options for their mitigation at three scales: translocation over long distances (assisted long-
distance migration), translocation just beyond the range limit (assisted range expansion), and
translocation of genotypes within the existing range (assisted population migration). From an
ecological perspective, we find that AM is a feasible management option for tree species and that
constraints and consequences can be minimized through careful application of available knowledge and
Key words: Adaptation, assisted migration, climate change, colonization, constraints, ecology,
establishment, mitigation, translocation, trees
Les écosystèmes forestiers sont déjà touchés par les changements climatiques, et les taux naturels de
migration ont du mal à suivre l'évolution rapide des conditions climatiques. La migration assistée a été
proposée comme option d’aménagement forestier permettant le maintien de la santé et la productivité
optimales des forêts du Canada. Cependant, une meilleure compréhension des implications écolo-
giques est nécessaire pour éclairer le processus décisionnel en matière de migration assistée. Cet article
examine les contraintes écologiques et les conséquences de la migration assistée, et discute des options
d'atténuation à trois échelles d’intervention: la translocation sur de longues distances (migration assis-
tée à longue distance); translocation juste au-delà de la limite de distribution (extension assistée de la
distribution), et la translocation de génotypes au sein de la distribution actuelle (migration assistée de la
population). Du point de vue écologique, nous constatons que la migration assistée est réalisable pour
des espèces d'arbres, et qu’une application rigoureuse de la connaissance et des outils disponibles peut
permettre d’atténuer les contraintes et les conséquences.
Mots clés: adaptation, arbres, atténuation, changements climatiques, colonisation, contraintes, écolo-
gie, établissement, migration assistée, translocation
1 Natural Resources Canada, Canadian Forest Service - Pacific Forestry Centre, 506 Burnside Road West, Victoria, British
Columbia, V8Z 1M5. Email:
2 Natural Resources Canada, Canadian Forest Service - Headquarters, 580 Booth Street,
Ottawa, Ontario, K1A 0E4. Email:
3 Natural Resources Canada, Canadian Forest Service - Atlantic Forestry Centre, 1350 Regent Street, P.O. Box 4000, Fre-
dericton, New Brunswick, E3B 5P7. Email:
Climate is projected to change dramatically in the coming decades, significantly impacting biological
systems (IPCC 2007). Forest growth is dependent on climate and environmental conditions, responding
closely to changes in temperature and water availability. Climate change in past epochs has resulted in
the natural migration of forests via an array of dispersal mechanisms (e.g. Webb 1987), but will the
pace of future natural migration match the pace of current climate change? Estimated rates of tree
migrations vary with species and location. Estimates of migration rates during the post-glacial
Holocene, range from 0.15 to 2 km y-1 (Huntley 1991; Pakeman 2001; Woodall et al. 2009). In the
current epoch, there are numerous projections of potential range shifts for tree species, particularly
those with commercial importance (e.g. Iverson and Prasad 1998; Crumpacker et al. 2001, Malcolm et
al. 2002, Hamann and Wang 2006; McKenney et al. 2007; Rehfeldt and Jaquish 2010). Regarding
actual rates of climate change, estimates again vary. Loarie and colleagues (2009) calculated that
climate zones or “envelopes” will shift 0.42 km y-1 on a global basis in the IPCC A1B moderate
emissions scenario, although the rate of change may be much greater (ca. 10 km y-1) at northern
latitudes (Hamann and Wang 2006). When habitat fragmentation (Collingham and Huntley 2000) and
other factors (Malcolm et al. 2002) are taken into account, model results do not paint an optimistic
picture for forests. For example, a model for 130 North American species indicated declines in forest
area of 12% to 58% (McKenney et al. 2007). Although forests may be resilient to climate change over
longer time scales, the projected changes are occurring at a much higher rate than modern forests have
experienced (IPCC 2007).
The human-assisted migration of tree species or populations has been proposed as a potential forest
management response to maintain optimal productivity and diversity under continued climate change
(O’Neill et al. 2008). This paper discusses the ecological considerations for assisted migration of tree
species in Canada, covering both the ecological constraints that could potentially limit successful
colonization, as well as the potential negative ecological consequences of translocation efforts. This
summary draws from relevant research pertaining to tree migrations across a range of ecosystems,
countries, and continents, while focusing on implications for North American forest management.
Within the forestry context, we will refer to three principle modes of assisted migration:
Assisted long-distance migration (or “assisted colonization” sensu Gray et al. 2011) describes human-
assisted movement well outside of the established species’ range into regions that are inaccessible
through natural dispersal processes, often through inhospitable areas or over large geographic barriers
such as oceans, lakes, deserts and mountain ranges. As the most aggressive option for assisting the
migration of plant species, it is likely the most difficult to employ with success. It is primarily
proposed as a last resort for endangered species with highly specialized habitats that are not easily
replaced or for species with current or projected disjunctive distributions (McLachlan et al. 2007,
Hoegh-Guldberg et al. 2008). Humans have already translocated tree species for long distances in
many parts of the world (Mack, 2001); although these translocations were not motivated by concerns
about climate change, they nevertheless provide relevant experience in this mode of assisted migration
and are discussed in greater detail later in this paper.
Assisted range expansion involves the human-assisted movement of species just outside of their range
to facilitate or mimic natural range expansion, and therefore occurs in an area adjacent to the native
range of the species in question. At the “advancing” limit of a range, deliberate efforts to extend the
range of a tree species may help to ensure the long-term survival of the tree species, its climate-adapted
alleles, and/or the size of its population, as well as facilitating maintenance or even enhancement of
species’ productivity and health. This form of assisted migration would pertain to situations in which
nearby habitat is available, but species would be unlikely to colonize optimal future habitat without
human aid due to limited dispersal abilities and/or highly fragmented habitat (Vitt et al. 2010). In
theory, physical factors (soil types, topography, photoperiod) and biotic factors (community
composition) of recipient ecosystems are not far beyond adaptive constraints occurring in the native
range of the species and may, in fact, correspond to the historical range of the species on longer time
scales (Hunter 2007; Hoegh-Guldberg et al. 2008). In a forestry context, assisted range expansion
could be invoked to maintain forest dominance on suitable landscapes, ensuring coverage of stable and
productive forests under conditions of continued climate change.
Assisted population migration involves the human-assisted movement of populations (genotypes)
within a species’ established range in order to maximize adaptation to climate change. Assisted
population migration can have the general goal of improving climate-related genetic diversity across
the natural range of the species, or a more targeted goal of matching projected climate scenarios with
“pre-adapted” populations from within the species’ range. Where forests are composed of a few
dominant tree species, there is a concern that climate impacts on those species could have a significant
impact on their overall biological and economical productivity (O’Neill et al. 2008). The practice of
assisted population migration has therefore already been implemented in forestry operations of several
Canadian provinces, with southern seed sources being planted further north to improve productivity
under warming conditions (q.v. Pedlar et al. in this issue).
There are several situations where the distinction between these three types of assisted migration
becomes somewhat ambiguous; for example, range expansion efforts are likely to require the
subsequent and ongoing translocation of appropriate genotypes as climate continues to change (assisted
population migration), or may involve translocation over inhospitable areas in some portions of the
range limit (assisted long-distance migration). In a forest management context, the three modes also
share a common set of objectives. The first would be to maintain the optimal productivity or “vigor” of
forest ecosystems and their self-sustaining cycles and services. The second would be to maintain the
optimal diversity of those same ecosystems and attributes in order to promote their “resiliency” to
various disturbances. In either case, optimal does not necessarily mean maximal; there are trade-offs
between these two objectives, as it very difficult to maximize one parameter without impacting the
other (Odum 1974). There are numerous studies that project or model climatic impacts on ecosystem
structure and productivity (reviewed by Johnston et al., 2010) within specific geographic or adaptive
contexts (e.g., Reed and Desanker 1992; Hamann and Wang 2006); these can provide valuable insight
into decision-making on assisted migration.
Beyond timber production, assisted migration has implications for the vital role of forests in
biogeochemical cycling. For example, northern forests are thought to be an important carbon sink
(Goodale et al. 2002; Pan et al. 2011); there are concerns that less robust forests could negatively
impact national carbon budgets (Jandl et al. 2007). Some studies suggest that carbon sinks should
increase as forests expand into tundra; however, these projections are long-term and based on
concomitant assumptions that forest productivity would increase (White et al. 2000). In the short-term,
climate effects may actually decrease forest cover and productivity (McKenney et al. 2007, Kurz et al.
2008); among various options, assisted migration could help to ameliorate this situation. Although the
above-mentioned goals are “ecologically-oriented objectives”, ecosystems do not have goals. Clearly,
people have a vested interest in the outcomes, where there are implications for human livelihoods and
social values (q.v. Aubin et al. in this issue).
Concerning the application of assisted migration, a variety of decision-making frameworks have been
proposed to help prioritize and/or determine the scope of assisted migration efforts (Hoegh-Guldberg et
al. 2008; Richardson et al. 2009). Commonly, the frameworks involve an assessment of species or
habitat vulnerability to climate change and a determination of which species may be potential
candidates for translocation; candidates are screened for both their need for assistance and the risk they
pose to recipient ecosystems. Hoegh-Guldberg et al. (2008) proposed a decision-making framework
that considers the risk to the species, practical issues (the feasibility of translocation), and the suitability
of assisted migration as a solution, considering biological and socioeconomic issues. In the particular
case of tree species impacted by climate change, relevant factors that will determine outcomes include
dispersal ability, genetic variability (adaptive capacity for necessary traits), and phenotypic plasticity
(Aitken et al. 2008). Outcomes for trees are also dependent on the spatial and temporal scale of assisted
migration efforts. With these specifics in mind, we present a slightly modified version of the Hoegh-
Guldberg et al. (2008) framework in Figure 1, adapted for tree species, and taking into account the
factors of scale that are important in forest management. The modified framework incorporates the
three principal modes of assisted migration that we have outlined, each carrying different ecological
implications. Decisions regarding assisted migration will have to be made using the best available
ecological knowledge (McLachlan et al. 2007; Hoegh-Guldberg et al. 2008; Vitt et al. 2010), with a
recognition that the ecological implications and outcomes of the three different modes of assisted
migration will differ substantially. There is a wealth of knowledge on the ecological constraints
inherent in species introductions, however, this ecological knowledge has a conservation bias, with the
intended outcome being the maintenance or restoration of an existing or pre-existing ecosystem. The
use of decision frameworks (such as the one presented in Figure 1) is essential to ensure that
appropriate ecological information is identified and applied in each assisted migration approach. In this
paper, we have attempted to collect and present a brief overview of this ecological information, with an
overall goal of informing decisions concerning the potential implementation of assisted migration in
Canadian forests.
Figure 1. A framework to inform decision-making for assisted migration of Canadian tree species (adapted from Hoegh-Guldberg et al. 2008). The first step is to assess the vulnerability of
the focal species, differentiating between those species that are at low, medium or high vulnerability under continued climate change. Species with low vulnerability are best served by
continued efforts to improve conventional forest management. Species with moderate vulnerability are likely candidates for assisted population migration efforts, which can enhance
resilience to climate change. Species with high vulnerability may benefit most from human assistance at range limits, either through the facilitation of natural migration and
establishment or through more direct human-assisted range expansion. In cases where range expansion is insufficient, as in the case of large geographical barriers at the species’ range
limit, assisted long-distance migration is also a consideration. Ex situ conservation through the creation of seed backs can provide “ecological insurance”, as well as support for future
assisted migration activities. Regardless of the management options considered, species monitoring is essential during and following implementation of any assisted migration
approach, and species vulnerability must be periodically assessed to determine whether the approach chosen was the appropriate one for meeting the current conditions, and to further
identify measures that could be implemented as part of an on-going adaptation strategy.
Ecological constraints of assisted migration
Genetic and physiological factors
Temperature and moisture are major environmental factors affected by climate change, but ecological
niches are multi-dimensional, and trees must also adapt to other environmental variables. A major
genetic factor impacting the success of population migration is adaptation to new site conditions; trees
possessing alleles suitable for a particular temperature regime may be less well adapted to the local
abiotic and biotic conditions associated with a new site (Aitken et al. 2008; O’Neill et al. 2008).
Although potentially able to survive, normal growth could be impeded or altered to such an extent that
productivity would be affected and the tree would be considered “maladapted” for the site (O’Neill et
al. 2008). With increasing distance from the native (or historical) range of a plant species, differences
in soil pH, nutrient levels, texture, drainage, moisture capacity, and mineral composition could also
greatly affect the successful establishment and/or growth of tree species, and competition with other
As ranges extend to higher latitudes, summer day-lengths increase, winter day-lengths decrease, and
daily light availability near the vernal and autumnal equinoxes changes more rapidly. These factors
combine to alter the length of the growing season, the rate of plant growth, and the timing of bud-break
and bud-set (Downs and Borthwick 1956). Temperature affects many physiological processes relating
to germination, growth, reproduction and survival (Morgenstern 1996). Trees are adapted to complete
all their growth processes within a particular length of growing season. Spring bud burst is particularly
influenced by temperature (e.g., Badeck et al. 2003), whereby northern or high-elevation provenances
require fewer degree-days than southern or low-elevation provenances for bud-flush (Worrall 1983).
Trees moved northward from southern provenances may therefore be less likely to be impacted by late
spring frosts; however there is a risk that they will be affected by late summer or early autumn frost if
shoots and buds are not sufficiently hardened.
If climate change scenarios proceed as expected, climate zones will continue to shift for decades after
translocation, with implications for the timing and duration of assisted migration efforts (Gray et al.
2011; Pedlar et al. in this issue). In some cases this will mean that the planning horizon will be
relatively brief (Gray et al. 2011). Moreover, assumptions that variability in climate extremes will be
proportional to climate means are by no means certain (Schneider 2004); increases in weather extremes
may have significant impacts on ecosystems, including risks for both bud-flush and bud-set. For
example, it has been noted in Arctic tree-line studies that minimum winter temperatures and frosts may
have more impact on survival than summer air temperatures (Bonan and Sirois 1992).
Tree-associated species
Optimal tree health and productivity is dependent on a range of associated species, and the movement
of these species is a complicating factor in assisted migration projects. Within the native range of a tree
species, insects and microbes are usually well-adapted to provide key ecological services. Soil-
building, decomposition, nutrient-cycling, pathogen suppression, pollination, and other ecosystem
services provided by this community are all attuned to the composition of the native forest community,
extending as far as the genotypic level within tree species (e.g. Finzi et al. 1998; Ste-Marie and Houle
2006). As with physical soil properties, different soil communities may constrain the success and
establishment of tree species outside their native range (reviewed by Lafleur et al. 2010); globally,
there are still knowledge gaps concerning the functional influences of soil organisms (Tiedje et al.
Ectomycorrhizal (EM) mushrooms are a prime example of the importance of the soil community.
These associates of conifers are essential tree symbionts, acquiring water and nutrients and protecting
roots from various stresses. Sometimes cosmopolitan and sometimes with specialized niches, these
fungi can disseminate via spores or spread through underground growth. Some EM mushrooms with
expanding ranges are estimated to migrate less than 10 km y-1 (Pringle et al. 2009), roughly in line with
the 2 km y-1 rate estimated for invasive plants (Williamson et al. 2005); this may imply that the
migration capacity of most EM fungi is roughly equivalent to that of their tree hosts. Even hypogeous
(underground) symbionts important to wildlife, such as truffles (Pyare and Longland 2002), may have
some capacity to respond to climate change by undergoing fairly rapid range shifts (Büntgen et al.
2011). The history of long-distance tree migrations, however, has shown that there is often a need for
the right mycorrhizal species to be present when trees are translocated to distant habitat; trees lacking
appropriate mycorrhizal partners may show poor productivity (Björkman 1970).
Pests and pathogens
Assisted migration may involve exposure and adaptation to new pest relationships, or alteration of
extant relationships. Soil pathogens, for example, are one of the major determinants of biodiversity and
natural community structure along large geographic gradients (Reynolds et al. 2003). Among soil
pathogens, Armillaria solidipes (formerly known as A. ostoyae) is an aggressive root-rot fungus that
primarily attacks Douglas-fir (Pseudotsuga menziesii) trees on mesic sites in the interior of British
Columbia (B.C.); it is not reported in B.C. north of the latitude of Prince George. Northward migration
of Douglas-fir could result in temporary release from the influence of A. solidipes and greater exposure
to Armillaria sinapina, a more opportunistic root-rot pathogen with different impacts on stand
structure. Novel pest interactions and adaptations may also occur as a result of overlapping or
intersecting ranges that were previously distinct. In North America, for example, loblolly pine (Pinus
taeda) is thought to have originated from two Pleistocene refugia, one in south-eastern Texas and
northern Mexico, the other in southern Florida. The species has expanded its range northward, but
western populations are distinct in their resistance to fusiform rust and southern pine beetle, possibly
through introgression with long-leaf pines (Pinus palustris). As forest managers continue to mix seed
sources for plantations and as long-leaf pines have expanded northward, there are now new
opportunities for the spread of adaptive traits throughout the range of both pine species (Schmidtling et
al. 1999). In shorter time frames, host–pest interactions may be impacted by rapid range expansions,
where migration of the pest species is not synchronous with that of the host. Subsequent changes in
pest relationships may have positive or negative implications (Dukes et al. 2009). For example, impacts
from pests sensitive to overwintering conditions may be exacerbated or diminished as average and
extreme winter temperatures continue to change within the new optimal range. Projected increases in
extreme weather events (IPCC 2007) may have unexpected effects on pest populations, potentially
resulting in outbreaks (or, alternatively, pest population collapse) despite only moderate changes in
average temperatures.
AM efforts may encounter situations where seedlings will be exposed to new interactions (or levels of
interaction) with herbivores endemic to the recipient ecosystem. Along tree-lines, it has been noted
that grazing pressure from deer and other animals may be a very important factor in determining actual
versus potential migration rates (Grace et al. 2002). In studies of establishment of lodgepole pine
(Pinus contorta) at the northern portion of its range, Johnstone and Chapin (2003) noted very high
levels of mortality (50%) from snowshoe hare herbivory, although herbivory in successive years was
expected to decline as trees matured. Assisted migration efforts may also face situations where
herbivores from the “donor” ecosystem migrate with shifting climates well in advance of natural plant
migration; while we found no particular reports of this phenomenon in the literature, it seems
reasonable that migrating herbivore populations could also present challenges for AM in the “recipient”
ecosystem. In such a case, increasing populations of migrated seedlings might alleviate grazing
pressures; on the other hand, migrating herbivores may prefer grazing on the migrated seedlings,
especially where other plants are sub-optimal. The detailed possibilities are probably beyond the scope
of this report, but it is worth noting that there could be implications for seedling establishment and
wildlife management.
Competing vegetation
Exposure to competing vegetation will almost certainly affect successful seedling establishment in
assisted migration efforts. Competing vegetation often thrives when forests are disturbed (Wagner
2005). Calamagrostis canadensis, for example, is a native North American grass with a nominal
presence in intact boreal forests, but it is readily dispersed, and it becomes dominant in disturbed
situations (harvested areas, roadsides, etc). These properties effectively cause it to become an
“invasive” suppressor in reforestation efforts (Macey et al. 2001). If assisted migration efforts occur in
a disturbed interface where a formerly dominant tree species is receding, the “invasive” properties of
some native plants may present a competitive challenge.
Interactions and surprises
Changes in site conditions may lead to interactive effects from the previously mentioned constraints.
Constraints that increase plant stress may also increase disease risks, for example, it has been reported
that various site-related stresses may trigger endophytic fungi in trees to become pathogenic (Dorworth
and Callan 1996). On the other hand, interactions between tree genotypes and changes in day-length,
temperature extremes, and microclimate could disrupt the niches of certain pests, for example those
that synchronize with bud phenology (Hulme 1995; Volney and Fleming 2007) or those that are
sensitive to overwintering conditions (Winder et al. 2010). Occupation of an extended range may occur
within the context of shifts in community structure or composition. These alterations may be deliberate
(e.g., planting according to commercial considerations) or unintentional (e.g. shifts in dominance due to
pest impacts and tree mortality). In cases where previously distinct ranges of other species start to
intersect or overlap due to climate shifts, or in cases where there is some maladaptation to site
characteristics, a confluence of interactive impacts may be unavoidable.
Even where the above-ground portion of a migrated tree appears to grow rapidly, constraints may
interact over time to limit success. Maladapted roots, for instance, may cause trees to be more
susceptible to wind-throw, or disease may become prevalent as the stand matures. Seed germination
may be disrupted by new edaphic factors, or herbivores may impact natural recruitment of seedlings.
This is just a partial list of possible issues; new and unexpected ecological constraints will likely
emerge as assisted migration projects are implemented and monitored.
Ecological constraints for assisted migration: The Canadian context
If climate continues to shift during the century, AM efforts would potentially have long-term
implications beyond the scope of single timber rotations or current forest boundaries. This is illustrated
in Figure 1, where AM in forestry is portrayed not as a “one-shot deal”, but as a continuing effort to
monitor success and to keep pace with changing conditions. While the increased scope of impacts (in
space and time) may be less concerning for local forest managers, a collective awareness of the issue
will ultimately be important if AM efforts are implemented.
In the Canadian context and in the longer term, there are wider ecological implications for AM that
stem from the geographic attributes of the country. For example, Canada occupies latitudes that are at
the extreme limits for northward expansion of ranges in North America. In this region, the AM strategy
may therefore become ultimately impossible for some northern genotypes; under such circumstances,
seed banks or similar strategies might become necessary to conserve genetic resources. Even before
reaching the shoreline of the Arctic Ocean, northerly migrating species would ultimately encounter
permafrost and tundra landscapes. These northerly regions often have under-developed soils (Tedrow
and Cantlon 1969), and experience high winds (Wilson 1959), and temperature extremes (Bliss, 1962)
currently inhospitable to tree species; some of these conditions are related to climate and may abate
with increasing temperatures. Although the lack of competing trees north of tree-lines essentially
provides an empty niche for the expansion of forest areas (Grace et al. 2002), moss or lichen cover,
nutrient-poor organic soils, and competing vegetation may limit seed germination and seedling
establishment (Hobbie and Chapin 1998). Although some warmed permafrost areas might eventually
attain nominal nutrient relations and facilitate the northward spread of forests (Lafleur et al. 2010),
warming in the shorter term may also produce water-logged, infertile soils inhospitable to tree
seedlings (Anderson 1991). Concomitantly, colder soil temperatures may be challenging to the growth
of some plant species as ranges approach the Arctic Circle (Hobbie and Chapin 1998); this may change
with increasing temperatures.
The Canadian landmass also narrows toward the north, spanning roughly 45o of longitude in the area
between 49o and 75o N latitude, amounting to a width of about 4,800 km across the southern border
region, but only about 1,300 km across the islands of the northern Canadian Arctic. Some geographic
barriers to lateral migration (mountain ranges, agricultural areas) diminish and disappear to the north,
but Hudson Bay emerges as a central geographic barrier. To the west, Alaska has space for potential
range extensions northward to the Arctic shore; range expansion efforts may therefore become
northwesterly in this area, although mountains would challenge expansion into the tundra of western
Alaska. East of Canada’s Arctic territories, coastal areas of Greenland would be a small additional area
for potential range expansion and conservation of species or genotypes with further warming; since
1953, experimental plantations of Siberian larch (Larix sibirica), white spruce (Picea glauca), hybrid
white spruce, Scots pine (Pinus sylvestris), subalpine fir (Abies lasiocarpa), Engelmann spruce (Picea
engelmanii), Norway spruce (Picea abies), and Swiss stone pine (Picea cembra) have already been
established in a few protected areas of southern Greenland (Ødum 2003).
Another geographic feature of Canada with long-term implications for AM is the east–west orientation
of the country, which does not align with the north–south distribution of many of North America’s tree
species. This circumstance is complicated by the existence of major geographic barriers (e.g., the Great
Lakes, the St. Lawrence River, and the Toronto metropolitan area) along or near the USA–Canada
border. Tree species partly or mainly distributed in the lower continental United States may ultimately
see their optimal ranges extend into Canada, in whole or in part (e.g., Gibson et al. 2009; McLane and
Aitken 2011). Other tree species of western Canada, e.g. lodgepole pine, may ultimately migrate into
Alaska (Alden 1988; Johnstone and Chapin 2003). Successful assisted migration efforts for North
American forest species can therefore be viewed as an international enterprise, where the management
policies and efforts in different jurisdictions must mesh with the existing biological imperatives.
A final geographic consideration is Canada’s size. As the second largest country in the world, Canada’s
landscapes offer a variety of forest types and sizes, from large expanses like the boreal forest, to the
forests of smaller montane or maritime ecozones. There are many significant differences between
forested areas of eastern and western Canada, as evidenced by the relief of the terrain or in the relative
size of biogeoclimatic zones. The boreal forest is a key feature that spans northern Canada. Its southern
boundaries vary by region: in the east, it is bounded by primarily deciduous forests with over 50 tree
species; to the west, it is bounded by aspen parklands (croplands interspersed with grasslands and
trees), and further to the west, coniferous forests. This situation leads to different regional
vulnerabilities in the context of assisted migration. For example, forests in the west may be more
sensitive to the impacts of climate change on moisture regimes or have different fire regimes (Hogg
and Bernier 2005); for assisted migration, these differences may translate to an emphasis on different
adaptive traits or to differences in the quality and scale of establishment opportunities (e.g., burned
forests). For Canada, assisted migration efforts will deal with a “range of ranges”, from relatively
small, to those that span the continent. The larger ranges may have particular challenges, in terms of the
scope of the effort needed to maintain optimal productivity. For example, whereas environmental
selection may act to define the fundamental niche of a tree species, density-dependent selection may
cause the actual or realized niche of various genotypes to be relatively localized, narrow, or discrete.
This implies that successful assisted migration would involve extensive redistribution of genotypes
across the landscape (Rehfeldt et al. 1999). Approaches to assisted migration in Canada will clearly
differ from region to region and will require that geographic considerations be taken into account.
Mitigation of ecological constraints at different scales of assisted migration
Given the large translocation distances involved, assisted long-distance migration approaches may face
substantial ecological constraints. A particular issue in the deliberate movement of species over long
distances is the loss of tree-associated species. In ex-situ plantations of Douglas-fir and ponderosa pine
(Pinus ponderosa) in Patagonia, the passive introduction of EM fungi has contributed to establishment
of the trees, but the resulting fungal communities are less diverse than the same communities in native
forests (Barroetaveña et al. 2007); other North American tree species have been less successful at
establishing in this region (Simberloff et al. 2002). It is possible to use more active methods to
deliberately inoculate nursery-produced tree seedlings with a variety of targeted mycorrhizal fungi (e.g.
Trappe 1977; Molina 1979; Kottke et al. 1987); these methods may improve both the productivity and
health of species transported over longer distances. There are several strategies that could be employed
to either sustain the in situ diversity of cosmopolitan mycorrhizal fungi or establish non-endemic
symbionts, both with the aim of ensuring tree establishment and long-term stability and productivity of
the migrated tree species. EM fungi have been shown to require “reservoirs” or “refugia” of intact host
habitat to maintain their diversity. It has been shown, for example, that maintaining “nurse shrubs” on
sites may maintain these species in situ (Kitzberger et al. 2000). Retention of living trees can also
provide refugia for mycorrhizal fungi; the number and size of retention patches is thought to be an
important consideration for preservation of mycorrhizal diversity, with patch sizes of only 10m or so
providing significant benefit in some cases (Outerbridge and Trofymow 2004, Cline et al. 2005, Luoma
et al. 2006, Jones et al. 2008). Although the impacts of mycorrhizal diversity on productivity depend
on site characteristics (Jonsson et al. 2001), the distribution and size of migrated forest patches will
likely be a consideration when establishing trees in new types of habitat.
Understory plant species associating with tree species may also have significant influences on
productivity, nutrient-cycling, and ecosystem stability (Hutchison et al. 1999; Nilsson and Wardle
2005), and therefore could be a consideration for co-migration with trees to aid in successful
establishment when they improve these forest attributes. Environmental restoration studies have
explored the in situ seeding and transplantation of understory plants (e.g., Rayfield et al. 2005; Mottl et
al. 2006) or even intact portions of forest floor (Anonymous 2010). In the overstory, tree retention is
thought to be important in sustaining epiphytic lichens and mosses associated with old-growth, where
such organisms possess limited dispersal capabilities (Muir et al. 2006). If one of the goals of assisted
range extension or assisted long-distance migration is to maintain diversity of species, retaining mature
trees on the landscape will be a consideration as establishment proceeds in the longer term. All of this
being said, complete replication of forest communities in distant (ex situ) ecosystems would seem to be
an unrealistic goal, irrespective of how well such communities may support or depend upon a focal
species in its current environment.
Assisted range expansion should present fewer ecological constraints, as the recipient ecosystems are
just beyond the species’ range limits. However, adjacent ecosystems can be substantially different, and
species may still experience significant ecological challenges in establishment. In forest management,
one might imagine a wide range of methods that could be employed to “set the stage” for the arrival of
tree species before actually planting them (Millar et al. 2007). Shade-tolerant species, for example,
would benefit from the presence of an overstory (Gradowski et al. 2008); shade-intolerant species
would conversely benefit from relatively open canopies near the advancing edge of the range (Leithead
et al. 2010). Some opportunistic tree species may readily colonize disturbed sites, whereas others may
require relatively undisturbed conditions. Managing habitat near disjunctive refugia or “pioneer” stands
of the advancing species might be a particular consideration for this strategy. Understanding barriers to
natural migrations may provide insights for mitigating constraints on assisted range extension or
assisted long-distance migration. For certain Canadian tree species, natural range expansion is already
underway. Leithead et al. (2010) have characterized the northward migration of species such as red oak
(Quercus rubra) and white pine (Pinus strobus) in Ontario, observing that tree-fall gaps along the
temperate-boreal ecotonal boundary are permitting the establishment of northward migration for these
species, with such establishment occurring more often in larger, older gaps. Along Arctic tree lines,
there is evidence that shrub species are undergoing northward migration in some areas, although tree
migration can be constrained by a variety of factors in other areas (Grace et al. 2002). Observations of
existing migration processes can aid in the planning of assisted range expansion projects, maximizing
success and reducing uncertainty in ongoing assisted migration programs.
For assisted population migration, ecological constraints are likely to be much less than those involved
in range expansion or long-distance migration efforts, however, the selection of appropriate genotypes
for establishment can be a complex process when genotypes adapted to future climates may not be
optimally adapted to recipient site conditions. For example, it has been suggested that regional
provenances of Norway spruce planted in some areas of the Black Forest of Germany were less productive than provenances with more distant
origins (Hosius et al. 2006). As climatic tolerances are rarely uniform across a species’ range, source populations must be chosen very carefully
(McLachlan et al. 2007). European sources mention “enrichment planting” and “diversification of tree species mixtures and management” to ensure
resiliency and to counter climate change (European Forest Institute et al. 2008); these strategies are also relevant considerations for North American
approaches. Careful management is needed across all assisted migration projects to ensure that a variety of seed sources are used, thus ensuring genetic
diversity in introduced populations. Where site-specific factors become limiting, mitigation options could include
the use of breeding to combine adaptive traits and climate tolerances better suited for the extended
range; there is extensive knowledge available in this area, especially for commercial tree species.
Parent species or genotypes may hybridize to produce plants that are much better adapted to extreme
environments than either parental type (Rieseberg et al. 2003). Although the impacts of climate change
on particular pathosystems may be uncertain (Sturrock et al. 2011), hybridization efforts could also
improve resistance to some foliar and soil pathogens throughout the native range of the species.
Information on the edaphic adaptation of tree genotypes is also limited at this point, but an example
would be the adaptation of spruce species in the East Kootenays to calcareous soils (Xie et al. 1998).
Some estimate that up to 25% of plant species may occasionally breed with another species (Carroll
2010); where previously separate species ranges begin to overlap with climate change, new
opportunities for spontaneous hybridization and improvement of fitness may actually occur – a process
which could be enhanced by introduction of new genotypes and species.
As pests and hosts expand their ranges, there are implications for conserving genetic traits pertaining to
host resistance. Pest resistance in trees may be heterogeneous throughout the range of the tree species
and may not always correlate with shifting climate envelopes (Garrett et al. 2006). At the range limit of
a tree species, pests might therefore encounter “naïve” (unadapted) or susceptible genotypes of a
species. In performing assisted range expansion and assisted population migration, monitoring of
alleles related to pest resistance is therefore desirable, and could conceivably lead to some form of pest-
oriented assisted population migration to augment the diversity of populations as the new range is
established. Although the ecological constraints experienced by candidate species in assisted migration
efforts are likely to be substantial, more attention has been paid to the possible ecological consequences
of human-assisted movement of species into entirely new ecosystems; this subject is explored in greater
detail below.
Ecological consequences of assisted migration
Invasion risk of introduced species
When candidate species for assisted migration efforts are translocated to recipient ecosystems, the new
environment may be so suitable (e.g. free of pests and competitors) that the “immigrant” species
becomes invasive or pernicious. Invasive species cause billions of dollars in economic damage
(Pimentel et al. 2005); Diamond (2005) lists invasive species as one factor among many that can
potentially contribute to the failure of societies. From a more ecological perspective, invasive or
pernicious species can cause a range of impacts on productivity, biodiversity and ecosystem services
(Gordon 1998). The establishment of large (ca. 300 km2) commercial Eucalyptus spp. plantations in
Portugal is a case in point. Although these established forests have provided economic benefits in
terms of increased pulp production and ecological benefits such as increased litter production, these
man-made ecosystems have displaced the native ecosystem, resulting in a number of deleterious
impacts, including a reduction of wildlife habitat, increased soil hydrophobicity, and increased nutrient
losses (Kardell et al. 1986; Doerr et al. 1998; Thomas et al. 2000; Anonymous 2007).
Although it is sometimes difficult to predict invasive behavior, invasive species are reported to exploit
some common parameters and attributes, such as: relatively long translocation distances (Mueller and
Hellmann 2008); increased disturbance (Elton 1958); relatively high genetic diversity and frequency of
introductions; ability to shed deleterious alleles or otherwise adapt or evolve; ability to be asexual or
self-fertile; sustained introduction; prolific reproduction potential (Frankham 2005); broad or
“generalist” niches and spatial heterogeneity; and interactions with other species (Hastings et al. 2005).
What is the likelihood that trees translocated by assisted migration will display any of these invasive
attributes? Although climate envelopes should be relatively favorable for growth and establishment in
typical scenarios for assisted migration, constraints mentioned in the previous section could very well
limit habitat suitability, thus mitigating invasiveness; invasion risk assessments will likely have to be
conducted on a case-by-case basis, taking all available ecological information into account.
Globally, issues concerning invasive tree species have focused on intercontinental movements, e.g.,
broad-leaved paperbark (Melaleuca quinquenervia) in the Florida Everglades (Turner et al. 1998);
Pinus spp., Eucalyptus spp., and others in South Africa (Richardson 1998; Le Maitre et al. 2002);
lodgepole pine in New Zealand (Ledgard 2001); and Scots pine in Nova Scotia and Ontario (Catling
and Carbyn 2005). On the other hand, some intercontinental introductions have resulted in
“naturalized” populations with less aggressive behavior, e.g. Norway spruce in the forests of eastern
North America (e.g. Stover and Marks 1998; Hunter and Mattice 2002; Pennsylvania Department of
Conservation and Natural Resources 2011); sycamore in England (Peterken 2001); or lodgepole pine in
Scandinavia (Knight et al. 2001). In some cases, the “invasive” behaviour can be relatively subtle or
localized, as in the tendency for Douglas-fir to occupy and shade-out previously unforested rocky
slopes in Europe (Klingenstein and Diwani 2005), or in the potential for Siberian larch to dominate
some Alaskan sites despite lower densities across the overall forest landscape (Alden 2006). In
Patagonia, some introduced species initially thought to be invasive now appear to be less aggressive
(Simberloff et al. 2002). A similar effect has been noted in Britain, where trees introduced over the last
400 years are changing genetically and assimilating into forest communities; the long-term ecological
impacts of these species are not yet fully manifest (Peterken 2001).
Regarding the intra-continental migration of tree species, it is difficult to find examples of assisted
long-distance migration resulting in invasive behavior. There are some reports of “invasive” behavior
in species native to North America, particularly in areas experiencing changes in land management. For
example, Douglas-fir is reported to become invasive in oak savannas as a result of long-term fire
suppression (Devine and Harrington 2007). Another example, resulting from anthropogenic and natural
causes, would be the expansion of American beech vs. the decline of sugar maple in some hardwood
forests of Québec (Messier et al., 2011). In other cases, the “invasiveness” of native species within
North America, for example eastern hemlock in sugar maple and basswood forests of upper Michigan,
may relate to localized patterns of natural migration responding to changes in climatic conditions
(Davis et al., 1998). Overall, very few species become invasive when introduced into a novel
environment (Mueller and Hellman 2008); for forestry, the greater risk in assisted migration efforts
may be the unintentional introduction or exacerbation of forest pests and pathogens, as discussed
Invasion risk of pests and pathogens of introduced species
When plants are moved, their insect, pathogenic, and weedy pests often move with them, either with
the host, or migrating via various dispersal mechanisms. Intercontinental introductions of forest pests
can result in severe disturbance and economic damage in forest ecosystems (Liebhold et al. 1995). A
well-known example would be chestnut blight, which infected the dominant American chestnut
(Castanea dentata) across the forests of eastern North America (Winder and Shamoun 2006), resulting
in serious economic impacts (Freinkel 2007) and large-scale ecological change. When this “foundation
species” lost dominance, impacts included changes in nutrient cycling, soil processes, hydrological
cycles, and other deleterious effects on macro-invertebrates, fish, and wildlife, as well as widespread
invasion by other riparian trees and shrubs (Ellison et al. 2005).
Intra-continental translocations can also result in episodes of invasive disturbance and damage to
ecosystems. White pine blister rust, for example, was originally introduced to the forests of eastern
North America from origins in Asia, and was subsequently introduced to western North America
despite quarantine measures; some of these introductions were via stock originating from eastern North
America (Hunt 2003). This introduction resulted in the decline of western white pine, which has in turn
been linked to a decline in populations of grizzly bear and Clark’s nutcracker, both of which depend on
pine seeds for food (Tomback and Kendall 2001).
Northward movement of tree pathogens can also be a concern. The “Lake States” strain of
Scleroderrus canker, for example, is extant throughout the eastern forests of Canada and the USA, and
has been found as far west as Alberta (Skilling et al. 1986), but causes little tree mortality. The
“European” strain, however, causes serious damage in several pine species in northern New York and
Vermont; this has provoked regulatory action to restrict movement of the strain from this area (Skilling
et al. 1986). Regulatory action has also been taken to limit the northward spread of Sudden Oak Death,
a disease caused by an alien fungal pathogen with a number of alternative hosts, which is currently
expanding its range in the forests of western North America (Rizzo and Garbelotto 2003). Among alien
insect pests, Asian gypsy moth is a prominent species expanding its range in North America; it is
continually targeted by regulators for prevention of further intra-continental and intercontinental
introductions (Liebhold et al. 1992).
Regarding the intra-continental movement of forest pests native to North America, we did not locate
any reports of large-scale invasive introductions within North America, suggesting that assisted
population migration, assisted range expansion, and intra-continental modes of assisted long-distance
migration may be relatively low-risk approaches. However, forest insect pests are a potential source of
novel intra-continental interactions under a changing climate, given that insect pests are mobile, have
short generation times, and almost invariably possess a capacity for migration that far outstrips that of
their tree hosts. Moreover, they have a relatively high capacity for adapting to new hosts when their
ranges expand (Lieutier 2008). In North America, the mountain pine beetle (Dendroctonus ponderosae)
could be considered the “poster child” for these issues (Carroll et al. 2004). The expanding range of the
mountain pine beetle now brings this pest species into contact with new host populations in the boreal
forests of Alberta, including exposure to a new host species, jack pine (Pinus banksiana; Cullingham et
al. 2011). In principle, forest management methods currently used to limit beetle outbreaks (Carroll
2007) might also be applied to novel host situations; for example, pure monocultures of pine in a single
age class could be avoided. However, where there is a lack of adaptation in tree hosts, or where
planting alternative tree species is not an option, outcomes become much less certain. Plant pathogens
may also present adaptive risks when they migrate into new forests. As with their tree hosts, introduced
tree pathogens also have the potential to hybridize, thereby increasing their fitness (Brasier 2001). In
evaluating the risks and benefits of any assisted migration effort, the movement of pest species within
the new range, and possible intersection with the ranges of other host and pest species, are essential
Invasion in a forestry context
On the surface, successful assisted migration efforts would bear some resemblance to “invasive”
behaviour as non-native species become established and potentially dominant in recipient ecosystems.
Where climate change threatens ecosystem stability and causes the decline of dominant species,
phenomena that resemble “invasiveness” at a superficial level might actually be considered a measure
of success for assisted migration. In declining ecosystems, field studies have observed that introduced
plants may behave more like “passengers” than “drivers”; in other words, invasive phenomena
sometimes reflect opportunism after ecosystem disturbance, rather than active disturbance by the
invader (MacDougall and Turkington 2005). Even where dominant species or ecosystems are not in
decline, the degradation of ecosystem productivity and diversity by invasive trees is only one
possibility, mainly pertaining to cases where understory vegetation is suppressed or where there are
deleterious effects on soils. In other cases, invasive plants may actually benefit overall productivity and
diversity; the introduction of new dominant species may sustain and improve ecosystem function, albeit
in a novel configuration. Invasive plants have even been called “ecosystem engineers”, whose precise
role varies depending on the ecosystem and species involved (Crooks 2002).
Overall, a minority of species become invasive when introduced into a novel ecosystem. For example,
Richardson and Rejmánek (2004) list 80 conifer taxa (13% of conifer species) naturalized in forests
around the world, with another 36 species (6% of all conifer species) considered to be invasive. The
invasion of associated pests and pathogens may actually be the greater risk in assisted migration
approaches. On the other hand, despite the fact that pathogens are often treated as a direct negative
influence on forest productivity or diversity, forests in general are resilient to their impacts, which often
result in increased plant diversity (Winder and Shamoun 2006; Dukes et al. 2009). In the context of
shifts in tree dominance caused by assisted migration, it may be useful to view some tree pests as
agents that help define or diversify community structure.
Mitigation of ecological consequences at different scales of assisted migration
Assisted long-distance migration is likely to pose the greatest invasion risk to recipient ecosystems, as
it involves the introduction of alien species to novel habitats. Intercontinental invasive introductions are
much more common than intra-continental introductions, despite a higher frequency of intra-
continental human movement (Mueller and Hellman 2008). These results suggest that assisted range
expansion efforts may carry lower risk, although there is still the possibility of invasion following
short-distance and regional-scale translocations (Mueller and Hellman 2008; Davidson and Simkanin
Assisted population migration has the lowest risk, but there can be uncertainty concerning the selection
of source ecotypes or populations (McLachlan et al. 2007). Maintaining optimal mixes of species, age
classes, and seral stages will also be important, as planting even-aged monocultures can result in
unintended impacts from pest organisms (insects and pathogens). An example of this can be found in
BC, where intensive planting has shifted dominance of the tree community to lodgepole pine in some
areas. In north-western BC, this shift in dominance is at least partly responsible for an epidemic of
Dothiostroma needle blight (Woods et al. 2005). Meanwhile, in areas of southern BC dominated by
lodgepole pine plantations, climate-related increases in a variety of diseases and insect pests are leading
to calls for more careful planning in selection and planting of species (Heineman et al. 2010).
There are established risk assessments for invasiveness that can be adapted for assisted migration
purposes, thus identifying high-risk projects before they are implemented (Vitt et al. 2010). In Figure 1,
this assessment is a decision-making step for the modes of AM with higher risk, portrayed by a key
question: “Do the benefits of translocation outweigh the biological, social, and economic costs and
constraints?” In contrast to smaller weed species, trees that become invasive following assisted
migration practices are also likely to be more easily identified and controlled through traditional
silvicultural methods (Ledgard 2001; Le Maitre et al. 2002); among the trees and shrubs that have been
introduced to Britain, it has been observed that extensive control programs have been able to eradicate
Rhododendron ponticum, despite its significant threat to ecosystems (Peterken 2001).
Assessments of species vulnerability with respect to changing climates are also underway (q.v.
Beardmore et al. in this issue). Where vulnerable or declining tree species and genotypes cannot be
successfully translocated in AM scenarios, ex situ conservation strategies may be the only alternative
option (Hoegh-Guldberg et al. 2008; Vitt et al. 2010). Vitt and colleagues (2010) advocate the
conservation of plant species’ gametes using carefully managed seed banks. Seed banks can allow
managers to delay decisions on assisted migration until more information is known, thus providing
“species insurance” (Vitt et al. 2010). Relative to intensive conservation efforts, preservation in seed
banks may be less expensive, and managers can prioritize the collection of species with low tolerance
to disturbance, high fidelity to habitat integrity, non-aggressive pioneer species (good for restoration
efforts), native species at their range limit, rare/threatened/endangered species, narrow endemics,
species that are highly conservative, and species with dispersal limitations (Vitt et al. 2010).
Certainly, assisted migration is not the only strategy for adaptation to climate change, and there other
forest practices that are potentially useful alternatives or adjuncts to this approach (Ogden and Innes
2007). Using various forest management practices to promote species and forest resiliency within
existing populations is a common option for coping with disturbances, for example, following fire or
insect and pathogen outbreak. Millar et al. (2007) propose that the capacity to maintain and improve
resiliency becomes more difficult and requires extensive intervention as climate changes progress over
time. Although maintaining resilience could also be considered a short-term option, it may be an
appropriate long-term strategy when an ecosystem appears to be insensitive to a changing climate, or
when climate change in a particular area is expected to be relatively modest. Where ranges begin to
extend into Canada, populations at the leading edge of a range may warrant special attention in
conservation efforts, because they may have genotypic or adaptive attributes important in further range
extension, irrespective of whether the extension is natural or assisted (Gibson et al. 2009). One could
also promote adaptation to the changing environment by assisting with natural ecosystem processes;
among myriad forest management options, this could include strategies such as facilitation of seed
dispersal and establishment (e.g., protecting vector species, controlling seed pathogens, employing
prescribed burning for fire-adapted species, etc.), altering community composition (e.g., controlling
populations of competing tree species or vegetation), or altering community structure (e.g., light and
temperature control through opening canopies, growing “nurse” crops for shading, etc.).
Preventing negative ecological consequences of assisted migration efforts will continue to be an active
area of research, however, the fear of invasion should not prevent implementation of assisted migration
programs. In fact, insight into mitigating these unexpected ecological consequences is most likely to
come from well-planned and carefully monitored assisted migration attempts.
Although there are significant knowledge gaps within our understanding of the ecological
considerations pertaining to assisted migration, this paper attempts to present the state of ecological
knowledge available to those making on-the-ground decisions about the implementation of human-
assisted translocation in response to climate change. This ecological information can be combined with
decision-making frameworks (Figure 1) to identify species- and site-specific research priorities, thus
closing these knowledge gaps on a case-by-case basis.
From a purely ecological perspective, assisted migration is a feasible option in many situations,
particularly where risks are minimized through shorter transfer distances and/or careful management.
This is especially true for commercial tree species, where established site preparation, planting and
monitoring methods are in place (q.v. Pedlar et al. in this issue). However, concerns about helping
natural migration are also inextricably interwoven with human values and questions of scale; over
millennial time-spans, we can expect forest ecosystems to re-establish optimal productivity and
stability without human assistance, just as these ecosystems colonized the far and isolated corners of
the planet prior to the current epoch (Ridley 1930). Unfortunately, ecosystems are already being
impacted by climate change and humans will not have millennia to consider their choices. Moreover, if
we accept humans as part of nature, it becomes vital to also address human values and concerns from
an ecological perspective. Taking all of that into consideration, assisted migration is an option that
remains on the table for ecological adaptation to rapid climate change. Although there will certainly be
caveats concerning its safe and effective implementation, assisted migration can contribute to optimal
productivity, stability, and condition of forest ecosystems in Canada, and throughout the biosphere.
We would like to thank the other members of the Assisted Migration team, who greatly aided the
development of this manuscript. Anna Dabros, Elaine Qualtiere, Catherine Ste-Marie, and Caroline
Simpson provided valuable comments on the manuscript, and Dale Simpson contributed important
material for the paper. We would like to especially thank Sylvie Gauthier for her excellent review and
for proposing a new and much-improved manuscript structure, and David Price for his thorough,
conscientious and constructive comments on our first and second drafts. We also thank Julie Piché for
assisting with illustration.
Literature cited
Aitken, S.N., S. Yeaman, J.A. Holliday, T. Wang, and S. Curtis-McLane. 2008. Adaptation, migration or extirpation: climate change
outcomes for tree populations. Evolutional Applications 1: 95-111.
Alden, J.N. 1988 Implications of research on lodgepole pine introduction in interior Alaska. U.S.D.A. Forest Service Research Paper
Alden, J.N. 2006. Field survey of growth and colonization of nonnative trees on mainland Alaska. U.S.D.A. Forest Service General
Technical Report PNW-GTR-664.
Anderson, J.M. 1991. The effects of climate change on decomposition processes in grassland and coniferous forests. Ecol. Appl. 1:326-
Anonymous. 2007. EU ‘put Portugal wildlife under threat’. BBC News, Accessed 20 Oct. 2010 at:
Anonymous. 2010. Regreening Greater Sudbury: five-year plan; Regreening Program 2011-2015. City of Greater Sudbury / Vegetation
Enhancement Technical Advisory Committee. 21 p.
Badeck, F-W., A. Bondeau, K. Böttcher, D. Doktor, D. Lucht, J. Schaber and S. Sitch. 2003. Responses of spring phenology to climate
change. New Phytologist 162: 295-309.
Barroetaveña, C., E. Cázares, and M. Rajchenberg. 2007. Ectomycorrhizal fungi associated with ponderosa pine and Douglas-fir: a
comparison of species richness in native western North American forests and Patagonian plantations from Argentina. Mycorrhiza 17:
Bliss, L.C. 1962. Adaptations of Arctic and alpine plants to environmental conditions. Arctic 15:117-144.
Björkman, E. 1970. Forest tree mycorrhiza: the conditions for its formation and the significance for tree growth and afforestation Plant
and Soil 32: 589-610,
DOI: 10.1007/BF01372897
Bonan, G.B. and L. Sirois. 1992. Air temperature, tree Growth, and the northern and southern range limits to Picea mariana Journal of
Vegetation Science 3: 495-506.
Brasier, C.M. 2001. Rapid evolution of introduced plant pathogens via interspecific hybridization. BioScience 51:123-133.
Büntgen, U., W. Tegel, S. Egli, U. Stobbe, L. Sproll, and N.C. Stenseth. 2011. Truffles and climate change. Frontiers in Ecology & the
Environment 9: 150-151.
Carroll, A.L. 2007. The Mountain Pine Beetle Dendroctonus ponderosae in western North America: potential for area-wide integrated
management. Pages 297-307 in M.J.B Vreysen, A.S. Robinson, and J. Hendrichs, editors. Area-Wide Control of Insect Pests From
Research to Field Implementation. Springer, New York, NY.
Carroll, A.L., S.W. Taylor, J. Régnière, and L. Safranyik. 2004. Effects of climate change on range expansion by the mountain pine beetle
in British Columbia. In: Shore, T.L., Brooks, J.E. and J.E. Stone, (Eds.), Mountain Pine Beetle Symposium: Challenges and
Solutions, October 30-31, 2003, Kelowna, British Columbia, Canada. Natural Resources Canada, Canadian Forest Service, Pacific
Forestry Centre, Victoria, British Columbia, Information Report BC-X-399. pp 223-232.
Carroll, S.B. 2010. Hybrids may thrive where parents fear to tread. New York Times, Sept. 13, 2010.
Catling, P.M. and S. Carbyn. 2005. Invasive Scots pine, Pinus sylvestris, replacing Corema, Corema conradii, heathland in the Annapolis
Valley, Nova Scotia. Canadian Field Naturalist. 119: 237-244.
Cline, E.T., J.F. Ammirati and R.L. Edmonds. 2005. Does proximity to mature trees influence ectomycorrhizal fungus communities of
Douglas-fir seedlings? New Phytologist 166: 993–1009.
Collingham, Y.C. and B. Huntley. 2000. Impacts of habitat fragmentation and patch size upon migration rates. Ecological Applications
10: 131–144.
Crooks, J.A. 2002. Characterizing ecosystem-level consequences of biological invasions: the role of ecosystem engineers. Oikos 97: 153-
Crumpacker, D.W., E.O. Box and E.D. Hardin. 2001. Implications of climatic warming for conservation of native trees and shrubs in
Florida. Conservation Biology 15:1008–1020.
Cullingham, C.I., J.E.K. Cooke, C.S. Davis, B.J. Cooke and D.W. Coltman. 2011. Mountain pine beetle host-range expansion threatens
the boreal forest. Molecular Ecology 20: 2157-2171.
Davidson, I. and C. Simkanin. 2008. Skeptical of assisted colonization. Science. 322: 1048-1049.
Davis, Margaret B., Randy R. Calcote, Shinya Sugita, and Hikaru Takahara. 1998. Patchy invasion and the origin of a hemlock-
hardwoods forest mosaic. Ecology 79: 2641–2659.
Devine, W.D. and C.A. Harrington. 2007. Release of Oregon white oak from overtopping Douglas-fir: effects on soil water and
microclimate. Northwest Science 81: 112-124.
Diamond, J. 2005. Collapse: How societies choose to fail or succeed. Penguin Group, New York. 575pp.
Doerr, S.H., R.A. Shakesby and R.P.D. Walsh. 1998. Spatial variability of soil hydrophobicity in fire-prone Eucalyptus and Pine Forests,
Portugal. Soil Science 163: 313-324.
Dorworth, C.E. and B.E. Callan. 1996. Manipulation of endophytic fungi to promote their utility as vegetation biocontrol agents. 1996.
Pp. 209-219 in: S.C. Redlin and L.M. Carris, eds. Endophytic fungi in grasses and woody plants: systematics, ecology, and
evolution, St. Paul, MN. APS Press, St. Paul, MN.
Downs, R.J. and H.A. Borthwick. 1956. Effects of photoperiod on growth of trees. Botanical Gazette 117: 310-326.
Dukes, J., J. Pontius, D. Orwig, J.R. Garnas, V.L. Rodgers, N. Brazee, B. Cooke, K.A. Theharides, E.E. Strange, R. Harrington, J.
Ehrenfeld, M.L. Gurevitch, K. Stinson R. Wick and M. Ayres. 2009. Responses of insect pests, pathogens, and invasive plant species
to climate change in the forests of northeastern North America: What can we predict? Canadian Journal of Forest Research. 39:
Ellison, A.M., M.S. Bank, B.D. Clinton, E.A. Colburn, K. Elliott, C.R. Ford, D.R. Foster, B.D. Kloeppel, J.D. Knoepp, G.M. Lovett, J.
Mohan, D.A. Orwig, N.L. Rodenhouse, W.V. Sobczak, K.A. Stinson, J.K. Stone, C.M. Swan, J. Thompson, B. Von Holle and J.
Webster. 2005. Loss of foundation species: consequences for the structure and dynamics of forested ecosystems. Frontiers in
Ecology and the Environment 3:479-486.
Elton, C.S. 1958. The ecology of invasions by animals and plants. Methuen & Co., London.
European Forest Institute, University of Natural Resources and Applied Life Sciences (Vienna) Institute of Silviculture and Institute of
Forest Entomology, Forest Pathology and Forest Protection, INRA - UMR Biodiversité Gènes et Communautés Equipe de
Génétique, and Italian Academy of Forest Sciences. 2008. Impacts of Climate Change on European Forests and Options for
Adaptation. Report to the European Commission Directorate-General for Agriculture and Rural Development, AGRI-2007-G4-06.
Finzi, A.C., N. van Breemen, N., and C.D. Canham. 1998. Canopy tree-soil interaction within temperate forests: species effect on soil
carbon and nitrogen, Ecological Application 8:440–446.
Frankham, R. 2005. Resolving the genetic paradox in invasive species. Heredity 94: 385.
Freinkel, S. 2007. American chestnut: the life, death, and rebirth of a perfect tree. Univ. of California Press, Los Angeles, CA. 284 pp.
Garrett, K.A., S.P. Dendy, E.E. Frank, M.N. Rouse, and S.E. Travers. 2006. Climate change effects on plant disease: Genomes to
ecosystems. Annual Review of Phytopathology 44: 489–509
Gibson, S.Y., R.C. Van Der Marel, and B.M. Starzomski. 2009. Climate change and conservation of leading-edge populations.
Conservation Biology 23: 1369-1373.
Goodale, C.L., M.J. Apps, R.A. Birdsey, C.B. Field, L.S. Heath, R.A. Houghton, J.C. Jenkins, G.H. Kohlmaier, W.A. Kurz, S. Liu, G-J.
Nabuurs, S. Nilsson and A.Z. Shvidenko. 2002. Forest carbon sinks in the norther hemisphere. Ecological Applications. 12: 891-899.
Gordon, D.R. 1998. Effects of invasive, non-indigenous plant species on ecosystem processes: Lessons from Florida. Ecological
Applications 8: 975-989.
Grace, J., F. Berninger, and L. Nagy. 2002. Impacts of climate change on the tree line. Annals of Botany 90: 537-544.
Gradowski, T., D. Sidders, T. Keddy, V.J. Lieffers, and M. Landhäusser. 2008. Effects of overstory retention and site preparation on
growth of planted white spruce seedlings in deciduous and coniferous dominated boreal plains mixedwoods. Forest Ecology and
Management 255: 3744-3749.
Gray, L.K., T. Gylander, M.S. Mbogga, P. Chen, and A. Hamann. 2011. Assisted migration to address climate change: recommendations
for aspen reforestation in western Canada. Ecological Applications 21: 1591–1603.
Hamann, A and T. Wang. 2006. Potential effects of climate change on ecosystem and tree species distribution in British Columbia.
Ecology. 87: 2773–2786.
Hastings, A., K. Cuddington, K.F. Davies, C.J. Dugaw, A.F. Elmendorf, S. Harrison, M. Holland, J. Lambrinos, U. Malvadkar, B.A.
Melbourne, K. Moore, C. Taylor and D. Thomson. The spatial spread of invasions: new developments in theory and evidence.
Ecology Letters 8: 91-101.
Heineman, J.L., D.L. Sachs, J. Mather and S.W. Simard. 2010. Investigating the influence of climate, site, location, and treatment factors
on damage to young lodgepole pine in southern British Columbia. Canadian Journal of Forest Research 40: 1109-1127.
Hobbie, S.E. and S.F. Chapin III. 1998. An experimental test of limits to tree establishment in Arctic tundra. Journal of Ecology 86:449–
Hoegh-Guldberg, O., L. Huges, S. McIntyre, D.B. Lindenmayer, C. Parmesan, H.P. Possingham and C.D. Thomas. 2008. Assisted
colonization and rapid climate change. Science 321: 345-346.
Hogg, E.H. and P.Y. Bernier. 2005. Climate change impacts on drought-prone forests in western Canada. Forestry Chronicle 81:675-682.
Hosius, B., L. Leinemann, M. Konnert and F. Bergmann. 2006. Genetic Aspects of Forestry in the Central Europe. European Journal of
Forest Research 125: 407-417, DOI: 10.1007/s10342-006-0136-4
Hulme, M. 1995. Resistance by translocated Sitka spruce to damage by Pissodes strobi (Coleoptera: Curculionidae) related to tree
phenology. Journal of Economic Entomology 88: 1525-1530.
Hunt, R. 2003. White pine blister rust. Recent Research Developments in Mycology 1:73-85.
Hunter, J.C. and J.A. Mattice. 2002. The spread of woody exotics into the forest of a northeastern landscape, 1938-1999. Journal of the
Torrey Botanical Society 129: 220-227
Hunter, M. L. Jr. 2007. Climate change and moving species: furthering the debate on assisted colonization. Conservation Biology. 21:
Huntley, B.1991. How plants respond to climate change: migration rates, individualism and the consequences for plant communities.
Annals of Botany 67 (suppl.): 15–22.
Hutchison, T.F., R.E.J. Boerner, L.R. Iverson, S. Sutherland and E.K. Sutherland. 1999. Landscape patterns of understory composition
and richness across a moisture and nitrogen mineralization gradient in Ohio (U.S.A.) Quercus forests. Plant Ecology 144: 177-189
Intergovernmental Panel on Climate Change. 2007, Climate Change 2007: The Physical Science Basis, Contribution of Working Group I
to the Fourth Assessment Report of the Intergovernmental Panel on Climate Change, edited by S. Solomon et al., Cambridge Univ.
Press, Cambridge,U. K.
Iverson, L.R. and A.M. Prasad. 1998. Predicting abundance of 80 tree species following climate change in the eastern United States.
Ecological Monographs. 68: 465–485.
Jandl, R., L. Vesterdal, M. Olsson, O. Bens, F. Badeck and J. Rock. 2007. Carbon sequestration and forest management. CAB Reviews:
Perspectives in Agriculture, Veterinary Science, Nutrition and Natural Resources. 2: 017.
Johnston, M., D, Price, S. L’Hirondelle, R. Fleming, and A. Ogden. 2010. Tree species vulnerability and adaptation to climate change:
Final technical report. SRC Publication No. 12416-1E10, Saskatchewan Research Council, Saskatoon, Saskatchewan.
Johnstone, J.L. and F.S. Chapin. 2003. Non-equilibrium succession dynamics indicate continued northern migration of lodgepole pine.
Global Change Biology 9:1401–1409.
Jones, M.D., B.D. Twieg, D. Durall and S.M. Berch. 2008. Location relative to a retention patch affects the ECM fungal community more
than patch size in the first season after timber harvesting on Vancouver Island, British Columbia. Forest Ecology and Management
255: 1342-1352.
Jonsson, L.M., M-C. Nilsson, D.A. Wardle and O. Zackrisson. 2001. Context dependent effects of ectomycorrhizal species richness on
tree seedling productivity. Oikos. 93: 353-364.
Kardell, L., E. Steen and A. Fabiao. 1986. Eucalyptus in Portugal – A threat or a promise? Ambio 15: 6-13.
Kitzberger, T., D.F. Steinaker and T.T. Veblen. 2000. Effects of climatic variability on facilitation of tree establishment in northern
Patagonia. Ecology 81:1914–1924.
Klingenstein, F. and T. Diwani. 2005. Invasive alien species from a nature conservation point of view in Germany. In: Identification of
risks and management of invasive alien species using the IPPC framework: Proceedings of the workshop on invasive alien species
and the International Plant Protection Convention, Braunschweig, Germany, 22–26 September 2003. 137-145 pp. IPPC Secretariat,
Rome, Italy.
Knight, D.H., W.L. Baker, O. Engelmark and C. Nilsson. 2001. A landscape perspective on the establishment of exotic tree plantations:
lodgepole pine (Pinus contorta) in Sweden. Forest Ecology and Management 141: 131-142.
Kottke, I., M. Guttenberger, R. Hampp and F. Oberwinkler. 1987. An in vitro method for establishing mycorrhizae on coniferous tree
seedlings. Trees 1: 191-194.
Kurz, W.A., G. Stinson, G.J. Rampley, C.C. Dymond, and E.T. Neilson. 2008. Risk of natural disturbances makes future contribution of
Canada’s forests to the global carbon cycle highly uncertain. Proceedings of the National Academy of Science 105: 1551-1555
Lafleur, B., D. Paré, A.D. Munson and Y. Bergeron. 2010. Response of northeastern North American forests to climate change: will soil
conditions constrain tree species migration? Environmental Reviews 18: 279-289.
Ledgard, N. 2001 The spread of lodgepole pine (Pinus contorta, Dougl.) in New Zealand. Forest Ecology and Management. 141: 43-57.
Le Maitre, D.C, B.W. van Wilgen, C.M. Gelderblom, C. Bailey, R.A. Chapman and J.A. Nel. 2002. Invasive alien trees and water
resources in South Africa: case studies of the costs and benefits of management. Forest Ecology and Management. 160: 143-159.
Leithead, M.D., M. Anand and L.C.R. Silva. 2010. Northward migrating trees establish in treefall gaps at the northern limit of the
temperate-boreal ecotone, Ontario, Canada. Oecologia 164: 1094-1106.
Liebhold, A.M., J.A. Halverson, and G.A. Elmes. 1992. Gypsy moth invasion in North America: A quantitative analysis. Journal of
Biogeography 19:513-520.
Liebhold, A.M., W.L. MacDonald, D. Begdahl, and V.C. Mastro. 1995. Invasion by exotic forest pests: a threat to forest ecosystems.
Forest Science 41: Supplement (Monograph 30).
Lieutier, F. 2008. Changing forest communities: Role of tree resistance to insects in insect invasions and tree introductions. In: T.D.
Paine (ed.) Invasive forest insects, introduced forest trees, and altered ecosystems: Ecological pest management of global forests in a
changing world. 15-52 pp. Springer, Dordrecht, Netherlands.
Loarie, S.R., P.B. Duffy, H. Hamilton, G.P. Asner, C.B. Field, and D.D. Ackerly. 2009. The velocity of climate change. Nature 462:
Luoma, D.L., C.A. Stockdale, R. Molina and J.L. Eberhart. 2006. The spatial influence of Pseudotsuga menziesii retention trees on
ectomycorrhiza diversity. Canadian Journal of Forest Research 36: 2561–2573.
MacDougall, A.S. and R. Turkington. 2005. Are invasive species the drivers or passengers of change in degraded ecosystems? Ecology
86: 42–55.
Macey, D.E. and R.S. Winder. 2001. Biological control and the management of Calamagrostis canadensis (bluejoint grass) Natural
Resources Canada, Can. For. Serv., Pacific Forestry Centre, Victoria, BC. Tech. Transfer Note 25. 6 p.
Mack, R.N. 2001. Motivations and consequences of the human dispersal of plants. Pp. 23-34 in: J.A. McNeely (ed.) The Great
Reshuffling: Human dimensions of invasive alien species. IUCN Publications Services Unit, Cambridge, UK. 242 pp.
Malcolm, J.R., A. Markham, R.P. Neilson, and M. Garacil. 2002. Estimated migration rates under scenarios of global climate change.
Journal of Biogeography. 29: 835–849.
McKenney, D.W., J.H. Pedlar, K. Lawrence, K. Campbell, and M.F. Hutchinson, M.F. 2007. Potential Impacts of Climate Change on the
Distribution of North American Trees. BioScience. 57: 939-948.
McLachlan, J.S., J.J. Hellmann and M.W. Schwartz. 2007. A framework for debate of assisted migration in an era of climate change.
Conservation Biology 21: 297–302.
McLane, S.C. and S.E. Aitken. 2011. Whitebark pine (Pinus albicaulis) assisted migration trial. P. 205 in: Keane, Robert E.; Tomback,
Diana F.; Murray, Michael P.; Smith, Cyndi M., eds. The future of high-elevation, five-needle white pines in Western North
America: Proceedings of the High Five Symposium. 28-30 June 2010; Missoula, MT. Proceedings RMRS-P-63. Fort Collins, CO:
U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station.
Messier, C., Bélanger, J., Brisson, M., Lechowicz, J., and Gravel, D. 2011. Comment on ‘‘Present-day expansion of American beech in
northeastern hardwood forests: Does soil base status matter?” Can. J. For. Res. 41: 649–653.
Millar, C.I., N.L. Stephenson, and S.L. Stephens. 2007. Climate change and forests of the future: managing in the face of uncertainty.
Ecological Applications. 17: 2145-2151.
Molina, R. 1979. Ectomycorrhizal inoculation of containerized Douglas-fir and lodgepole pine seedlings with six Isolates of Pisolithus
tinctorius. Forest Science 25:585-590.
Morgenstern, E.K. 1996. Geographic variation in forest trees: genetic basis and application of knowledge in silviculture. University of
British Columbia Press, Vancouver, BC. 209 p.
Mottl, L.M., C.M. Mabry, and D.R. Farrar. 2006. Seven-year survival of perennial herbaceous transplants in temperate woodland
restoration. Restoration Ecology. 14: 330–338.
Mueller, J.M., and J.J. Hellmann. 2008. An assessment of invasion risk from AM. Conservation Biology 22: 562–567.
Muir, P.S., T.R. Rambo, R.W. Kimmerer and D.B. Keon. 2006. Influence of overstory removal on growth of epiphytic mosses and
lichens in Western Oregon. Ecological Applications 16: 1207-1221.
Nilsson, M.-C. and D.A. Wardle. 2005. Understory vegetation as a forest ecosystem driver: evidence from the northern Swedish boreal
forest. Frontiers in Ecology and the Environment 3:421-428.
Ødum, S. 2003. Choice of conifer plant material for south west Greenland. Acta Horticulturae (ISHS) 615:273-279
Odum, E.P. 1974. The strategy of ecosystem development: an understanding of ecological succession provides a basis for resolving man’s
conflict with nature. Pp. 224-240 in P.E. Black and L.P. Herrington, eds. Readings in environmental impacts. MSS Information
Corporation, NY. 353 p.
Ogden, A.E. and J. Innes. 2007. Incorporating climate change adaptation considerations into forest management planning in the boreal
forest. International Forestry Review 9:713-733.
O’Neill, G., A. Hamann, and T. Wang. 2008. Accounting for population variation improves estimates of the impact of climate change on
species’ growth and distribution. Journal of Applied Ecology 45: 1040-1049.
Outerbridge, R.A. and J.A. Trofymow. 2004. Diversity of ectomycorrhizae on experimentally planted Douglas-fir seedlings in variable
retention forestry sites on southern Vancouver Island. Canadian Journal of Botany 82: 1671–1681.
Pan, Y., R.A. Birdsey, J. Fang, R. Houghton, P.E. Kauppi, W.A. Kurz, O.L. Phillips, A. Shvidenko, S.L. Lewis, J.G. Canadell, P. Ciais,
R.B. Jackson, S. Pacala, A.D. McGuire, S. Piao, A. Rautiainen, S. Sitch, and D. Hayes. 2011. A Large and persistent carbon sink in
the world’s forests. Science 333: 988-993.
Pakeman, R.J. 2001. Plant migration rates and seed dispersal mechanisms. Journal of Biogeography 28: 795-800.
Pennsylvania Department of Conservation and Natural Resources. 2011. Common trees of Pennsylvania. Accessed 10 March 2011 at:
Peterken, G.F. 2001. Ecological effects of introduced tree species in Britain. Forest Ecology and Management 141: 31-42.
Pimentel, D., R. Zuniga and D. Morrison. 2005. Update on the environmental and economic costs associated with alien-invasive species
in the United States. Ecological Economics 52: 273-288.
Pringle, A., R. Adams, H. Cross, and T. Bruns. 2009. The ectomycorrhizal fungus Amanita phalloides was introduced and is expanding
its range on the west coast of North America. Molecular Ecology 18: 817-833.
Pyare, S., and W.S. Longland. 2002. Interrelationships among northern flying squirrels, truffles, and microhabitat structure in Sierra
Nevada old-growth habitat. Canadian Journal of Forest Research 32: 1016-1024.
Rayfield, B., M. Anand, and S. Laurence. 2005. Assessing simple versus complex restoration strategies for industrially disturbed forests.
Restoration Ecology 13: 639-650.
Reed, D.D. and P.V. Desanker. 1992. Ecological implications of projected climate change scenarios in forest ecosystems in northern
Michigan, USA. International Journal of Biometeorology 2: 97-107.
Rehfeldt, G.E. and B.C. Jaquish. 2010. Ecological impacts and management strategies for western larch in the face of climate-change.
Mitigation and Adaptation Strategies for Global Change. 15: 283-306.
Rehfeldt G.E., C.C. Ying, D.L. Spittlehouse and D.A. Hamilton. 1999. Genetic responses to climate in Pinus contorta: niche breadth,
climate change, and reforestation. Ecological Monographs 69: 375-407.
Reynolds, H.L., A. Packer, J.D. Bever and K. Clay. 2003. Grassroots ecology: plant–microbe–soil interactions as drivers of plant
community structure and dynamics. Ecology 84: 2281–2291.
Richardson, D.M. 1998. Forestry trees as invasive aliens. Conservation Biology 12: 18-26.
Richardson, D.M., J.J. Hellmann, J. S. McLachlan, D.F. Sax, M.W. Schwartz, P. Gonzalez, E.J. Brennen, A. Camacho, T.L. Root, O.E.
Sala, S. H. Schneider, D. M. Ashe,J. R. Clark, R. Early, J.R. Etterson, E.D. Fielder, J.L. Gill, B.A. Minteer, S. Polasky, H.D. Safford,
A.R. Thompson and M. Vellend. 2009. Multidimensional evaluation of managed relocation. Proceedings of the National Academy
of Sciences 106: 9721-9724.
Richardson, D.M. and M. Rejmánek. 2004. Conifers as invasive aliens: a global survey and predictive framework. Diversity and
Distributions 10: 321-331.
Ridley, H.N. 1930. The dispersal of plants throughout the world. L. Reeve and Co., Ltd., Ashford, UK. 744 p.
Rieseberg, L.H., O. Raymond, D.M. Rosenthal, Z. Lai, K. Livingstone, T. Nakazato, J.L. Durphy, A.E. Schwarzbach, L.A. Donovan and
C. Lexer. 2003. Major ecological transitions in wild sunflowers facilitated by hybridization. Science 301: 1211-1216
Rizzo, D.M., and M. Garbelotto. 2003. Sudden oak death: endangering California and Oregon forest ecosystems. Frontiers in Ecology and
the Environment 1: 197–204.
Schmidtling, R.C., E. Carroll, and T. LaFarge. 1999. Allozyme diversity of selected and natural loblolly pine populations. Silvae Genetica
48: 35-45.
Schneider, S.H. 2004. Abrupt non-linear climate change, irreversibility and surprise. Global Environmental Change 14: 245–258.
Simberloff, D., M. Relva and M. Nuñez. 2002. Gringos en el bosque: introduced tree invasion in a native Nothofagus/Austrocedrus forest.
Biological Invasions 4: 35-53.
Skilling, D.D., B. Scheider, and D. Fasking. 1986. Biology and control of Sclerroderrus canker in North America. U.S.D.A. Forest
Service Research Paper NC-275.
Ste-Marie, C., and D. Houle. 2006. Forest floor gross and net nitrogen mineralization in three forest types in Quebec, Canada. Soil
Biology and Biochemistry 38: 2135-2143
Stover, M.E. and P.L. Marks. 1998. Successional vegetation on abandoned cultivated and pastured land in Tompkins County, New York.
Journal of the Torrey Botanical Society 125: 150-164.
Sturrock, R.N., S.J. Frankel, A.V. Brown, P.E. Hennon, J.T. Kliejunas, K.J. Lewis, J.J. Worrall and A.J. Woods. 2011. Climate change
and forest diseases. Plant Pathology 60: 133-149.
Tedrow, T.C.F. and J.E. Cantlon. 1969. Concepts of soil formation and classification in Arctic regions. Pp. 125-138 in: Nelson, J.G. and
Chambers, M.J. Process and method in Canadian geography: vegetation, soils and wildlife. Methuen Publications, Toronto, Canada.
375 p.
Thomas, A.D., R.P.D. Walsh and R.A. Shakesby. 2000. Solutes in overland flow following fire in Eucalyptus and pine forests, northern
Portugal. Hydrological Processes 14: 971-985.
Tiedje, J.M., S.A. Brempong, K. Nüsslein, T.L. Marsh, and S.J. Flynn. 1999. Opening the black box of soil microbial diversity. Applied
Soil Ecology 13: 109-122.
Tomback, D.F. and K.C. Kendall. 2001. Biodiversity losses: the downward spiral. pp. 243-2623 in D.F. Tomback, ed. Whitebark pine
communities: ecology and restoration. Island Press, Washington, D.C. 441 p.
Trappe, J.M. 1977. Selection of fungi for ectomycorrhizal inoculation in nurseries. Annual Review of Phytopathology 15:203-222.
Turner C.E., T.D. Center, D.W. Burrows and G.R. Buckingham. 1998. Ecology and management of Melaleuca quinquenervia, an invader
of wetlands in Florida, U.S.A. Wetlands Ecology and Management 5: 165–178.
Volney, J.A., and R.A. Fleming. 2007. Spruce budworm (Choristoneura spp.) biotype reactions to forest and climate characteristics.
Global Change Biology 13: 1630-1643.
Wagner, R.G. 2005. Top ten principles for managing competing vegetation to maximize regeneration success and long-term yields. Pp.
31-35 in: Colombo, S.G., ed. The thin green line: Symposium on the state-of-the-art in reforestation (Proceedings). July 25-28 2005,
Thunder Bay, Ontario, Canada. Ontario Forest Research Institute Forest Research Information Paper 160.
Webb., T., III. 1987. The appearance and disappearance of major vegetational assemblages: Long-term vegetational dynamics in eastern
North America. Plant Ecology 69: 177-187.
White, A., M.G.R. Cannell and A.D. Friend. 2000. The high-latitude terrestrial carbon sink: a model analysis. Global Change Biology 6:
Williamson, M., P. Pyšek, V. Jarošík, and K. Prach. 2005. On the rates and patterns of spread of alien plants in the Czech Republic,
Britain, and Ireland. Ecoscience 12: 424-433.
Wilson, J.W. 1959. Notes on wind and its effects in Arctic-alpine vegetation. Journal of Ecology 47: 415-427.
Winder, R.S., D.E. Macey, and J. Corese. 2010. Characterization of bacteria associated with larvae of mountain pine beetle,
Dendroctonus ponderosae (Coleoptera: Scolytidae). Journal of the Entomological Society of British Columbia 107: 43-56.
Winder, R.S. and S.F. Shamoun. 2006. Forest pathogens: friend or foe to biodiversity? Can. J. Plant Pathology 28: S221-S227.
Woodall, C.W., C.M. Oswalt, J.A. Westfall, C.H. Perry, M.D. Nelson and A.O. Finley. 2009. An indicator of tree migration in forests of
the eastern United States. Forest Ecology and Management 257: 1434-1444.
Woods, A., K.D. Coates and A. Hamann. 2005. Is an unprecedented Dothiostroma needle blight epidemic related to climate change?
BioScience 55: 761-769.
Worrall, J. 1983. Temperature and bud-burst relationships in amabilis and sub-alpine fir provenance tests replicated at different
elevations. Sylvae Genetica 32: 203-209.
Vitt, P., K. Havens, A.T. Kramer, D. Sollenberger, and E. Yates. 2010. Assisted migration of plants: Changes in latitudes, changes in
attitudes. Biological Conservation. 143: 18-27.
Xie, C.-Y., A.D. Yanchuk and G.K. Kiss. 1998. Genetics of interior spruce in British Columbia: Performance and variability of open-
pollinated families in the East Kootenays. Research Report 07. B.C. Ministry of Forests Research Branch, Victoria, B.C.
... Assisted migration is the human-assisted movement of species to habitats that they otherwise cannot currently colonize. It can be applied when it is predicted that a threatened habitat will in the future have the same climate as the current habitat of the target species and it can be separated into three approaches: assisted population migration, assisted range expansion, and assisted long-distance migration (Winder et al. 2011). ...
... On the contrary, Q. frainetto is not native to Germany and Q. pubescens is found only in limited locations in SW Germany (Bussotti et al. 1998, Pasta et al. 2016. Introducing seed material of these oak species from an Italian provenance in Greece would be equivalent to the approach of assisted population migration to enlarge the gene pool of the already existing species (Winder et al. 2011). The attempt to establish the species in central European ecosystems (e.g., in Germany) corresponds to the approach of assisted range expansion, which imitates natural migration processes, but is unlikely to occur at the velocity of climate change due to landscape fragmentation (Vitt et al. 2010, Winder et al. 2011 or slow migration speed of the target species. ...
... Introducing seed material of these oak species from an Italian provenance in Greece would be equivalent to the approach of assisted population migration to enlarge the gene pool of the already existing species (Winder et al. 2011). The attempt to establish the species in central European ecosystems (e.g., in Germany) corresponds to the approach of assisted range expansion, which imitates natural migration processes, but is unlikely to occur at the velocity of climate change due to landscape fragmentation (Vitt et al. 2010, Winder et al. 2011 or slow migration speed of the target species. To test the performance of these oak species, coming from different regions of the Mediterranean area, from the perspective of assisted migration, a replicate common garden experiment has been established (Bantis et al. 2021). ...
Full-text available
Keywords: assisted migration; chlorophyll fluorescence; diurnal variation; gas exchange; Quercus frainetto; Quercus pubescens. Abbreviations: DOY-day of the year; gs-stomatal conductance; OLY-Olympiada/Greece; PIabs-performance index on absorption basis; PNmax-light-saturated net photosynthetic rate; RC/ABS-active reaction centers on absorption base; SWA-Schwanheim/ Germany; WUE-water-use efficiency; φP0-Fv/Fm-maximum quantum efficiency of the reduction of QA; ψE0-the probability that an absorbed photon leads to a reduction further than QA. Oaks may contribute to the stabilization of European forests under climate change. We utilized two common gardens established in contrasting growth regimes, in Greece (Olympiada) and Germany (Schwanheim), to compare the diurnal photosynthetic performance of a Greek and an Italian provenance of two Mediterranean oaks (Quercus pubescens and Q. frainetto) during the 2019 growing season. Although the higher radiation in the southern common garden led to a strong midday depression of chlorophyll a fluorescence parameters (maximum quantum efficiency of PSII, performance index on absorption basis), comparable light-saturated net photosynthetic rates were achieved in both study areas. Moreover, both species and provenances exhibited analogous responses. Q. pubescens had enhanced chlorophyll a fluorescence traits but similar photosynthetic rates compared to Q. frainetto, whereas the provenances did not differ. These findings indicate the high photosynthetic efficiency of both oaks under the current climate in Central Europe and their suitability for assisted migration schemes. Highlights • High radiation led to strong photoinhibition of both oaks at the southern site • At both sites, only small differences were observed between species and provenances • Both oak species show high potential for assisted migration to Central Europe
... A conservative approach for assisted migration would be to favor species that are currently present at a planting site or nearby (M. I. Williams & Dumroese, 2013;Winder et al., 2011). Furthermore, when faced with a choice of seedlots with similar climatic matches, a conservative approach would be to favor the seedlot that is geographically closer. ...
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The Seedlot Selection Tool and Climate‐Smart Restoration Tool are web‐based tools designed to match seedlots with planting sites assuming that seedlots are adapted to the past climates in which they evolved, primarily with respect to temperature and aridity. The tools map the climatic match of seedlots with the past or projected climates of planting sites. The challenge is that future climates are a moving target, which means that seedlots must be adapted to the near‐term climates as well as the climates of the mid‐ to late‐21st century. Because climate projections are uncertain, the prudent approach is to aim for the warmest climate that may be expected while ensuring that seedlots moved from warmer to colder locales are not moved so far that they risk cold damage. Uncertainty in climate projections may be mitigated by ensuring genetic diversity through mixing seed sources and having collections from many parents per seed source. Three examples illustrate how to effectively use the web tools: (1) choosing seedlots targeting different future climates for a mid‐elevation Douglas‐fir site in the Washington Cascades, (2) finding current and future seed sources for restoration of big sagebrush after fires in the Great Basin and Snake River Plain, and (3) planning to ensure that a Douglas‐fir seed inventory includes seedlots suitable for future climates in western Oregon and Washington.
... In the Northern hemisphere, assisted migration usually means moving seedlings from southern to more northern locations, with managers aiming to match genotypes and plantations sites using the climatic information at the source site, such as temperature and mean precipitation e.g., [7]. Genotypes can also be selected for their resistance to specific climate-related threats, such as drought [8]. However, climatic conditions are not the only problems faced by translocated seedlings. ...
Full-text available
To facilitate forest transition to future climate conditions, managers can use adaptive silvicultural tools, for example the assisted translocation of tree species and genotypes to areas with suitable future climate conditions (i.e., assisted migration). Like traditional plantations, however, assisted migration plantations are at risk of failure because of browsing by ungulate herbivores. The ability of seedlings to tolerate browsing could also be hampered by low water availability, as is expected under climate change. Using a greenhouse experiment with five eastern North American tree species, we evaluated the effects of simulated winter browsing and reduced water availability on the growth (total biomass, shoot:root ratio), survival, and chemical composition (nitrogen, total phenolics, flavonoids) of seedlings. We compared seedlings from three geographic provenances representing three climate analogues, i.e., locations with a current climate similar to the climate predicted at the plantation site at a specific time (here: current, mid-century and end of the century). We hypothesized that seedlings would allocate resources to the system (shoots or roots) affected by the most limiting treatment (simulated browsing or reduced water availability). Additionally, we evaluated whether the combination of treatments would have an additive or non-additive effect on the growth, survival and chemical composition of the seedlings. Quercus rubra seedlings reacted only to the water reduction treatment (changes in biomass and N concentration, dependent on geographic provenance) while Pinus strobus reacted only to the simulated browsing treatment (biomass and chemical composition). We also observed non-additive effects of reduced water availability and simulated browsing on Prunus serotina, Acer saccharum and Thuja occidentalis. In general, shoot:root ratio and investment in chemical defense did not vary in response to treatments. The regrowth response observed in Q. rubra and A. saccharum suggests that these species could tolerate periodic browsing events, even when water availability is reduced. More information is required to understand their long-term tolerance to repeated browsing events and to harsher and more frequent water stress. We highlight the importance of species-specific growth and allocation responses that vary with geographic provenance, which should be considered by managers when planning climate-adapted strategies, such as assisted migration.
... Planting options for assisted migration range from single-species enrichment planting under partial forest cover, to multi-species plantations in open conditions. Such forestry assisted migration (sensu Pedlar et al., 2012) could increase forest resilience to disturbance (e.g drought, insect outbreaks) and maintain ecosystem services and productivity (Aubin et al., 2011;Ste-Marie et al., 2011;Winder et al., 2011). Scientists and land managers are increasingly recognizing the role assisted migration could play in adaptation scenarios of forest ecosystems (Johnston et al., 2009;Lal et al., 2012;Government of Canada, 2019). ...
To promote the sustainability of forest ecosystems and maintain ecosystem services to human populations in the context of climate change, forest managers are considering several adaptation tools. One of those, assisted migration, consists of displacing tree species and/or populations to locations with suitable future climate conditions. Assisted migration plantations, however, could fail to produce viable forests under high herbivory pressure from mammalian herbivores. Thus, selecting species and genotypes with low susceptibility to herbivores could be a key condition for the recruitment of translocated seedlings. We developed an approach to predict susceptibility to mammalian herbivores, based on the seedlings’ chemical content (a proxy of phytochemical defence and susceptibility to herbivores). We used the approach on eight North American tree species of three climate analogues regions each (i.e. locations where the current climate is similar to the future climate at plantation site). We built chemical profiles and ranked species and climate analogue based on their potential susceptibility. To assess the reliability of our approach, we compared the chemical profiles to a systematic review of these species’ chemistry and of mammalian browsing throughout their native geographic range. For most species, our chemical profiles and browse susceptibility rankings were congruent with information available in the literature, both for phytochemical defence and for browsing. Two of the eight species (Pinus strobus and Thuja occidentalis) were more susceptible than predicted based on their chemical profile. These discrepancies could be linked to specific mammalian herbivores, that were unaffected by the phytochemical defence of these species. We observed a generally higher susceptibility of broadleaf species, which could be taken into account when devising adaptive silvicultural strategies. Furthermore, we propose the chemical profiling approach as a preliminary screening tool to identify species more resistant to mammalian herbivores, but also potentially to other herbivores and pathogens. Our chemical profile approach, based on the objective assessment of multivariate analyses results, could be replicated to compare the potential susceptibility of other species. This approach could be especially useful when contending with novel plant-herbivore relationships, such as forestry and conservation assisted migration or species invasion.
... Ste-Marie et al., 2011). Winder et al. (2011) discussed the ecological constraints and consequences of AM and options for their mitigation at three scales: translocation over long distances (assisted long-distance migration), translocation just beyond the range limit (assisted range expansion) and translocation of genotypes within the existing range (assisted population migration). They concluded that, from an ecological perspective, AM is a feasible management option for tree species. ...
This book contains a series of chapters reviewing the current scientific knowledge on RNAi, methods for developing RNAi systems in transgenic plants and a range of applications for crop improvement, crop production and crop protection. Some chapters examine both endogenous systems in transgenic plants and exogenous systems where interfering RNAs are applied to target plants, pests and pathogens. The biosafety of these different systems is examined and methods for risk assessment for food, feed and environmental safety are discussed. Finally, aspects of the regulation of technologies exploiting RNAi and the socioeconomic impacts of RNAi technologies are discussed.
... Incertitude des modèles utilisés pour l'identification des sites récepteurs. 20,22,29,32,33,38,51,62 ...
... To help increase the adaptive capacity of vulnerable tree populations, human-assisted movement of tree species' populations via assisted gene flow, i.e. the migration of pre-adapted alleles/genotypes, is envisioned 15,16 . While assisted gene flow, as an adaptive measure, could help maintain and even enhance the health and growth of forests, it might also contribute to climate mitigation through more permanent sequestration of forest carbon 17,18 . Notably, by influencing the rates of tree survival and growth, assisted gene flow could help increase the carbon stored during stand development. ...
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Assisted gene flow between populations has been proposed as an adaptive forest management strategy that could contribute to the sequestration of carbon. Here we provide an assessment of the mitigation potential of assisted gene flow in 46 populations of the widespread boreal conifer Picea mariana, grown in two 42-year-old common garden experiments and established in contrasting Canadian boreal regions. We use a dendroecological approach taking into account phylogeographic structure to retrospectively analyse population phenotypic variability in annual aboveground net primary productivity (NPP). We compare population NPP phenotypes to detect signals of adaptive variation and/or the presence of phenotypic clines across tree lifespans, and assess genotype‐by‐environment interactions by evaluating climate and NPP relationships. Our results show a positive effect of assisted gene flow for a period of approximately 15 years following planting, after which there was little to no effect. Although not long lasting, well-informed assisted gene flow could accelerate the transition from carbon source to carbon sink after disturbance.
Technical Report
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In this report, we summarize the current state of knowledge and best estimates of how climate change is expected to impact Norwegian forest ecosystems from now to the year 2100
Anthropogenic climate warming is undisputed and yet, there is much that is unknown regarding biological impacts of changing temperature and precipitation, and the management options presented as solutions to biodiversity losses—such as translocations of plants and seeds—are often controversial. This Special Focus presents five new studies and two recently published articles in Journal of Ecology and Ecological Solutions and Evidence that analyse the use of existing plant translocations, assess the potential to use translocations to offset past declines and future population losses, and articulate the need to involve a broader range of stakeholders in the use of translocations. Our recommendations include improving the monitoring of plant translocations; harnessing existing translocations to form a global monitoring array for climate change impacts on biological diversity and strengthening research–practice linkages improving data sharing and broadening the stakeholders involved in translocations. Synthesis. Further study of translocations and integration of information from translocation projects holds the opportunity to provide a wealth of new knowledge and improved ways to counter climate change impacts on biological diversity and ecosystems.
The assisted migration (AM) of trees is increasingly being proposed and trialed to adapt forest management to the impacts of climate change. While institutional and risk perception dimensions of AM are increasingly well-studied, a key gap that remains is to understand how current institutional practices shape the types of knowledge that are considered in AM policy development, and how this in turn makes visible different risks and benefits. In this study, we use a politics of knowledge lens applied to the case of British Columbia, Canada, where AM policy is currently in place, to examine the types of knowledge informing AM thus far, and how that knowledge shapes perceived AM risks and ways of addressing them. Based on 27 in-depth, semi-structured interviews with key government employees and forest industry professionals involved with the development and implementation of AM, we find an overall optimistic view of AM. However, the type of knowledge deemed credible to inform AM decision-making is restricted to biophysical, model-based, scientific knowledge. This primarily biophysical framing of AM arises from the objectives and worldviews of actors working in the AM space and gives rise to relatively narrow ways of understanding potential AM risks and solutions to them. While policymakers and government scientists recognize the need to engage industry, Indigenous Peoples, and the general public, these groups are seen as knowledge receivers. We argue that these beliefs about what counts as credible expertise (and who can produce it) have served to exclude other knowledge forms from being considered in decision-making, and in so doing, have limited possibilities for generating transformative change.
Loblolly pine (Pinus taeda L.) megagametophytes and embryos were examined electrophoretically to compare the extent and distribution of genetic variability in allozymes of selected and wild populations. Range-wide collections of three different types were investigated in this study. These consisted of seed sampled from, 1. a provenance test established in 1953, 2. bulk seed sampled from collections obtained from natural stands, and 3. seed harvested from clones used to produce improved seed in a tree improvement program. All 18 loci tested were found to be polymorphic. The average number of alleles overall (N(a)) was 3.8. Expected heterozygosities (H(e)) varied from 0.193 in the 70-year old orchard clones, to 0.174 in the 40-year-old provenance test samples, to 0.163 in the embryos of the bulk collections. The maximum F(ST) was 0.066 for the provenance test populations, which indicates that only a small proportion (6.6 %) of the total variation in allozymes was attributed to population differences. In spite of this, the populations were well differentiated in multivariate analysis. In controlled-pollinated progeny tests of the orchard selections, there was a negative association between growth and the presence of rare alleles in the parent. A rare allele at the IDH locus was associated with slower growth, probably because it indicated hybridization with the slower-growing shortleaf pine (P. echinata MILL.). Allozyme variation as well as variation in cortical monoterpenes and fusiform rust resistance suggests that loblolly pine resided in two refugia during the Pleistocene; one in south Texas / northeast Mexico and one in south Florida / Caribbean. The two populations migrated to the northern Gulf Coastal Plain at the beginning of the Holocene and merged just east of the Mississippi River.
The first experiences of growing conifers in Greenland are reported. At the interior fjords of SW Greenland around 61°N, away from the more or less ice-filled outer coasts, the average temperature for July is <10°C and the heat sum for the growing season (threshold 5°C) is <600 degree-days. However, the natural vegetation indicates a potential for the occurrence of a subarctic-subalpine conifer tree-line. Furthermore, small scale planting trials since 1953 have resulted in groups of 7-9 m tall Larix sibirica (from Krasnoyarsk and Ural provenances), 4-7 m tall groups of Picea abies as well as Pinus sylvestris (northern Scandinavia) and Picea glauca and P. glauca × P. sitchensis (Picea × lutzii) (coastal southern Alaska). On this basis, the Greenland Arboretum at Narssarssuaq International Airport has been developed. It was started in 1976 within a 200 ha area. It is developing into a representation of primarily tree-line forming species and provenances from circumpolar regions. Until now, Russian material has been raised from seed whereas Fennoscandian and North American material has mostly been collected as seedlings. At present, 100,000 trees have been planted, 95% of which are conifers. Of the 450 provenances represented, 257 are conifers. To secure a common character and quality of the landscape, 50,000 seedlings of Larix sibirica var. sukaczewii (from seven provenances) have been planted in a mosaic with the natural scrub. The results show that for large scale planting in SW Greenland Larix sibirica var. sukaczewii is the best choice, followed by successful provenances of Abies lasiocarpa, Picea engelmannii, P. glauca and its hybrids with P. engelmannii and P. sitchensis, Pinus cembra var. sibirica, Picea abies and Pinus sylvestris. Like other arboreta, the collections serve scientific, educational and amenity purposes.
Projected climate warming will potentially have profound effects on the earth's biota, including a large redistribution of tree species. We developed models to evaluate potential shifts for 80 individual tree species in the eastern United States. First, environmental factors associated with current ranges of tree species were assessed using geographic information systems (GIS) in conjunction with regression tree analysis (RTA). The method was then extended to better understand the potential of species to survive and/or migrate under a changed climate. We collected, summarized, and analyzed data for climate, soils, land use, elevation, and species assemblages for >2100 counties east of the 100th meridian. Forest Inventory Analysis (FIA) data for >100 000 forested plots in the East provided the tree species range and abundance information for the trees. RTA was used to devise prediction rules from current species-environment relationships, which were then used to replicate the current distribution as well as predict the future potential distributions under two scenarios of climate change with twofold increases in the level of atmospheric CO2. Validation measures prove the utility of the RTA modeling approach for mapping current tree importance values across large areas, leading to increased confidence in the predictions of potential future species distributions. With our analysis of potential effects, we show that roughly 30 species could expand their range and/or weighted importance at least 10%, while an additional 30 species could decrease by at least 10%, following equilibrium after a changed climate. Depending on the global change scenario used, 4-9 species would potentially move out of the United States to the north. Nearly half of the species assessed (36 out of 80) showed the potential for the ecological optima to shift at least 100 km to the north, including seven that could move >250 km. Given these potential future distributions, actual species redistributions will be controlled by migration rates possible through fragmented landscapes.
Some alien tree species used in commercial forestry, and agroforestry cause major problems as invaders of natural and seminatural ecosystems. The magnitude of the problem has increased significantly over the past few, decades, with a rapid increase in afforestation and changes in land use. Trends can be explained by analyzing natural experiments created by the widespread planting of a small number of species in different parts of the world. The species that cause the greatest problems are general those that have been planted most widely and for the longest time. The most affected areas have the longest histories of intensive planting. Pinus spp. are especially problematic, and at least 19 species are invasive over large areas in the southern hemisphere, where some species cause major problems. The most invasive Pinus species have a predictable set of life-history, attributes, including low, seed mass, short juvenile period, and short interval between large seed crops. Pine invasions have severely, impacted large areas of grassland and scrub-brushland in the southern hemisphere by causing shifts in life-form dominance, reduced structural diversity, increased biomass, disruption of prevailing vegetation dynamics, and changing nutrient cycling patterns. The (unavoidable) negative impacts of forestry with alien species are thus spilling over into areas set aside for conservation or water production. There is an urgent need to integrate the various means available for reducing the negative impacts of current invaders and to implement protocols to regulate the translocation of species that are known to be invasive.
Examination of air photos from 1930, 1970 and 2002 revealed stands of the European Scots Pine (Pinus sylvestris) invading remnants of natural Corema (Corema conradii) heathland in the Annapolis valley. To document the impact of the introduced pines, four natural habitats were compared with two adjacent habitats already invaded by the pines. All surveyed habitats had been dominated by Corema heath based on air photos taken in 1930. Twenty 1 m2 quadrats were used to record presence and cover of vascular plants at each site. The invasive alien pines reduce the native cover to 12%. Vascular plant biodiversity is reduced to less than 42% and the cover of the heathland dominant, Corema conradii, is reduced from over 100 % to less than 2%. with Deschampsia flexuosa becoming the dominant species. The modified ecosystem and loss of biodiversity has economic impacts through loss of pollinators of agricultural crops and loss of germplasm of native crop relatives.
We outline the major steps involved in implementing assisted migration (AM) and assess, in a general way, the capacity to carry out each step in Canadian forests. Our findings highlight the fact that capacity to implement AM differs between forest species; in particular, the existence of established provenance trials, seed transfer guidelines, seed procurement systems, and plantation establishment protocols makes AM considerably more feasible for most commercial tree species than for most species of conservation concern. We report on several AM efforts involving commercial tree species that are already underway in Canada and identify a number of initiatives that could be undertaken to help build AM capacity. This paper is not intended as an endorsement of the AM approach; however, we feel there is considerable value in discussing implementation issues at this point in the AM debate.
Current models of climate change predict a reduction of area covered by northern coniferous forests and tundra, and an increase in grasslands. These scenarios also indicate a northerly shift in agricultural regions, bringing virgin soils under cultivation. The direct effects of man on tundra, boreal forest, and temperate grassland ecosystems are likely to result in less carbon mobilization from soils and vegetation than from tropical forests. However, as a consequence of climate change, carbon mineralization rates from arctic and sub-arctic soils could be very rapid under warmer and drier conditions because of low stabilization of soil organic matter (SOM) and enhanced microbial responses to small changes in soil moisture and temperature. Predicting the response of these systems to climate change is complicated where the edaphic environment regulating SOM dynamics is not a direct function of macroclimatic conditions. The primary recommendation for future research is for integrated studies on plant and soil processes. -from Author
This study documents changes in the distribution of non-native woody species from 1938 to 1999 within 30 forests of a $54\>km^2$ landscape in Monroe County, New York. Within these forests, the mean number of exotic species increased nearly three-fold from 1938 to 1999, and two species not naturalized within the landscape in 1938, Lonicera morrowii and Rosa multiflora, had become widespread by 1999. In 1999, the most abundant exotic was Lonicera morrowii, which had greater cover within forests on wetter portions of the landscape where much of the surrounding agricultural land had been abandoned. Though exotics accounted for < 10 % of relative cover within the shrub layer and < 1 % of the tree layer, their cover was higher in portions of some forests, especially near edges. In 1999, six species had covers > 25 % within patches $> 100\>m^2$: Acer platanoides, Crataegus monogyna, Ligustrum vulgare, Lonicera morrowii, Robinia pseudoacacia and Rosa multiflora. These species in particular may represent on-going invasions that could alter this landscape's forested habitats.