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The Deciduous Forest – Boreal Forest Ecotone


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Ecotones have been subject to significant attention over the past 25 years as a consensus emerged that they might be uniquely sensitive to the effects of climate change. Most ecotone field studies and modeling efforts have focused on transitions between forest and non-forest biomes (e.g. boreal forest to Arctic tundra, forest to prairie, subalpine forests to alpine tundra) while little effort has been made to evaluate or simply understand forest–forest ecotones, specifically the deciduous forest – boreal forest ecotone. Geographical shifts and changes at this ecotone because of anthropogenic factors are tied to the broader survival of both the boreal and deciduous forest communities as well as global factors such as biodiversity loss and dynamics of the carbon cycle. This review summarizes what is known about the location, controlling mechanisms, disturbance regimes, anthropogenic impacts, and sensitivity to climate change of the deciduous forest – boreal forest ecotone.
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The Deciduous Forest – Boreal Forest Ecotone
David Goldblum*and Lesley S. Rigg
Department of Geography, Northern Illinois University
Ecotones have been subject to significant attention over the past 25 years as a consensus emerged
that they might be uniquely sensitive to the effects of climate change. Most ecotone field studies
and modeling efforts have focused on transitions between forest and non-forest biomes (e.g. boreal
forest to Arctic tundra, forest to prairie, subalpine forests to alpine tundra) while little effort has
been made to evaluate or simply understand forest–forest ecotones, specifically the deciduous forest
– boreal forest ecotone. Geographical shifts and changes at this ecotone because of anthropogenic
factors are tied to the broader survival of both the boreal and deciduous forest communities as well
as global factors such as biodiversity loss and dynamics of the carbon cycle. This review summarizes
what is known about the location, controlling mechanisms, disturbance regimes, anthropogenic
impacts, and sensitivity to climate change of the deciduous forest – boreal forest ecotone.
Over the past century there have been numerous descriptive field-based studies cataloging
patterns of boundaries (ecotones) between vegetation types (biomes). These studies were
generally followed by research focused on understanding the biological processes and or
environmental conditions creating the patterns. Recent research has addressed the tempo-
ral and spatial dynamics of ecotones, particularly in light of anthropogenic disturbances
such as climate change, logging, agriculture, and altered fire regimes. Ecotones may be
obvious (Arctic alpine treelines or prairie forest boundaries) or more subtle (between for-
ested zones). Because of quantifiable geographic shifts in ecotone locations over recent
decades (e.g. Beckage et al. 2008; Parmesan 2006), centuries (e.g. Vallee and Payette
2004), and millennia (e.g. Hupy and Yansa 2009; Kullman 1995), many ecologists and
biogeographers have suggested that ecotones may be well suited to detect human impacts
on terrestrial ecosystems, including signs of anthropogenic climate change (Kupfer and
Cairns 1996; Loehle 2000; Neilson 1993; Noble 1993). This review summarizes the
current state of knowledge about an infrequently studied ecotone between forested zones,
namely the deciduous forest – boreal forest ecotone (DBE). We consider the position of
the DBE since the end of the last ice age (18,000 years BP) across the northern hemi-
sphere, the environmental, ecological, and biological variables responsible for the transi-
tion from deciduous to boreal forest, the dynamics of the DBE during the Holocene, and
anthropogenic impacts on the DBE. Lastly, we include a discussion of the ecotone’s
future given modeled anthropogenic climate change.
According to Ries et al. (2004), the earliest reference to edge-related ecology was by the
influential ecologist geographer Clements (1907) who first introduced the term ‘ecotone’,
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whereas Livingston (1903) described a distinct boundary between forest types as a ‘zone
of tension’. In both cases, according to van der Maarel (1990), these descriptions refer to
a sharp ‘stress zone’, which is distinctly different from a ‘gradient zone’ which can be
termed an ecocline, a term coined by Clements (1937) and is associated with large-scale
community change. A third type of boundary has been proposed by van der Maarel
(1990), the mosaic, to represent areas with intermixed fragments of the adjacent communi-
ties, which may be most appropriate for describing the deciduous boreal forest transition
discussed in this article (Figure 1). We refer to boundaries between vegetation biomes as
ecotones, a term commonly used to describe boundaries and biome transitions at the global
scale (Kent et al. 1997). However, implied in our use of this term is that biome-level
boundaries may be abrupt, gradual, or composed of vegetation mosaics.
Generally, climate driven air mass activity creates global-scale biome patterns (Risser
1995). However, at progressively finer spatial scales, a hierarchy of biotic and abiotic con-
straints on vegetation community types becomes evident (Gosz 1993) leading to fine scale
transition zones (van der Maarel 1990). At regional and local scales, soil characteristics,
microclimatology, microtopography, competition, and population genetics ultimately
determine the exact position of the ecotone (van der Maarel 1990).
For much of the 20th century many ecologists viewed ecotones as anomalies (Fortin
et al. 2000; Yarrow and Marin 2007), but an interest in ecotones arose in the 1990s.
Firstly, ecotones were seen as controlling the flux of materials between ecosystems (Risser
1995), population dynamics, and biodiversity (Naiman and De
´camps 1990; Risser 1995).
Secondly, because ecotones contained species pushed to their physiological tolerance,
ecotones should be especially sensitive to environmental fluctuations, and biological
changes would be detectable (Arris and Eagleson 1994; Loehle 2000), and thus could be
bellwethers of anthropogenic impacts (Fortin et al. 2000; Neilson 1993), although this
may not apply to all ecotones (Kupfer and Cairns 1996; Noble 1993).
The deciduous forest – boreal forest ecotone
Unlike other vegetation zones on Earth, only the boreal forest biome encircles the
globe (Woodward 2003). The biomes at the southern boundary of these forests vary
Fig. 1. A generalized diagram expressing the variable nature that might exist at community (or biome) transitions
(Source: Kent et al. 1997).
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geographically. Where oceanic climates predominate, the boreal zone borders deciduous
forests; while under more xeric climates, steppe, grassland, or semi-desert adjoin boreal
forests (Breckle 2002) as is the case in large tracts of northern Asia and central North
America. Generally, there is no distinct boundary between boreal and deciduous forests;
instead, a broad transition zone exists composed of mixed stands of coniferous and
deciduous species, or a ‘macromosaic-like arrangement’ with pure stands of deciduous
trees on favorable sites and pure coniferous stands on less favorable sites found on poor
soils (Breckle 2002).
Several studies have described the complexities of locating the DBE in Asia. Sukachev
(1928) describes the DBE in European Russia, coinciding with chernozem soils, as a
broad zone associated with a general decrease in oak, maple, and lime and dominance by
Picea. Breckle (2002) describes the European DBE coinciding with the northern limit of
oak at around 60N. Given the mountainous terrain of Japan, the DBE is expressed as
elevational transitions (Ohsawa 1984, 1990; Yoshino 1978), similar to mountainous China
(Tang and Ohsawa 1997, 2002). Hou (1983), surveying China’s vegetation zones, sug-
gests that the DBE is found only in the extreme northeastern portion of the continent;
yet it occurs as an altitudinal boundary on mountainous terrain (Pastor and Mladenoff
1992). The DBE in North America is largely intact running through the Great Lakes
from Minnesota, Wisconsin, east-central Ontario (Figure 2), and ultimately into southern
Quebec and northern Maine (Breckle 2002; Pastor and Mladenoff 1992) (Figure 3).
Additionally, similar ecotones are found along the Appalachian Mountains in eastern
North America (Beckage et al. 2008). The DBE in Eurasia is highly anthropogenically
Fig. 2. A southward view from the boreal forest towards the northern limit of the deciduous forest in Lake Supe-
rior Provincial Park, Ontario, Canada. The trees (with a slight red tinge) on the ridges of the distant hills are the
northernmost sugar maple in the region, marking the transition to boreal forest. Photo credit: David Goldblum.
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disturbed, making its description somewhat speculative compared with the more quantifi-
able and intact North American DBE (Pastor and Mladenoff 1992).
Factors determining the location of the DBE
A suite of environmental factors determine the location of the DBE across space and
time. The extent of the ecotone is not consistent globally; some regions in Europe and
North America have a much more developed mixed-forest community (Pastor and
Mladenoff 1992). In some areas the latitudinal ecotone coincides with an elevational
ecotone (i.e. parts of China), but for the purposes of this discussion the focus will be on
variables associated with changing latitude rather than elevation. At a continental scale
ecotones are strongly correlated with climate factors such as temperature, monthly precip-
itation, seasonality (i.e. growing degree-days), and potential evapotranspiration (Parmesan
et al. 2005; Sowell 1985; Stephensen 1998; Woodward and Williams 1987). While stud-
ies that broadly correlate vegetation–climate associations are common in the literature,
regional scale determinants may include regional water balance deficits (Stephensen
1998), ecophysiological plant response (Arris and Eagleson 1994; Stephensen 1998;
Woodward and Williams 1987), actual evapotranspiration (Stephensen 1998; Thornthwa-
ite 1948), and extreme minimum temperature (Woodward and Williams 1987). At the
landscape scale, boreal and deciduous species within the DBE tend to establish and persist
along environmental gradients determined not by climate alone but rather by subtle varia-
tion in substrate, drainage (local watershed dynamics), physical soil properties and nutrient
availability (Pastor and Mladenoff 1992).
Climate is frequently identified as an important variable in determining the location of
the DBE (Prentice et al. 1992). Arris and Eagleson (1989) discuss the coincidence of the
DBE with the )40 C average annual minimum temperature isotherm (Figure 4). Trees
common in the deciduous forest experience cellular damage with temperatures below
)40 C, whereas boreal and northern tree species tolerate colder extremes through deep
Fig. 3. The location of the deciduous-boreal forest ecotone in eastern North America. Vegetation zones are based
on several sources (Sources: Minnesota Department of Natural Resources, Natural Resources Canada, USDA Natural
Resources Conservation Service, Wisconsin Geological and Natural History Survey, Scott 1995; Stearns 1997;
Watkins 2006).
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super-cooling to avoid intercellular freezing (Lee et al. 2005; Lenihan and Neilson 1993;
Sakai 1975; Sakai and Weiser 1973; Woodward and Williams 1987). In China, Liu et al.
(1998) found that accumulated seasonal warmth for months over 5 C (warmth index),
not necessarily minimum temperatures, explained the location of the DBE. Additionally,
growing season length, growing degree-days (GDD), and frost-free period have been
offered as factors in determining the location of the DBE (Arris and Eagleson 1994; Kup-
fer and Cairns 1996; Lenihan and Neilson 1993; Neilson 1995; Pastor and Mladenoff
1992; Prentice et al. 1992). Kupfer and Cairns (1996) describe a growing season length
threshold (generally four months), and Prentice et al. (1992) delineate a 1200 GDD
threshold, leading to a transition to conifers because of the greater water and growing
season length requirements of the more photosynthetically efficient, but thermally sensi-
tive deciduous leaves. While these constraints explain the northern limit of the deciduous
forest, competitive limitations have been proposed as accounting for the southern limit of
the boreal forest as a longer growing season, coupled with higher angiosperm photosyn-
thesis rates effectively leads to competitive exclusion of conifers (Arris and Eagleson 1994)
and dominance by deciduous species in the deciduous forest biome.
Soil formation is coincident with vegetation development but responsive to the underly-
ing substrate, climate, topography, and time (Jenny 1994). Throughout the DBE the
dominant soils are spodosols and are relatively young, having formed as the most recent
deglaciation, with sparse mineral soil (Kellman 2004). In North America, these soils are
developing in some locations on the Canadian Shield, and in other locations on relatively
unweathered glacial deposits or till or outwash (Kellman 2004; Pastor and Mladenoff
1992). Spodosols throughout the DBE in both North America and Europe (Podosols) are
characterized by a sandy texture, generally low nutrient status, low pH, and organic
matter accumulation (Elgersma and Dhillion 2002; Kellman 2004). The location of the
DBE is responsive to climate, but changes in the edaphic conditions within the ecotone
play a role in the distribution of species because of changing soil nutrient status and pH
(Barras and Kellman 1998; Demers et al. 1998; Elgersma and Dhillion 2002; Messaoud
et al. 2007; Pastor and Mladenoff 1992). The connection between edaphic conditions
Fig. 4. The observed northern limit of the deciduous forest (short dashed line) in North America and the )40 C
average annual minimum temperature isotherm (long dashed line) (Source: Arris and Eagleson 1989).
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and climate is paramount, with reduced soil temperatures slowing the rates of decomposi-
tion causing an increased accumulation of organic matter and increased acidity (Messaoud
et al. 2007).
Feedbacks between litter quality (Barras and Kellman 1998) and nutrient availability
(particularly soil nitrogen) have been found to strongly impact the location of boreal
versus deciduous stands within the ecotone (Pastor and Mladenoff 1992). Nitrogen is the
critical limiting nutrient (Fonara and Tilman 2008; Reich and Oleksyn 2004) in both the
deciduous and boreal forest biomes (Pastor and Mladenoff 1992; Reich et al. 1995). The
feedback between soil quality (including microbial communities), litter quality, and
species presence absence is crucial within the DBE. Soil moisture may also be a factor in
explaining the location of the DBE (Weishampel et al. 1999). Hogg (1994) found that
the southern limit of several boreal forest conifers coincided with zero isoline of annual
precipitation minus potential evaporation.
Maycock and Curtis (1960), who surveyed the forests around Lake Superior to quantify
the forest composition associated with the DBE, identified a mosaic of stands dominated
by conifers distinguishable from stands dominated by deciduous hardwoods. Where the
mediating influence of Lake Superior accentuates topographically induced changes in
climatic conditions at the hillslope scale, sugar maple dominated stands are found on
ridges, whereas conifers dominate low-lying waterlogged sites (Barras and Kellman 1998;
Boucher et al. 2009; Goldblum and Rigg 2002; Pastor and Mladenoff 1992). Similarly,
Hayes et al. (2007) found that the DBE in the Appalachian Mountains varied based on
topographic position, and Messaoud et al. (2007) found that species distributions at the
DBE in Quebec were also governed in part by topographic position. So, while the general
location of the DBE may be climatically controlled, fine scale environmental heterogeneity
creates pockets of stands of one forest type or another, as well as accounts for the broad
transition between the two biomes in North America (Arris and Eagleson 1994).
Fire regimes
Similar to topography, fire may modify the climatically mediated location of the ecotone
by altering competitive interactions and resource availability, yet the role fire plays in
modifying the location of the DBE is poorly understood (Pastor and Mladenoff 1992).
The flammability and role of fire varies dramatically between the boreal and deciduous
forest biomes (Heinselman 1973; Runkle 1990) with fire return intervals increasing as
one moves southward across the DBE from 50 to 80 years in the boreal forest to
>300 years in the deciduous forests of North America (Pastor and Mladenoff 1992).
Despite the contrast in return intervals, Bergeron et al. (2004) determined that the transi-
tion between the mixed and coniferous forests in the ecotone cannot be simply explained
by a difference in fire frequency over the past three centuries, but rather is due to fire
size and severity, with small fires favoring deciduous dominance and larger, intense fires
favoring boreal communities. In central Sweden Axelsson et al. (2002) describe that
anthropogenic modification of the fire regime dramatically altered the location and pres-
ence of deciduous species within the boreal forest matrix. Similarly, Clark and Royall
(1995) demonstrate changes to forest composition in the Canadian DBE associated with
Native American burning, sufficient to tip dominance between biome types. Time
since fire, as is true elsewhere affects species composition, but in the ecotone these
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compositional changes may lead to a transition from deciduous forest to boreal forest, or
vice versa (Gauthier et al. 2000), a phenomenon not often associated with successional
processes outside the DBE.
Historical legacies
While environmental variables are extremely important in determining species distribu-
tions, much remains unexplained at the local scale. Messaoud et al. (2007) identify site
history as explaining a significant portion of the variation in the distribution of a common
ecotone species (balsam fir) at the North American DBE. Specifically, Messaoud et al.
(2007) identify initial site colonization patterns and accidental elimination and replace-
ment as important agents modifying species distributions, altering the subtle location of
the ecotone. Further, Bergeron et al. (2004) describe lags in changes to disturbances
regimes and positive feedbacks between forest structure and fire regimes that may con-
found the association of forest type with environmental conditions.
Anthropogenic and natural disturbances
Ecotones reflect ongoing competitive tensions between species living at the extremes of
their range, and whereas large-scale climate fluctuations may be critical in determining
the broad establishment of ecotones, subtle environmental changes may tip the balance in
favor of one species (or one biome) or another. External factors that may affect the com-
petitive relationship between ecotonal species include fire, treefall, species-specific insect
outbreaks, and the ability of one species to create microenvironments that inhibit the
establishment of potential competitors (Barras and Kellman 1998; Wilson and Agnew
1992). This has occurred over the past centuries and millennia as climate fluctuations and
human impacts on disturbance regimes (Colombaroli et al. 2008; Miettinen et al. 2002)
have given certain species a slight competitive advantage reflected in the establishment
and survival of ecotonal species (Noble 1993).
Paleo-records and mitochondrial DNA chronologies place humans in Europe 40–
60,000 years BP and North America approximately 20,000–18,000 years BP (Forster
2004). Coincident with the onset of ice sheet retreat in North America and Europe
(18,000 years BP), human populations spread from their glacial maximum refugia (For-
ster 2004), forever changing the forests of these two continents. Paleontological, archeo-
logical and paleobotanical evidence supports the notion that a combination of climate
factors and human hunting were responsible for the extinction of Pleistocene megafauna
(Barnosky et al. 2004) that were integrated into the forest ecosystems present at the time
(Donlan et al. 2006). Those flora–fauna biological interactions are absent today.
While early human populations most likely played a key role in altering forest structure
of the DBE, it is the movements, more recently, of colonial and post-colonial Europeans
who have had the greatest impact on the DBE globally (Delcourt and Delcourt 1987). In
both Europe and North America, the boreal forest remains largely intact, whereas the
deciduous forest has been extensively utilized for many centuries for farming and habita-
tion (Pastor and Mladenoff 1992). This is not to say the boreal forest, especially at its
southern margins is not highly managed for timber and other activities, but in compari-
son, the history of human land-use within this ecotone, is more pronounced on the
deciduous side of the DBE.
The DBE has shifted as glaciers have advanced and retreated during the past 2–3
million years across North America and Eurasia (Brubaker 1988; Delcourt and Delcourt
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1987; Hewitt 2000), and the DBE has been shown to migrate with Holocene climate
fluctuations (Hupy and Yansa 2009) as well. In the past few decades there has been
extensive research examining the climatic conditions during and since the last glacial
maximum (Davis 1983;2000; MacDonald 2003; McLachlan et al. 2005; Williams et al.
2007). Ice cores from Greenland and Antarctica provide the most detailed information
today for reconstructing past climates using the relative abundances of oxygen isotopes
(MacDonald 2003). Reconstructed proxy temperatures suggest that 140,000 years BP was
as warm as today, but that by 100,000 years BP the temperature had cooled by 6 C.
Temperatures began to warm approximately 10,000 years BP, with the last glacial maxi-
mum occurring approximately 20,000 years BP, coinciding with a period of particularly
cold temperatures (MacDonald 2003). The ice sheets at their maximum extended far into
central Europe and northern North America.
When Europeans began to colonize North America, the forest was not untouched by
human influence; the area today that is largely farmland, urban centers, and secondary
forest, was at the time dominated by deciduous and boreal forest (Davis 1983), but was
inhabited by Native Americans. Based on the period immediately prior to European pres-
ence, there is evidence from fire-scarred trees that Native Americans were a source of
frequent fires in the North American Great Lakes region (Loope and Anderton 1988),
although their impacts may have been localized (Drobyshev et al. 2008). More broadly,
throughout the deciduous forests of eastern North America, Native American activity
(i.e. burning and agriculture) selected for disturbance-tolerant trees leading to forest com-
munities dominated by those species (Black et al. 2006). Profound impacts, such as
Native American contribution to Holocene migration of forest tree species, are likely to
have occurred at some level (MacDougall 2003), but are challenging to quantify.
Fire is the dominant disturbance in boreal forests and windthrow is the dominant
disturbance in hardwood forest communities (Pastor and Mladenoff 1992; Runkle 1990).
While treefall gaps may occur in the boreal forest (Drobyshev 2001), the deciduous
forests of eastern North America are rarely subject to fire except near the DBE (Runkle
1990) where little difference exists between fire regimes in the two forest types (Bergeron
et al. 2004). The contrasting disturbance regimes of the two biomes may have a signifi-
cant impact on the spatial pattern of the two communities within the DBE. Barras and
Kellman (1998) demonstrated that small-scale micro-site factors (e.g. litter depth, moss
cover) affect the ability of both boreal and deciduous species to establish within the DBE.
Furthermore, as climate change progresses, disturbance regimes (Krawchuk et al. 2009)
and pathogen dynamics (Logan et al. 2003) will likely be altered in both forests.
In some areas, for approximately 4000 years (McLauchlan 2003) temperate deciduous
forest composition in eastern North America has been profoundly altered by Native
American agriculture. Native Americans were practicing swidden agriculture and arbori-
culture in portions of the eastern United States partially accounting for shifts in hardwood
species composition (Black et al. 2006) associated with forest thinning, clearance, and
removal of undesirable trees. Foster et al. (1998) found that tree species distributions in
the deciduous forests of Massachusetts are no longer tied to broad climatic gradients, a
condition attributable to the effects of post-colonization agricultural practices.
The forests of the DBE in North America have been logged extensively for
100 + years (Boucher et al. 2009; Friedman and Reich 2005) and until recently little was
known about how logging activity might alter forest composition of the post-logging for-
ests. The general post-logging pattern for much of the DBE in Ontario and the Great
Lakes states is for shade-intolerant hardwoods to replace boreal species in areas that have
been clear cut (Jackson et al. 2000; Schulte et al. 2007) and previously well-established
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topographic and altitudinal segregation of conifers (lowlands) and hardwoods (upper
slopes) to be erased (Boucher et al. 2009). Further, Schulte et al. (2007) document a
trend towards structural simplification, lower species diversity, and functional diversity of
post-logging forests in the North American DBE compared with the pre-European settle-
ment forest. In some cases, community types (species assemblages) that did not exist
before logging are now extremely common (Friedman and Reich 2005).
Forest decline associated with air pollution is often a concern within the deciduous for-
est, particularly within and near the DBE (Gawel et al. 1996; Jones 2006; McLaughlin
1998; McLaughlin and Percy 1999; Houle et al. 2007; Watmough 2002). The extent of
forest decline occurring within the DBE is most pronounced in Eastern Europe (e.g.
Godbold et al. 1988; Percy and Ferretti 2004) but also in eastern North America (e.g.
Houle et al. 2007; Watmough 2002).
Within both the boreal and deciduous zones, pollution is a key factor affecting forest
change, as exposure to subtle, but long-term, pollutants such as acid deposition, weaken
trees leaving them susceptible to disease, insect outbreaks, and extremes in weather condi-
tions (Drohan et al. 2002; Duchesne et al. 2002; Watmough 2002; Watmough et al.
1998). In North America, studies of sugar maple forests south of the DBE have noted
changes in elemental concentration of Ca, Al, Mg, Mn, and K, over time in the woody
tissue of trees growing in regions experiencing forest decline (e.g. Watmough 2002). In
Europe, spruce seedling root growth has been found to be dramatically reduced in the
presence soil soluble Al (Godbold et al. 1988). In the northeast of North America, fine
root production in both conifer and hardwood stands was found to decline with cation
leaching (particularly Ca) as a result of acid precipitation (Park et al. 2008). The long-
term impacts of pollutants on forested communities within the DBE include forest
decline, because of either nutrient deficiencies or Al trace metal toxicity, and dieback of
species particularly sensitive to changes in soil environmental chemistry or species grow-
ing in marginal soils (Bondietti et al. 1989; Jones 2006; Kogelmann and Sharpe 2006).
DBE dynamics through the holocene
Climate changes over at least the past 20,000 years resulted in massive biome shifts in
terms of geographic location and spatial extent (Amundson and Wright 1979; Davis
1983; Williams et al. 2004, 2007). At the height of the most recent glaciation much of
the area in North America, Europe, and Asia currently occupied by the boreal forest and
the northern deciduous forest was ice covered. In Europe, the extent of the shift was
more pronounced than in North America with only pockets of boreal and deciduous spe-
cies surviving in small populations in protected locations (MacDonald 2003). In North
America the fragmentation, restriction, and marginalization of many of the boreal and
deciduous forest species resulted in ephemeral biomes that do not exist today (Williams
et al. 2004) and the location of what might be considered the DBE is difficult to identify
before 10,000 years BP (Webb 1988). As climates warmed in the post-glacial maximum
period, tree species migrated northward out of their southerly refugia (although see
McLachlan et al. 2005) and shifted into geographic locations and ranges currently associ-
ated with the boreal and deciduous forests (Webb 1988).
Pollen records in North America show that spruce initially colonized the post-glacial
landscapes across the northern United States by approximately 12,500 years BP reaching
the current DBE by 10,000 years BP (Davis 1983; Jacobsen et al. 1987). Maple and birch
species, currently associated with the DBE reached their current northerly limit
approximately 6000 and 7–10,000 years BP, respectively (Davis 1983; Webb et al. 1983).
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By 7000 years BP, within the Great Lakes region of North America, beech had migrated
into the oak hickory forests of Lower Michigan, and hemlock, beech and white pine
formed extensive stands throughout southern Ontario (Davis 1983). Around this period,
white pine was replacing jack pine to the north along with northern white cedar, birch
species and alder (Anderson 1995), forming many of the currently established species asso-
ciations coincident with the mixed forests of the contemporary DBE. Using pollen analy-
sis, Anderson (1995) suggests that regional warming and reduced precipitation with
increased fire activity, between 5000 and 7000 years BP, resulted in a northward shift in
the DBE by 140 km in the Great Lakes region of North America. In Europe, Barnekow
et al. (2008) studied the past 10,000 years of forest change in northeastern Sweden, using
pollen, and found that an initial northward expansion of species associated with warmer
climates (oak, elm, and linden) was mediated by a cooler period, around 3200 years BP,
when spruce, pine, and birch become established beyond the current DBE. Nearer to the
present, during the late Holocene, in response to the Little Ice Age (500 to 150 years
BP) and Medieval Warm Period (1000 to 700 years BP), Hupy and Yansa (2009) docu-
ment northward and southward shifts in the DBE in response to small temperature
changes of 1–2 C.
Pollen evidence suggests that most species migrated rapidly across the continents under
post-glacial climates, at rates of between 100 and 1000 metersyear (Anderson 1995;
Davis 1983; Jacobsen et al. 1987; McLachlan et al. 2005). More recently, the use of
molecular indicators has shown that certain species which currently reside in the DBE,
such as red maple, may have survived during the last glacial maximum in limited popula-
tions within 500 km of the Laurentide ice sheet (McLachlan et al. 2005). Ultimately, the
close proximity of species to the ice sheets means that the post-glacial migration rates
were slower than the pollen records suggest; likely less than 100 meters year. (McLachlan
et al. 2005). A key feature of vegetation migration in relation to climate change was the
individualistic nature of the response (Brubaker 1988; Williams et al. 2004) of the species.
For example, Williams et al. (2004, 2007) postulates that biomes emerge and vanish as
different species shift through space and time, temporally intersecting and forming com-
munities. The implications of this suggestion are that forest biomes, and therefore the
boundaries and ecotones where they meet, will change in terms of their species composi-
tion and abundance depending on individual species tolerances and competitive abilities
given the suite of biotic and abiotic factors to which they are responding. As the climate
of the DBE changes in the near future, the region of the current DBEs around the world
may be moving into periods of novel climates and therefore the present-day coexistence
of species within biomes will become segregated. Changes in climatic conditions may
lead to positive climate change feedbacks as carbon is released due to forest dieback (King
and Neilson 1992).
The DBE and climate change
While spatial resolution from general circulation models (GCM) has improved over the
past two decades, the problem remains that many of the climatic variables responsible for
the location of the DBE are at the synoptic scale (Pastor and Mladenoff 1992) and the
location may also be tied closely to disturbance regimes, neither of which are readily
extracted from GCM output. Further, vegetation responds to climate at a range of spatial
and temporal scales (Tang and Bartlein 2008), so what may be an ecological control at
the global or regional scale (i.e. )40 C average annual minimum temperature isotherm)
will likely differ dramatically from local scale controls on species distributions. Even the
710 Deciduous – boreal forest ecotone
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finest spatial GCM resolutions of 0.5(lat long) are too coarsely scaled for identifying the
local location of the ecotone. Nonetheless, efforts have been made to model biome-level
changes at both regional (e.g. Bachelet and Neilson 2000; Frelich and Reich 2009; Koca
et al. 2006; Prentice et al. 1991; Sykes et al. 1996; Xu et al. 2007) and global scales (e.g.
Cramer et al. 2001; Scholze et al. 2006).
High-latitude vegetation, given its sensitivity to temperature, may exhibit the largest
response to climate change of all the world’s biomes (Pastor and Post 1988; Prentice et al.
1991). For the DBE in Scandinavia (Prentice et al. 1991), East Asia (He et al. 2005;
Zhang et al. 2009) and North America (Frelich and Reich 2009; Solomon 1986) a north-
ward movement in deciduous forest is modeled to occur at the expense of the southern
boreal forest. Pastor and Post (1988) and Xu et al. (2007) found the nature of the forest
change to be conditional on soil moisture, with boreal forests being competitively
replaced only if soil moisture was adequate. However, these transitions may take several
centuries given lags due to ecological processes (Solomon 1986; Sykes and Prentice
1996). Given species-specific responses to past climate changes, the nature of the north-
ward expansion of the deciduous forest is unlikely to be simple because it will be medi-
ated by climatic conditions, changes in competitive interactions, invasive species,
herbivore distributions, and disturbance regimes, as well as exogenous factors such as acid
precipitation, CO
fertilization, and altered disease and pathogen dynamics (Bergeron
et al. 2004; Frelich and Reich 2009; Loehle 2000; Price and Apps 1996; Solomon 1986;
Sykes and Prentice 1996; Sykes et al. 1996). A general concern is that plant migration
rates may be inadequate to track rapid anthropogenic climate change potentially leading
to depauperate forest ecotone communities (Solomon 1986; Solomon and Kirilenko
1997), a probable scenario if migration rates described by McLachlan et al. (2005) are
Nearly all the research on potential climate change driven dynamics of the DBE is
based on forest simulations. However, a few field-based empirical studies have been
conducted, although at fine spatial scales and generally not at the DBE. Some studies
(e.g. Bronson et al. 2009; Edwards and Norby 1999; Farnsworth et al. 1995; Gunderson
et al. 2000; Norby et al. 2003; Wan et al. 2004) have manipulated air and or soil temper-
ature in field environments in one of the two forest types to assess the impact that
warmer climates might have on plant communities. Similarly, field experiments enhanc-
ing CO
levels (free-air CO
experiments: FACE experiments) have been conducted in
both boreal and deciduous forests (Nowak et al. 2004), or in boreal forests containing
deciduous species (Rasmussen et al. 2002), but none have been conducted at the DBE.
Research at the ecotone is somewhat limited. Goldblum and Rigg (2005) employed past
growth rates derived from tree rings to predict future growth rates of the dominant tree
species at the ecotone in Ontario, and Goldblum and Rigg (2002) described the stand
structure and demography at the ecotone, also in Ontario, but little else has been
published. Given the substantial body of research focused on the impact of climate
change on forest communities around the world, there continues to be a need to pursue
field-based research on climate change impacts in ecotone areas, including the DBE.
Short Biographies
David Goldblum’s research focuses on the role of disturbances on natural plant communi-
ties, most recently focusing on the impact of anthropogenic climate change; he has
authored or co-authored papers in these areas for Dendrochronologia, Physical Geography,
Canadian Journal of Forest Research, Bulletin of the Torrey Botanical Society, Journal of Vegetation
Deciduous – boreal forest ecotone 711
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Journal Compilation ª2010 Blackwell Publishing Ltd
Science, Journal of Biogeography, and Plant Ecology. Current research involves a field experi-
ment simulating climate change in the forest at the deciduous forest – boreal forest ecotone
in Ontario, Canada and studying the dynamics of herbaceous plants in the understories of
boreal and deciduous forests. Before coming to Northern Illinois University, where he
presently teaches, Goldblum taught at the University of Wisconsin – Whitewater, Northern
Michigan University, and University of Melbourne. He holds a BS in Geography from
UCLA and a MS and PhD in Geography from the University of Colorado.
Lesley Rigg’s research is currently focused on the population dynamics of the boreal
forest deciduous forest boundary in Lake Superior (Ontario, Canada) and potential spe-
cies range shifts associated with climate change. Ongoing studies include the population
ecology of tree species growing in mixed angiosperm conifer communities associated
with ultramafic soils, in New Caledonia (South Pacific) and the regeneration status of
oak hickory woodlands in Northern Illinois. She has authored or co-authored papers in
these areas for Biotropica, Physical Geography, Austral Ecology and Canadian Journal of Forest
Research. Rigg currently is the Chair of the Biogeography Specialty Group of the Associa-
tion of American Geographers and is active in the area of women in science. Rigg holds
a BA in Physical Geography from York University in Toronto, Canada, and a master’s
degree in Geography from the University of Colorado, Boulder. She completed her PhD
work in Australia at the University of Melbourne before joining the faculty at Northern
Illinois University where she currently holds the position of Associate Professor.
*Corresponding address: D. Goldblum, Department of Geography, Northern Illinois University, DeKalb, IL
60115, USA. E-mail:
Amundson, D. C. and Wright, H. E. Jr (1979). Forest changes in Minnesota at the end of the Pleistocene. Ecological
Monographs 49, pp. 1–16.
Anderson, T.W. (1995). Forest changes in the Great Lakes region at 5-7 ka BP. Geographie
´Physique et Quaternaire
49, pp. 99–116.
Arris, L. L. and Eagleson, P. S. (1989). Evidence of a physiological basis for the boreal-deciduous forest ecotone in
North America. Vegetatio 82, pp. 55–58.
Arris, L.L. and Eagleson, P. S. (1994). A water use model for locating the boreal deciduous forest ecotone in east-
ern North America. Water Resources Research 30, pp. 1–9.
Axelsson, A.-L., O
¨stlund, L. and Hellberg, E. (2002). Changes in mixed deciduous forests of boreal Sweden 1866–
1999 based on interpretation of historical records. Landscape Ecology 17, pp. 403–418.
Bachelet, D. and Neilson, R. P. (2000). Biome redistribution under climate change. USDA Forest Service Gen.
Tech. Rep. RMRS-GTR-59.
Barnekow, L., Brage
´e, P., Hammarlund, D. and St. Amour, N. (2008). Boreal forest dynamics in north-eastern
Sweden during the last 10,000 years based on pollen analysis. Vegetation History and Archaeobotany 17, pp. 687–
Barnosky, A. D., et al. (2004). Assessing the Causes of Late Pleistocene Extinctions on the Continents. Science 306,
pp. 70–75.
Barras, N. and Kellman, M. (1998). The supply of regeneration micro-sites and segregation of tree species in a
hardwood boreal forest transition zone. Journal of Biogeography 25, pp. 871–881.
Beckage, B., et al. (2008). A rapid upward shift of a forest ecotone during 40 years of warming in the Green
Mountains of Vermont. Proceedings of the National Academy of Sciences 105, pp. 4197–4202.
Bergeron, Y., Gauthier, S., Flannigan, M. and Kafka, V. (2004). Fire regimes at the transition between mixedwood
and coniferous boreal forest in northwestern Quebec. Ecology 85, pp. 1916–1932.
Biondietti, E. A., Baes, C.F. III and McLaughlin, B. (1989). Radial trends in cation ratios in tree rings as indicators
of the impact of atmospheric deposition on forests. Canadian Journal of Forest Research 19, pp. 586–894.
Black, B. A., Ruffner, C. M. and Abrams, M. D. (2006). Native American influences on the forest composition of
the Allegheny Plateau, northwest Pennsylvania. Canadian Journal of Forest Research 36, pp. 1266–1275.
712 Deciduous – boreal forest ecotone
ª2010 The Authors Geography Compass 4/7 (2010): 701–717, 10.1111/j.1749-8198.2010.00342.x
Journal Compilation ª2010 Blackwell Publishing Ltd
Boucher, Y., Arseneault, D., Sirois, L. and Blais, L. (2009). Logging pattern and landscape changes over the last
century at the boreal and deciduous forest transition in Eastern Canada. Landscape Ecology 24, pp. 171–184.
Breckle, S.-W. (2002). Walter’s vegetation of the earth: the ecological systems of the geo-biosphere. Springer-Verlag: Berlin.
Bronson, D. R., Gower, S. T., Tanner, M. and Van Herk, I. (2009). Effect of ecosystem warming on boreal black
spruce bud burst and shoot growth. Global Change Biology 15, pp. 1534–1543.
Brubaker, L. B. (1988). Vegetation history and anticipating future vegetation change. In: Agee, J. K. and Johnson, D.
R., (eds.) Ecosystem management for parks and wilderness. Seattle, WA: University of Washington Press, pp. 41–61.
Clark, J. S. and Royall, P. D. (1995). Transformation of a northern hardwood forest by aboriginal (Iroquois) fire:
charcoal evidence from Crawford Lake, Ontario, Canada. The Holocene 5, pp. 1–9.
Clements, F. E. (1907). Plant physiology and ecology. New York: University of Washington Press.
Colombaroli, D., et al. (2008). Fire-vegetation interactions during the Mesolithic-Neolithic transition at Lago
dell’Accesa, Tuscany, Italy. The Holocene 18, pp. 679–692.
Cramer, W., et al. (2001). Global response of terrestrial ecosystem structure and function to CO
and climate
change: results from six dynamic global vegetation models. Global Change Biology 7, pp. 357–373.
Davis, M. B. (1983). Holocene vegetational history of the eastern United States. In: Wright, H. E. Jr, (ed.) Late-
quaternary environments of the United States. Minneapolis: University of Minnesota Press, pp. 166–181.
Delcourt, P. A. and Delcourt, H. R. (1987). Late-quaternary dynamics of temperate forests: applications of paleo-
ecology to issues of global environmental change. Quaternary Science Reviews 6, pp. 129–146.
Demers, J. D., Lee, T. D. and Barrett, J. P. (1998). Substrate type and the distribution of sugar maple at its eleva-
tional limit in the White Mountains, New Hampshire. Canadian Journal of Forest Research 28, pp. 494–498.
Donlan, C. J., et al. (2006). Pleistocene rewilding: an optimistic agenda for twenty-first century conservation. The
American Naturalist 168, pp. 660–681.
Drobyshev, I. V. (2001). Effect of natural disturbances on the abundance of Norway spruce (Picea abies (L.)
Karst.) regeneration in nemoral forests of the southern boreal zone. Forest Ecology and Management 140, pp.
Drobyshev, I., et al. (2008). Pre- and post-European settlement fire history of red pine dominated forest eco-
systems of Seney National Wildlife Refuge, Upper Michigan. Canadian Journal of Forest Research 38, pp.
Drohan, P. J., Stout, S. L. and Petersen, G. W. (2002). Sugar maple (Acer saccharum Marsh.) decline during 1979–
1989 in northern Pennsylvania. Forest Ecology and Management 170, pp. 1–17.
Duchesne, L., Ouimet, R. and Houle, D. (2002). Basal area growth of sugar maple in relation to acid deposition,
stand health, and soil nutrients. Journal of Environmental Quality 31, pp. 1676–1683.
Edwards, N. T. and Norby, R. J. (1999). Below-ground respiratory processes of sugar maple and red maple saplings
to atmospheric CO
enrichment and elevated air temperature. Plant and Soil 206, pp. 85–97.
Elgersma, A. M. and Dhillion, S. S. (2002). Geographical variability of relationships between forest commu-
nities and soil nutrients along a temperature-fertility gradient in Norway. Forest Ecology and Management 158,
pp. 155–168
Farnsworth, E. J., Nu
˜ez-Farfan, J., Careaga, S. A. and Bazzaz, F. A. (1995). Phenology and growth of three tem-
perate forest life forms in response to artificial soil warming. Journal of Ecology 83, pp. 967–977.
Fonara, D. A. and Tilman, D. (2008). Plant functional composition influences rates of soil carbon and nitrogen
accumulation. Journal of Ecology 96, pp. 314–322.
Forster, P. (2004). Ice ages and the mitochondrial DNA chronology of human dispersals: a review. Philosophical
Transactions of the Royal Society of London B 359, pp. 255–264.
Fortin, M.-J., et al. (2000). Issues related to the detection of boundaries. Landscape Ecology 15, pp. 453–466.
Foster, D. R., Motzkin, G. and Slater, B. (1998). Land-use history as long-term broad-scale disturbance: regional
forest dynamics in central New England. Ecosystems 1, pp. 96–119.
Frelich, L. E. and Reich, P. B. (2009). Wilderness conservation in an era of global warming and invasive species: a
case study from Minnesota’s Boundary Waters Canoe Area Wilderness. Natural Areas Journal 29, pp. 385–393.
Friedman, S. K. and Reich, P. B. (2005). Regional legacies of logging: departure from presettlement forest condi-
tions in northern Minnesota. Ecological Applications 15, pp. 726–744.
Gauthier, S., DeGrandpre
´, L. and Bergeron, Y. (2000). Differences in forest compostion in two boreal forest ecore-
gions of Quebec. Journal of Vegetation Science 11, pp. 781–790.
Gawel, J. E., Ahner, B. A., Friedland, A. J. and Morel, F. M. M. (1996). Role for heavy metals in forest decline
indicated by phytochelatin measurements. Nature 381, pp. 64–65.
Godbold, D. L., Fritz, E. and Hu
¨ttermann, A. (1988). Aluminum toxicity and forest decline. Proceedings of the
National Academy of Sciences 85, pp. 3888–3892.
Goldblum, D. and Rigg, L. S. (2002). Age structure and regeneration dynamics of sugar maple at the decidu-
ous boreal forest ecotone, Ontario, Canada. Physical Geography 23, pp. 115–129.
Goldblum, D. and Rigg, L. S. (2005). Tree growth response to climate change at the deciduous-boreal forest
ecotone, Ontario, Canada. Canadian Journal of Forest Research 35, pp. 2709–2718.
Gosz, J. R. (1993). Ecotone hierarchies. Ecological Applications 3, pp. 370–376.
Deciduous – boreal forest ecotone 713
ª2010 The Authors Geography Compass 4/7 (2010): 701–717, 10.1111/j.1749-8198.2010.00342.x
Journal Compilation ª2010 Blackwell Publishing Ltd
Gunderson, C. A., Norby, R. J. and Wullschleger, S. D. (2000). Acclimation of photosynthesis and respiration to
simulated climate warming in northern and southern populations of Acer saccharum: laboratory and field evidence.
Tree Physiology 20, pp. 87–96.
Hayes, M., Moody, A., White, P. S. and Costanza, J. L. (2007). The influence of logging and topography on the
distribution of spruce-fir forests near their Southern limits in Great Smoky Mountains National Park, USA. Plant
Ecology 189, pp. 59–70.
He, H. S., Hao, Z., Mladenoff, D. J. and Shao, G. (2005). Simulating forest ecosystem response to climate warming
incorporating spatial effects in north-eastern China. Journal of Biogeography 32, pp. 2043–2056.
Heinselman, M. L. (1973). Fire in the virgin forests of the Boundary Waters Canoe Area, Minnesota. Quaternary
Research 3, pp. 329–382.
Hewitt, G. (2000). The genetic legacy of the Quaternary ice ages. Nature 405, pp. 907–913.
Hogg, E. H. (1994). Climate and the southern limit of the western Canadian boreal forest. Canadian Journal of Forest
Research 24, pp. 1835–1845.
Hou, H.-Y. (1983). Vegetation of China with reference to its geographical distribution. Annals of the Missouri
Botanical Garden 70, pp. 509–549.
Houle, D., Tremblay, S. and Ouimet, R. (2007). Foliar and wood chemistry of sugar maple along a gradient of soil
acidity and stand health. Plant Soil 300, pp. 173–183.
Hupy, C. M. and Yansa, C. H. (2009). Late Holocene vegetation history of the forest tension zone in central lower
Michigan, USA. Physical Geography 30, pp. 205–235.
Jackson, S. M., Pinto, F., Malcolm, J. R. and Wilson, E. R. (2000). A comparison of pre-European settlement
(1857) and current (1981–1995) forest composition in central Ontario. Canadian Journal of Forest Research 30, pp.
Jacobsen, G. L. Jr, Webb, T. III and Grimm, E. C. (1987). Patterns and rates of vegetation change during the
deglaciation of eastern North America. In: Ruddiman, W. F. and Wright, H. E. Jr, (eds.) North America and
adjacent oceans during the last deglaciation. Boulder, Colorado: Geological Society of North America, pp. 277–
Jenny, H. (1994). Factors of soil formation: a system of quantitative pedology. New York: Dover Publications.
Jones, L. S. (2006). Use of dendroanalysis to study environmental change in three unique geological ecological settings. Ph.D.
Thesis, Department of Geology and Environmental Geosciences. Northern Illinois University.
Kellman, M. (2004). Sugar maple (Acer saccharum Marsh.) establishment in boreal forest: results of transplantation
experiment. Journal of Biogeography 31, pp. 1515–1522.
Kent, M., Gill, W. J., Weaver, R. E. and Armitage, R. P. (1997). Landscape and plant community boundaries in
biogeography. Progress in Physical Geography 21, pp. 315–353.
King, G. A. and Neilson, R. P. (1992). The transient response of vegetation to climate change: a potential source
of CO
to the atmosphere. Water, Air, and Soil Pollution 64, pp. 365–383.
Koca, D., Smith, B. and Sykes, M. T. (2006). Modelling regional climate change effects on potential natural ecosys-
tems in Sweden. Climatic Change 78, pp. 381–406.
Krawchuk, M. A., Cumming, S. G. and Flannigan, M. D. (2009). Predicted changes in fire weather suggest
increases in lightning fire initiation and future area burned in the mixedwood boreal forest. Climatic Change 92,
pp. 83–97.
Kullman, L. (1995). Holocene tree-limit and climate history from the Scandes Mountains, Sweden. Ecology 76, pp.
Kupfer, J. A. and Cairns, D. M. (1996). The suitability of montane ecotones as indicators of global climatic change.
Progress in Physical Geography 20, pp. 253–272.
Lee, T. D., Barrett, J. P. and Hartman, B. (2005). Elevation, substrate, and the potential for climate-induced tree
migration in the White Mountains, New Hampshire, USA. Forest Ecology and Management 212, pp. 75–91.
Lenihan, J. M. and Neilson, R. P. (1993). A rule-based vegetation formation model for Canada. Journal of Biogeogra-
phy 20, pp. 615–628.
Liu, Q.-J., Kondoh, A. and Takeuchi, N. (1998). Study of changes in life zone distribution in north-east China by
climate-vegetation classification. Ecological Research 13, pp. 355–365.
Livingston, B. E. (1903). The distribution of the upland societies of Kent County, Michigan. Botanical Gazette 35,
pp. 36–55.
Loehle, C. (2000). Forest ecotone response to climate change: sensitivity to temperature response functional forms.
Canadian Journal of Forest Research 30, pp. 1632–1645.
Logan, J. A., Re
`re, J. and Powell, J. A. (2003). Assessing the impacts of global warming on forest pest dynam-
ics. Frontiers in Ecology and the Environment 1, pp. 130–137.
Loope, W. L. and Anderton, J. B. (1988). Human vs. lightning ignition of presettlement surface fires in coastal pine
forests of the upper Great Lakes. The American Midland Naturalist 140, pp. 206–218.
van der Maarel, E. (1990). Ecotones and ecoclines are different. Journal of Vegetation Science 1, pp. 135–138.
MacDonald, G. (2003). Biogeography: introduction to space, time and life. New York: John Wiley & Sons, Inc.
714 Deciduous – boreal forest ecotone
ª2010 The Authors Geography Compass 4/7 (2010): 701–717, 10.1111/j.1749-8198.2010.00342.x
Journal Compilation ª2010 Blackwell Publishing Ltd
MacDougall, A. (2003). Did native Americans influence the northward migration of plants during the Holocene?
Journal of Biogeography 30, 633–647.
Maycock, P. F. and Curtis, J. T (1960). The phytosociology of boreal conifer-hardwood forests of the Great Lakes
region. Ecological Monographs 30, pp. 1–36.
McLachlan, J. S., Clark, J. S. and Manos, P. S. (2005). Molecular indicators of tree migration capacity under rapid
climate change. Ecological Society of America 86, pp. 2088–2098.
McLauchlan, K. (2003). Plant cultivation and forest clearance by prehistoric North Americans: pollen evidence from
Fort Ancient, Ohio, USA. The Holocene 13, pp. 557–266.
McLaughlin, D. (1998). A decade of forest tree monitoring in Canada: evidence of air pollution effects. Environmen-
tal Reviews 6, pp. 151–171.
McLaughlin, S. and Percy, K. (1999). Forest health in North America: some perspectives on actual and potential
roles of climate and air pollution. Water, Air and Soil Pollution 116, pp. 151–197.
Messaoud, Y., Bergeron, Y. and Leduc, A. (2007). Ecological factors explaining the location of the boundary
between the mixedwood and coniferous bioclimatic zones in the boreal biome of eastern North America. Global
Ecology and Biogeography 16, pp. 90–102.
Miettinen, J., Gro
¨nlund, E., Simola, H. and Huttunen, P. (2002). Palaeolimnology of Lake Pieni-Kuuppalanlampi
(Kurkijoki, Karelian Republic, Russia): isolation history, lake ecosystem development and long-term agricultural
impact. Journal of Paleolimnology 27, pp. 29–44.
Naiman, R. J. and De
´camps, H. (eds). (1990). The ecology and management of aquatic-terrestrial ecotones. Paris: The
Parthenon Publishing Group.
Neilson, R. P. (1993). Transient ecotone response to climatic change: some conceptual and modelling approaches.
Ecological Applications 3, pp. 385–395.
Neilson, R. P. (1995). A model for predicting continental-scale vegetation distribution and water balance. Ecological
Applications 5, pp. 362–385.
Noble, I. R. (1993). A model of the responses of ecotones to climate change. Ecological Applications 3, pp. 396–403.
Norby, R. J., Hartz-Rubin, J. S. and Verbrugge, M. J. (2003). Phenological responses in maple to experimental
atmospheric warming and CO
enrichment. Global Change Biology 9, pp. 1792–1801.
Nowak, R. S., Ellsworth, D. S. and Smith, S. D. (2004). Functional responses of plants to elevated atmospheric
– do photosynthetic and productivity data from FACE experiments support early predictions? New Phytolo-
gist 162, pp. 253–280.
Ohsawa, M. (1984). Differentiation of vegetation zones and species strategies in the subalpine region of Mt. Fuji.
Vegetatio 57, pp. 15–52.
Ohsawa, M. (1990). An interpretation of latitudinal patterns of forest limits in south and east Asian mountains. Jour-
nal of Ecology 78, pp. 326–339.
Park, B. B., et al. (2008). Fine root dynamics and forest production across a calcium gradient in Northern Hard-
wood and conifer ecosystems. Ecosystems 11, pp. 325–341.
Parmesan, C. (2006). Ecological and evolutionary response to recent climate change. Annual Review of Ecology, Evo-
lution, and Systematics 37, pp. 637–669.
Parmesan, C., et al. (2005). Empirical perspectives on species borders: from traditional biogeography to global
change. Oikos 108, pp. 58–75.
Pastor, J. and Mladenoff, D. J. (1992). The southern boreal-northern hardwood forest border. In: Shugart, H. H.,
Leemans, R. and Bonan, G. B., (eds.) A systems analysis of the global boreal forest. Cambridge: Cambridge Univer-
sity Press, pp. 216–240.
Pastor, J. and Post, W. M. (1988). Response of northern forests to CO
-induced climate change. Nature 334, pp.
Percy, K. E. and Ferretti, M. (2004). Air pollution and forest health: toward new monitoring concepts. Environmen-
tal Pollution 130, pp. 113–126.
Prentice, I. C., Sykes, M. T. and Cramer, W. (1991). The possible dynamic response of northern forests to global
warming. Global Ecology and Biogeography Letters 1, pp. 129–135.
Prentice, I. C., et al. (1992). A global biome model based on plant physiology and dominance, soil properties and
climate. Journal of Biogeography 19, pp. 117–134.
Price, D. T. and Apps, M. J. (1996). Boreal forest responses to climate-change scenarios along an ecoclimatic tran-
sect in central Canada. Climatic Change 34, pp. 179–190.
Rasmussen, L., Beier, C. and Bergstedt, A. (2002). Experimental manipulations of old pine forest ecosystems to
predict the potential tree growth effects of increased CO2 and temperature in a future climate. Forest Ecology and
Management 158, pp. 179–188.
Reich, P. B. and Oleksyn, K. (2004). Global patterns of plant leaf N and P in relation to temperature and latitude.
Proceedings of the National Academy of Sciences 101, pp. 11001–11006.
Reich, P. B., Kloeppel, B. D. and Ellsworth, D. S. (1995). Different photosynthesis-nitrogen relations in deciduous
hardwood and evergreen coniferous tree species. Oecologia 104, pp. 24–30.
Deciduous – boreal forest ecotone 715
ª2010 The Authors Geography Compass 4/7 (2010): 701–717, 10.1111/j.1749-8198.2010.00342.x
Journal Compilation ª2010 Blackwell Publishing Ltd
Ries, L., Fletcher, R. J. Jr, Battin, J. and Sisk, T. D. (2004). Ecological responses to habitat edges: mechanisms,
models, and variability explained. Annual Review of Ecology, Evolution, and Systematics 35, pp. 491–522.
Risser, P. G. (1995). The status of the science examining ecotones. BioScience 45, pp. 318–325.
Runkle, J. R. (1990). Gap dynamics in an Ohio Acer-Fagus forest and speculations on the geography of disturbance.
Canadian Journal of Forest Research 20, pp. 632–641.
Sakai, A. (1975). Freezing resistance of evergreen and deciduous broad-leaf trees in Japan with special reference to
their distributions. Japanese Journal of Ecology 25, pp. 101–111.
Sakai, A. and Weiser, C. J. (1973). Freezing resistance of trees in North America with reference to tree regions.
Ecology 54, pp. 118–126.
Scholze, M., Knorr, W., Arnell, N. W. and Prentice, I. C. (2006). A climate-change risk analysis for world ecosys-
tems. Proceedings of the National Academy of Sciences 103, pp. 13116–13120.
Schulte, L. A., et al. (2007). Homogenization of northern U.S. Great Lakes forests due to land use. Landscape Ecol-
ogy 22, pp. 1089–1103.
Scott, G. A. J. (1995). Canada’s vegetation: a world perspective. Montreal: McGill-Queen’s University Press.
Solomon, A. M. (1986). Transient response of forests to CO2-induced climate change: simulation modeling experi-
ments in eastern North America. Oecologia 68, pp. 567–579.
Solomon, A. M. and Kirilenko, A. P. (1997). Climate change and terrestrial biomass: what if trees do not migrate!
Global Ecology and Biogeography Letters 6, pp. 139–148.
Sowell, J. B. (1985). A predictive model relating North American plant formations and climate. Vegetatio 60, pp.
Stearns, F. W. (1997). History of the lake states forests: natural and human impacts. In: Vasievich, J. M. and Webster, H.
H. (eds) Lake states regional forest resources assessment: Technical papers Gen. Tech Rep. NC-189. St Paul, MN:
USDA, Forest Service, North Central Forest Experiment Station, pp. 8–29.
Stephenson, N. L. (1998). Actual evapotranspiration and deficit: biologically meaningful correlates of vegetation
distribution across spatial scales. Journal of Biogeography 25, pp. 855–870.
Sukachev, V. N. (1928). Principles of classification of the spruce communities of European Russia. Journal of Ecology
16, pp. 1–18.
Sykes, M. T. and Prentice, I. C. (1996). Climate change, tree species distributions and forest dynamics: a case study
in the mixed conifer northern hardwoods zone of northern Europe. Climatic Change 34, pp. 161–177.
Sykes, M. T., Prentice, I. C. and Cramer, W. (1996). A bioclimatic model for the potential distributions of north
European tree species under present and future climates. Journal of Biogeography 23, pp. 203–233.
Tang, C. Q. and Ohsawa, M. (1997). Zonal transition of evergreen, deciduous, and coniferous forests along the
altitudinal gradient on a humid subtropical mountain, Mt. Emei, Sichuan, China. Plant Ecology 133, pp. 63–78.
Tang, C. Q. and Ohsawa, M. (2002). Consistence mechanisms of evergreen, deciduous, and coniferous trees in a
mid montane forest on Mt. Emei, Sichuan, China. Plant Ecology 161, pp. 215–230.
Tang, G. and Bartlein, P. J. (2008). Simulating the climatic effects on vegetation: approaches, issues and challenges.
Progress in Physical Geography 32, pp. 543–556.
Thornthwaite, C. W. (1948). An approach toward a rational classification of climate. Geographic Review 38, pp. 55–
Vallee, S. and Payette, S. (2004). Contrasted growth of black spruce (Picea mariana) forest trees at treeline associated
with climate change over the last 400 years. Arctic, Antarctic, and Alpine Research 36, pp. 400–406.
Wan, S. Q., et al. (2004). CO
enrichment and warming of the atmosphere enhance both productivity and mortal-
ity of maple tree fine roots. New Phytologist 162, pp. 437–446.
Watkins, L. (2006). Forest resources of Ontario 2006: state of the forest report 2006. Ontario: Ontario Ministry of Natu-
ral Resources.
Watmough, S. A. (2002). A dendrochemical survey of sugar maple (Acer saccharum Marsh) in south-central Ontario,
Canada. Water, Air, and Soil Pollution 136, pp. 165–187.
Watmough, S. A., Hutchinson, T. C. and Sager, E. P. S. (1998). Changes in tree ring chemistry in sugar maple
(Acer saccharum) along an urban-rural gradient in southern Ontario. Environmental Pollution 101, pp. 381–390.
Webb, T. III (1988). Eastern North America. In: Huntley, B. and Webb, T. III, (eds.) Vegetation history. Dordrecht:
Kluwer Academic Publishers, pp. 385–414.
Webb, T. III, Cushing, E. J. and Wright, H. E. Jr (1983). Holocene changes in the vegetation of the Midwest. In:
Wright, H. E. Jr, (ed.) Late-quaternary environments of the United States. Minneapolis: University of Minnesota
Press, pp. 142–165.
Weishampel, J. F., Knox, R. G. and Levine, E. R. (1999). Soil saturation effects on forest dynamics: scaling across
a southern boreal northern hardwood landscape. Landscape Ecology 14, pp. 121–135.
Williams, J. W., et al. (2004). Late-quaternary vegetation dynamics in North America: scaling from taxa to biomes.
Ecological Monographs 74, pp. 309–334.
Williams, J. W., Jackson, S. T. and Kutzbach, J. E. (2007). Projected distributions of novel and disappearing cli-
mates by 2100 AD. Proceedings of the National Academy of Sciences 104, pp. 5738–5742.
716 Deciduous – boreal forest ecotone
ª2010 The Authors Geography Compass 4/7 (2010): 701–717, 10.1111/j.1749-8198.2010.00342.x
Journal Compilation ª2010 Blackwell Publishing Ltd
Wilson, J. B. and Agnew, A. D. Q. (1992). Positive feedback switches in plant communities. Advances in Ecological
Research 23, pp. 263–336.
Woodward, S. L. (2003). Biomes of the Earth: terrestrial, aquatic, and human-dominated. Westport, Connecticut: Green-
wood Press.
Woodward, F. I. and Williams, B. G. (1987). Climate and plant distribution at global and local scales. Vegetatio 69,
pp. 189–197.
Xu, C., Gertner, G. Z. and Scheller, R. M. (2007). Potential effects of interaction between CO
and temperature
on forest landscape response to global warming. Global Change Biology 13, pp. 1469–1483.
Yarrow, M. M. and Marin, V. H. (2007). Toward conceptual cohesiveness: a historical analysis of the theory and
utility of ecological boundaries and transition zones. Ecosystems 10, pp. 462–476.
Yoshino, M. M. (1978). Altitudinal vegetation belts of Japan with special reference to climatic conditions. Arctic and
Alpine Research 10, pp. 449–456.
Zhang, N., Shugart, H. H. and Yan, X. (2009). Simulating the effects of climate changes on Eastern Eurasia forests.
Climatic Change 95, pp. 341–361.
ª2010 The Authors Geography Compass 4/7 (2010): 701–717, 10.1111/j.1749-8198.2010.00342.x
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... The northern border of the projected invasive range was observed as roughly following the deciduous-boreal forest ecotone [143] in the East and the prairie-forest biome border [144] in the Alberta-Saskatchewan region in the West. The rapid warming in the boreal ...
... The northern border of the projected invasive range was observed as roughly following the deciduous-boreal forest ecotone [143] in the East and the prairie-forest biome border [144] in the Alberta-Saskatchewan region in the West. The rapid warming in the boreal forests (approximately twice as fast as the global average [145]) and the projected northward shifts of warmer climate zones [146], e.g., the climatic shifts from prairies to the boreal forests [147,148], are expected to cause significant disturbances that can affect individual species and ecosystems and can lead to biome-level changes [149]. ...
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Biological invasions are a major component of global environmental change with severe ecological and economic consequences. Since eradicating biological invaders is costly and even futile in many cases, predicting the areas under risk to take preventive measures is crucial. Impatiens glandulifera is a very aggressive and prolific invasive species and has been expanding its invasive range all across the Northern hemisphere, primarily in Europe. Although it is currently spread in the east and west of North America (in Canada and USA), studies on its fate under climate change are quite limited compared to the vast literature in Europe. Hybrid models, which integrate multiple modeling approaches, are promising tools for making projections to identify the areas under invasion risk. We developed a hybrid and spatially explicit framework by utilizing MaxEnt, one of the most preferred species distribution modeling (SDM) methods, and we developed an agent-based model (ABM) with the statistical language R. We projected the I. glandulifera invasion in North America, for the 2020–2050 period, under the RCP 4.5 scenario. Our results showed a predominant northward progression of the invasive range alongside an aggressive expansion in both currently invaded areas and interior regions. Our projections will provide valuable insights for risk assessment before the potentially irreversible outcomes emerge, considering the severity of the current state of the invasion in Europe.
... This allows for transitional zones at the temperate-boreal biome interface due to interspecific competitive limitations. In these northern temperate landscapes, species living at the edge of their range coexist in highly unstable and competitive ecosystems in which small variations in climatic or microclimatic conditions can result in changes in forest structure and composition (Goldblum and Rigg, 2010). Northern temperate landscapes are therefore affected by small-gap dynamics that creates irregular structures within stands. ...
... Anthropogenic disturbances such as harvesting may disrupt the natural carbon cycling of an ecosystem and thus the forest carbon budget, i.e. the balance between carbon sequestration and emission processes (Malhi et al., 1999;Goldblum and Rigg, 2010). Moreover, the compounding effects of wood harvesting in a changing climate can lead to significant variations in the forest landscape characteristics that may ultimately alter ecosystem capacity to store carbon (Steenberg et al., 2013;Brice et al., 2019). ...
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Background Forest based climate mitigation emerged as a key component of the Paris Agreement, and thus requires robust science to reduce uncertainties related to such strategies. The aim of this study was to assess and compare the cumulative effects on carbon dynamics of forest management and climate change on boreal and northern temperate forest sector in eastern Canada for the 2020-2100 period. Methods We used the spatially explicit forest landscape model LANDIS-II and its extension Forest Carbon Succession, in conjunction with the Carbon Budget Model for Harvested Wood Products framework. We simulated the dynamics of forest composition and carbon flows from forest ecosystems to wood products and their substitution effect on markets under increasing climate forcing, according to a tonne-year approach. Simulations were conducted for a series of forest management scenarios based on realistic practices principally by clearcut in the boreal territory and continuous-cover forestry in the northern temperate one. These scenarios included: i) a business-as-usual scenario (BaU), representing the current management strategy, ii) increased harvesting by 6.3 to 13.9%, iii) increased conservation (i.e. reduced harvesting by by 11.1 to 49.8%), iiii) and a scenario representing the natural evolution of the forest landscape (i.e. without any management activity). Results Our study revealed that increasing harvesting levels had contrasting effects on the mitigation potential in northern temperate (enhance net sequestration) and boreal forest sector (enhance net emissions) in comparison to the BaU from 2040 onwards, regardless of the future climate. Carbon storage in wood products and the substitution effect were not sufficient to offset carbon emissions from ecosystems. Moreover, climate change had a strong impact on the capacity of both landscapes to act as carbon sinks. Northern temperate landscapes became a net source of carbon over time due to their greater vulnerability to climate change than boreal landscapes. Conclusions Our study highlights the need to consider the initial landscape characteristics in simulations to maximize the mitigation potential of alternative forest management strategies. The optimal management solution can be very different according to the characteristics of forest ecosystems. This opens the possibility of optimizing management for specific forest stands, with the objective of maximizing the mitigation potential of a given landscape.
... Other climate drivers include moisture stress, warmer temperatures, increased insect infestations, N deposition, and CO 2 fertilization (Kint et al., 2012;Silva et al., 2010). Drier, warmer boreal forests will store less carbon due to moisture stress (Ma et al., 2012), becoming a net source of greenhouse gasses (Flannigan et al., 2000(Flannigan et al., , 2009Kurz et al., 2013), despite increased productivity in northern open taiga forests Goldblum & Rigg., 2010). A warming climate may result in release of the huge carbon store in frozen boreal peat soils (Schaefer et al., 2011). ...
... The forests of our study region are located at the northern edge of the deciduous forest ecozone in Northeastern Ontario (Goldblum and Rigg, 2010). We first derived a list of 130 candidate species for restoration based on plant taxa found in the Sudbury region and species planted by Sudbury's Regreening Program (City of Greater Sudbury, 2016, 2017SARA Group, 2009). ...
We present an applied model that helps restoration practitioners select an ideal mix of species to plant in order to meet their restoration objectives. The model generates virtual plant communities designed to optimize the delivery of multiple ecosystem functions. We used an optimization approach to find the most cost-effective combinations of species to plant to optimize the delivery of four ecosystem functions: rapid establishment of vegetation cover, soil building, biological soil health and resistance to invasion. We used trait-function relationships to characterize species' effects on ecosystem functions. This model accounts for key operational constraints selected by the user, including budget, the number of species to plant, and which functions to consider. The user can also decide whether or not to maximize the functional diversity of the species mix to increase its resilience to global environmental change. To demonstrate the practicality of this approach, we derived optimal species mixtures for the restoration of forests damaged by Cu-Ni smelters in the City of Greater Sudbury (Ontario, Canada). The species mixtures generated by the model varied according to which functions and operational constraints were selected. Results show that the species mixtures that were the most effective at delivering multiple functions were also cost-effective, but were less functionally diverse. This tool provides restoration practitioners with cost-effective restoration strategies for managing the recovery of multi-faceted socio-economic and environmental values in disturbed landscapes.
... Natural and human disturbances can be defined as relatively discrete events in time that disrupt the ecosystem and cause pronounced changes in resource availability or the physical environment (White & Pickett 1985, Attiwill 1994, and they are a major driver of forest community dynamics (Goldblum & Rigg 2010). Although highly variable in space and time, disturbances can have a lasting effect on forests and are increasingly affecting forest management (Seidl et al. 2011a, Grecs & Kolšek 2017, Danneyrolles et al. 2019. ...
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Kutnar L., Kermavnar J., Pintar A.M., 2021. Climate change and disturbances will shape future temperate forests in the transition zone between Central and SE Europe. Ann. For. Res. 64(2): 67-86. Abstract It is expected that climate change as well as abiotic and anthropogenic disturbances will strongly influence temperate forests. Besides changes in the main climate variables, various disturbance factors may significantly worsen conditions for mesic Slovenian forests (SE Europe) dominated by European beech (Fagus sylvatica), Norway spruce (Picea abies) and European silver fir (Abies alba). In Slovenia, the climate has warmed in recent decades, with an average annual rate of increase of about 0.4°C per decade or even more than 0.5°C per decade in summer. In addition, disturbances have caused considerable damage to trees in the most extensive forest types in Slovenia, starting with a widespread ice storm in 2014, followed by bark beetle outbreaks, windthrows and salvage logging interventions. After 2014, salvage logging increased from about one third to two thirds of the total annual felling. Over the last two decades, we have observed a decline in Norway spruce growing stock, with the highest rate of decrease in areas below 500 m a.s.l., and an increasing trend for European beech. Overall, the three dominant species (beech, spruce, silver fir), which together account for more than 70% of the total growing stock, have shown a declining trend over the last 20 years. The patterns observed are broadly consistent with earlier predictions developed for different climate change scenarios and with those reported in many other European countries. Adaptive forest management, which implements close-to-nature silviculture, has been traditionally practised in the region under study and has the potential to play an important role in reducing the risks associated with the impacts of climate change and disturbances in the future.
... Forest phenology is a reliable indicator to analyze the impacts of climate change within temperate and boreal zones (Fernandez et al., 2015;Evans & Brown, 2017;Heyder et al.,2011). This transitional region, between temperate and boreal, is bracketed by species that are close to the limits of their environmental range, both to the north and the south, resulting in zones that are particularly susceptible to small changes in climate (Fisichelli et al., 2013, Goldblum and Rigg, 2010, Froelich et al., 2015. There is evidence for an overall longer growing season, particularly an earlier spring Greenup, across temperate and boreal forests. ...
Rapid climate change in recent decades has impacted forest, coastal, and social systems globally. In the northeastern U.S., alterations to the seasonal timing and duration of phenology cycles are a direct result of increasing temperatures, and monitoring these changes serves as a valuable indicator to analyze the impacts of climate change. Furthermore, increasing temperatures can influence when and how visitors recreate in natural landscapes. In the past decade, outdoor spaces have seen an increase in the number of visitors, partly as a result of climate change, that has influenced how resource managers and tourism suppliers plan for and respond to the impacts of visitation changes. In Maine, increased visitation and usage of public lands and coastal tourism destinations, such as Mount Desert Island (MDI), have altered the locations and timing of when people visit and how they interact and recreate within these spaces. For resource managers and tourism operators to successfully adapt and plan for continued changes to phenology and park visitation it is necessary to understand (1) how increasing temperatures will impact forests at different scales and (2) how to effectively apply both short and long-term visitation and natural resource management plans. Here we use an interdisciplinary approach to integrate biophysical and social science methods to: (1) estimate forest phenology response to multiple climate variables at different spatial and temporal scales across Maine, (2) understand resource managers’ perceptions of the impacts of climate change and the perceived barriers to incorporating adaptation strategies into decision-making, and (3) identify climate change impacts in Maine and develop planning priorities for tourism operators. To accomplish these goals, we first analyzed three vegetation phenology metrics derived from satellite imagery. We built linear mixed effects models to identify relevant climate and environmental variables which most influence the onset of the three phenology metrics. Using two emission scenarios, RCP 4.5 and 8.5, our results indicate that by 2100 the range of the onset of Greenup will occur 19-33 days earlier, Peak 13-21 days earlier, and Dormancy 5 days earlier than their 16-year average (2001-2017). In addition, an online questionnaire of 61 management personnel within the Maine Bureau of Parks and Lands revealed that the most significant barriers to adopting effective adaptation strategies include uncertainties of the effects of climate change, insufficient staffing, and lack of time. Furthermore, managers observed a dramatic increase in the number of visitors to lands managed by the PBL during 2020 as a result of the COVID-19 pandemic. Finally, to understand how some tourism operators on Mount Desert Island, Maine, are preparing for observed changes in climate and visitation, we conducted a series of participatory workshops and found that community engagement and cohesive communication are key to cope with the impacts of climate change and increased visitation. The interdisciplinary approach used here further quantifies how climate change is influencing the timing and duration of key phenological events in Maine and can be used to predict how those trends will continue through the century. Our results provide insights for tourism operators and recreation managers to prepare and adapt for continued changes to Maine’s natural landscapes resulting from global stressors, like climate change.
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Climate warming has the potential to influence forest composition and species recruitment over the course of the 21 st century. Although many of these impacts are expected to occur during the growing season, important life history events, like seed dormancy release, may be affected during the winter. For seeds of balsam fir (Abies balsamea (L.) Mill.) to germinate, they require a lengthy cold stratification period to break seed dormancy, which may not be experienced under warmer winters. Moreover, within Atlantic Canada, balsam fir populations experience very different climates. Dissimilarities among the genetics of these balsam fir populations and adaptations to their local environments may engender variations in population response to winter warming. In this study, we selected three balsam fir seedlots each from four different seed origin zones within Atlantic Canada and subjected them to simulated winter warming in outdoor seed plots that were heated to ≈ 6 • C above the ambient temperature from December to April. Contrary to our hypotheses, germination success of the heated balsam fir seeds did not significantly decrease relative to the controls, and there was no interaction between warming and seed origin zone. Seedlots of some seed origin zones exhibited variable responses to warming, suggesting that dormancy levels substantially differ among populations from similar climates. This diversity in phenotype expression within balsam fir populations may improve this species resilience under future climate change.
•Uniquely important aspects of the boreal forest carbon budget relative to other regions include: smaller human populations with less direct anthropogenic influences and management, extensive areas of slow-growing coniferous forests, large areas of peatland and wetland complexes, substantial amounts of carbon stored below-ground in the region’s soils, and the presence of permafrost in many of these soils. •Boreal forest carbon budget accounting and reporting relies heavily on the national forest inventory programs of the major countries in the region, with other modeling approaches used to fill in the gaps in undersampled geographies and component pools. •Both top-down, atmospheric inversion modeling and bottom-up, terrestrial biosphere modeling have been challenged by a paucity of available data over the large extent of the mostly remote forest lands of the boreal region. •Uncertainties are being addressed—and confidence in budget assessments is improving—as new methods and expanded data collections are coming online, particularly with remote sensing. •How the land base is defined and reported as “managed forest” in boreal nations (that have large areas of noninventoried forest) will have important, global-scale implications for policy actions to mitigate GHG emissions.
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A new core from the centre of Lago dell'Accesa (Tuscany, Italy) was sampled for pollen and charcoal analyses to provide a high-resolution sequence from 8400 to 7000 cal. yr BP. We combined series of microscopic charcoal, macroscopic charcoal and pollen to address the response of vegetation to fire at different spatial scales. Before 7900 cal. yr BP, broadleaved evergreen forests of Quercus ilex were the most important vegetational type in the area of Lago dell'Accesa. The subsequent decline of Q. ilex occurred when human-induced fires increased at the Mesolithic/Neolithic transition (c. 8000 cal. yr BP). Cross-correlation analyses show that fire was a key factor for vegetational change. Higher fire incidence affected the forest composition, converting evergreen forests to high-diversity open, partly deciduous forests and shrubby communities. The correlation is more pronounced at a local scale (macroscopic charcoal), whereas at a regional scale (microscopic charcoal) the vegetation followed the fire intervals with a more marked time lag (10–100 years). Climatic change, such as wetter periods inferred from lake levels, may have directly influenced the vegetational change, exacerbating the effect of human impact. Our study suggests that the disruption of evergreen broadleaved forests occurred when mean fire interval reached values as high as those of today's highly disturbed Mediterranean ecosystems. Hence broadleaved evergreen forests may not be as fire-resilient as assumed according to modern ecological paradigms. In view of the projected increase in fire frequency as a consequence of global warming, the present relict forests of Quercus ilex will be strongly affected.
The boreal forests of the world, geographically situated to the south of the Arctic and generally north of latitude 50 degrees, are considered to be one of the earth's most significant terrestrial ecosystems in terms of their potential for interaction with other global scale systems, such as climate and anthropologenic activity. This book, developed by an international panel of ecologists, provides a synthesis of the important patterns and processes which occur in boreal forests and reviews the principal mechanisms which control the forests' pattern in space and time. The effects of cold temperatures, soil ice, insects, plant competition, wildfires and climatic change on the boreal forests are discussed as a basis for the development of the first global scale computer model of the dynamical change of a biome, able to project the change of the boreal forest over timescales of decades to millennia, and over the global extent of this forest.
To recover direct evidence of surface fires before European settlement, we sectioned fire-scarred logging-era stumps and trees in 39 small, physically isolated sand patches along the Great Lakes coast of northern Michigan and northern Wisconsin. While much information was lost to postharvest fire and stump deterioration, 147 fire-free intervals revealed in cross-sections from 29 coastal sand patches document numerous close interval surface fires before 1910; only one post-1910 fire was documented. Cross-sections from the 10 patches with records spanning >150 yr suggest local fire occurrence rates before 1910 ca. 10 times the present rate of lightning-caused fire. Since fire spread between or into coastal sand patches is rare, and seasonal use of the patches by Native people before 1910 is well documented, both historically and ethnographically, ignition by humans probably accounts for more than half of the pre-1910 fires recorded in cross-sections.
Patterns in modern vegetation can be viewed from a paleoecological perspective as patterns on the upper surface of a box whose horizontal axes are longitude and latitude and whose vertical axis is time (Fig. 1 in Grimm, this volume). Paleoecological studies based on pollen data allow ecologists to peer beneath this surface and see how the changing location, abundance, and association of the individual taxa have produced different plant assemblages and influenced the development of the modern vegetation patterns. During the past 18,000 years, some taxa have always grown abundantly over wide areas; the histories of these taxa can be traced by the changing location of constant-abundance surfaces that stretch continuously from the top to the bottom of the box. Other taxa appear to emerge from almost nothing and expand suddenly to become abundant over wide areas. Some of these taxa may later decrease in abundance and extent only to reappear abundantly elsewhere at a later date. The three-dimensional patterns of taxon distributions represent the history of individual taxa and Illustrate how they have changed in abundance and location through time.