CHAPTER 2
Secondary Successional Responses to Natural Disturbance, Forest
Management, and Climate Change in British Columbia’s Forests
Shikun Ran, Ecora Resource Group Ltd; Kathie Swift, FORREX
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2.1 EXECUTIVE SUMMARY
Natural and human-induced disturbances, such as wildfire, insect and disease outbreak, and forest harvesting,
are important drivers for forest renewal, stand structure, and ecosystem function. The use of natural
disturbance dynamics and historical variability as guidance for sustainable forest management practices have
been widely promoted (B.C. Ministry of Forests & B.C. Ministry of Environment, Land and Parks 1995;
DeLong & Tanner 1996; Drever et al. 2006; DeLong et al. 2007; DeLong 2010). The underlying assumption is
that forest ecosystems have adapted to the conditions created by natural disturbances and thus should be more
resilient to ecological changes associated with forest management activities if the pattern and structure created
by these activities resemble those of natural disturbance (Bunnell 1995; Angelstam 1998; Drever et al. 2006). It
is generally recognized that planning and operational management should maintain structural and
compositional heterogeneity at multiple scales, consistent with natural disturbance regimes over the landscape
and at the stand level (Drever et al. 2006).
Using physiographic regions, defined by the Canadian Committee on Ecological Land Classification framework
as “ecoprovinces,” this synthesis provides natural resource decision makers with scientific information on
potential secondary successional pathways following a variety of natural and human disturbances in different
ecosystem types across British Columbia. Secondary succession is defined as succession that takes place
subsequent to a disturbance which disrupts rather than destroys an existing biotic community (Spurr & Barnes
1980). This synthesis also provides information on projected ecological shifts resulting from changes in
temperature and precipitation and outlines some of the ramifications for current forest management practices.
With an uncertain climate future, decision makers will need to manage more proactively to reduce the
vulnerability and increase the resilience of future forest ecosystems. By providing more knowledge on how
various stand attributes within the forest communities are changed by natural disturbance agents, it is
envisioned that better resource management decisions will be possible.
2.2 INTRODUCTION
Natural and human-induced disturbances, such as wildfire, insect and disease outbreak, and forest harvesting,
are important drivers for forest renewal, stand structure, and ecosystem function. The use of natural
disturbance dynamics and historical variability as guidance for sustainable forest management practices has
been widely promoted (B.C. Ministry of Forests & B.C. Ministry of Environment, Land and Parks 1995;
DeLong & Tanner 1996; Drever et al. 2006; DeLong et al. 2007; DeLong 2010). The underlying assumption is
that forest ecosystems have adapted to the conditions created by natural disturbances and thus should be more
resilient to ecological changes associated with forest management activities if the pattern and structure created
by these activities resemble those of natural disturbance (Bunnell 1995; Angelstam 1998; Drever et al. 2006). It
is generally recognized that planning and operational management should maintain structural and
17
compositional heterogeneity at multiple scales, consistent with natural disturbance regimes over the landscape
and at the stand level (Drever et al. 2006).
The objective of this synthesis is to provide natural resource decision makers with scientific information on
potential secondary successional pathways following a variety of natural and human disturbances in different
ecosystem types across British Columbia. With an uncertain climate future, decision makers will need to
manage more proactively to reduce the vulnerability and increase the resilience of future forest ecosystems. By
providing more knowledge on how various stand attributes within the forest communities are changed by
natural disturbance agents, it is envisioned that better resource management decisions will be possible.
The species variation that exists within a forest community can change over time and space. This change, or
succession, within the forest community can range from days and years to decades and centuries, depending on
the nature of the natural forces at play in the system. Mimicking the pattern and dynamics created by historic
natural disturbance regimes in future forest management is partially based on the assumption that the climate
will remain relatively stable through the next cycle of forests. Recent mountain pine beetle outbreaks in British
Columbia illustrate the vulnerability of forest ecosystems and potential catastrophic impacts caused by
changing climate on ecological communities and economic and social aspects of human society. The potential
for range shifts by tree species and associated ecosystems over the next century under the changing climate are
dramatic (Hamann & Wang 2006; Wang et al. 2006). Incorporation of climate change adaptation in future
forest management is thus considered an integral part of sustainable forest management for desired ecosystem
products and services (Spittlehouse & Stewart 2003; Spittlehouse 2005; Burton 2010). The central theme of the
“Future Forest Ecosystem Initiative” introduced by British Columbia’s Chief Forester is to synthesize and
expand our knowledge base to manage for resilient forest ecosystems under changing climatic, social, and
economic conditions. Management that maintains or enhances ecological resiliency in the face of a changing
environment has received increasing recognition as a way to attain ecological, social, and economic
sustainability goals (Campbell et al. 2009).
2.3 THE SCOPE OF THE SYNTHESIS
The physiographic regions we outline in this synthesis are defined using the Canadian Committee on
Ecological Land Classification framework of “ecoprovince.” An ecoprovince is an area with consistent climate,
physiological characteristics, relief, and regional landforms; is smaller than an ecozone; and consists of multiple
ecoregions. Figure 2.1 shows the ecoprovinces of British Columbia. To describe natural disturbance events
related to terrestrial ecosystems, we amalgamated the “Georgia Depression” ecoprovince with the “Coast and
Mountains” ecoprovince. Since no terrestrial ecosystems occur within the “Northeast Pacific” ecoprovince, we
do not discuss it here. To organize and present information pertinent to this synthesis, we used the following
eight ecoprovinces.
1. Coast and Mountains (includes Georgia Depression ecoprovince)
2. Southern Interior
3. Central Interior
4. Southern Interior Mountain
5. Sub-Boreal Interior
6. Boreal Plains
7. Taiga Plains
8. Northern Boreal Mountains
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FIGURE 2.1 Ecoprovince locations within British Columbia (Source:
http://www.env.gov.bc.ca/ecology/ecoregions/province.html )
A given ecoprovince typically encompasses several complex ecosystem types and several natural disturbance
regimes. Ecosystem types are presented through the concept of biogeoclimatic unit, which is a broad ecosystem
classification at the regional level (Meidinger & Pojar 1991). As expected, a broad ecosystem type at a regional
level may range from dry and warm to wet and cold at the sub-regional level, with several major natural
disturbance regimes and many successional pathways. Similar ecosystems and overlapping disturbance agents
also occur in different ecoprovinces. To narrow the scope of this synthesis, we focussed on characteristic
forested ecosystems and key disturbance drivers within each of the eight ecoprovinces in British Columbia.
We used a variety of information sources to prepare this synthesis, ranging from peer-reviewed journal articles
and technical reports to unpublished discussion papers and personal observations. The description of each
ecoprovince is based on work from Demarchi (1996); Meidinger & Pojar 1991) provided the context for the
ecosystems of British Columbia. Wong et al. (2004) carried out a comprehensive literature review on natural
agents of disturbance and their historical variability for all broad ecosystems in the province; their review formed
the bulk of information sources and literature on natural disturbance components. The B.C. Ministry of Forests
and Range compiled tree species composition data over several decades from pre- and post-harvested stands
(2008). Spittlehouse (2008) presented the climate change scenarios and potential adaptation strategy for the
province’s forest and range ecosystems, while Hamann and Wang (2006) used an ecosystem-based climate
envelope modelling approach to study the potential effect of climate change and redistribution of ecosystems and
tree species. The unprecedented outbreaks of mountain pine beetle (Dendrochonus ponderosae) across the British
Columbia landscape pose major challenges to forest managers and decision makers. Burton (2010) reviewed the
overall management strategies adopted by the Province in the context of sustainability and articulated the postdisturbance successional pathways under various natural disturbance and management regimes.
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2.4 DOMINANT DISTURBANCE REGIMES AND FOREST SECONDARY SUCCESSION
British Columbia’s complex and distinct landforms and wide range of climate and vegetation types makes it
one of the most ecologically diverse places on Earth. The climate is dominated by major flows from both the
Pacific Ocean and the Arctic. Marked vegetation belts are a striking feature of the regional vegetation. Fourteen
biogeoclimatic zones occur across the province, ranging from some of the largest temperate rainforests on
Earth to arid grassland, and from oak savannahs to alpine tundra and boreal forests (Meidinger & Pojar 1991).
Wildfire, wind, drought, insects, and disease are recognized dominant natural agents, causing the most
widespread ecosystem disturbance. Management activities, such as harvesting and salvage logging, have also
rapidly become a major form of disturbance. Most ecosystems in the province are affected by these disturbance
agents, but with varying frequency or magnitude (Wong et al. 2004), and thus different pathways of secondary
successional responses.
Organized by ecoprovince, the following sections summarize the potential secondary successional responses of
forested ecosystems to the dominant natural or human-induced disturbances under the changing climate.
2.5 COAST AND MOUNTAINS ECOPROVINCE
Ecoprovince Description
The terrestrial component of the “Coast and Mountains” ecoprovince extends from coastal Alaska to coastal
Oregon. In British Columbia, it includes the windward side of the Coast Mountains and Vancouver Island, and
all of Haida Gwaii. The Coast and Mountains ecoprovince consists of the large coastal mountains; a broad
coastal trough and the associated lowlands, islands, and continental shelf; and the insular mountains on
Vancouver Island and the Queen Charlotte Islands (Haida Gwaii) archipelago.
The predominant climatic influences in this ecoprovince are derived from moisture-laden frontal systems off
the Pacific Ocean and the subsequent lifting of these systems over the coastal mountains. As these air masses lift
over the mountains, precipitation is deposited on the windward slopes. In winter, oceanic low-pressure systems
dominate the area and pump moist, mild air onto the south and central coast. In summer, high-pressure
systems occur over the North Pacific Ocean and frontal systems become less frequent and tend to strike the
coast further north.
A broad range of coastal ecosystems, from very dry to very wet, are recognized by the biogeoclimatic
classification system (Green & Klinka 1994). Dominant ecosystems include the Coastal Western Hemlock
(CWH), Coastal Douglas-fir (CDF), and Mountain Hemlock (MH) biogeoclimatic zones. The CWH occurs at
low to mid-elevations on Vancouver Island, Haida Gwaii, and the Coast Mountains up to southern Alaska
(Meidinger & Pojar 1991). The MH zone occupies higher elevations above the CWH throughout the range of
the ecoprovince, while the CDF occurs only at low elevations and is geographically restricted to the southern
mainland, islands in the Georgia Strait, and southeastern Vancouver Island. The CWH is the most structurally
complex ecosystem with the largest area in the coastal region. It includes some of the wettest and most
productive ecosystems on Earth, where mean annual precipitation ranges from 1000 to 4400 mm. Summers are
cool and winters are mild.
Coastal Western Hemlock Zone
Coniferous forests of western redcedar (Thuja plicata) and western hemlock (Tsuga heterophylla) dominate the
landscape of the CWH. These forests are commonly referred to as “temperate rainforests” because of the mild,
wet climate in which they grow; these forests are complex and often very productive ecosystems (Meidinger &
Pojar 1991; Hallett et al. 2003).
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The Biodiversity Guidebook (B.C. Ministry of Forests & B.C. Ministry of Environment, Lands and Parks 1995)
classified the ecosystems of the CWH into two natural disturbance types (NDT). The lower-elevation dry to
moist ecosystems of the CWH, where infrequent stand-initiating events occur at a mean return interval of
approximately 200 years, are classified as NDT2. The principal disturbance agent here is wildfire. Wet and very
wet subzones or variants, including those of the higher-elevation MH zone, are grouped into NDT1 with rare
stand-initiating events and a mean disturbance return interval of approximately 250 years (B.C. Ministry of
Forests & B.C. Ministry of Environment, Lands and Parks 1995). The area disturbed during these events is
generally relatively small. Large-scale forest fire is typically a rare natural event in the wet and very wet
ecosystems. The interval between these disturbance events can be centuries to millennia (Lertzman 1992, 2001;
Wong et al. 2004; Daniels & Gray 2006) and, as a result, old-growth forests dominate the landscape. In the
absence of large-scale forest fires, gap dynamics (Kimmins 1997) are the dominant processes driving forest
renewal and succession of wet coastal ecosystems (Daniels & Gray 2006). In gap dynamics, small canopy
openings (gaps) are created as a result of individual tree mortality events and are subsequently filled by
understorey tree species through advance regeneration and (or) seed germination.
Wind is a primary form of natural disturbance for creating canopy gaps, especially on the exposed outer coast,
where storms are often accompanied by strong winds (Meidinger & Pojar 1991). By blowing down individual
trees or small patches of trees, canopy openings or gaps are created in the forest canopy. These gaps allow light
to penetrate to the forest floor, stimulating the growth of shrubs and tree seedlings. More rarely, larger patches
are partially or completely blown down and the recovery cycle occurs on a larger scale.
Geomorphic disturbances such as landslides, debris flows, and avalanches also play an important role in
susceptible parts of the landscape and are the most frequent stand-initiating events, although wind, flooding,
and fire occasionally create larger canopy openings (Banner et al. 1983; Septer & Schwab 1995; Pearson 2000,
2003). Shade-tolerant tree species, such as western hemlock, mountain hemlock, and western redcedar,
gradually establish under the canopy gaps through advance regeneration or seeding from adjacent trees. This
process typically creates stands of a multi-aged, complex structure with a mixed species composition.
Processes that create, expand, and fill canopy gaps vary across wet to very wet ecosystems because of differences
in topography and vegetation patterns (Ott & Juday 2002). For example, fire-return intervals in the very wet
and maritime CWH can vary greatly with topography. Sites with high susceptibility to fire such as south-facing
slopes have substantially shorter fire return intervals compared to those with low susceptibility to fire, such as
north-facing sites and low terraces (see Gavin 2000; Gavin et al. 2003). The occurrence of isolated patches of
Douglas-fir (Pseudotsuga menziesii var. menziesii) on certain sites (Green et al. 1999; Gavin 2000), and the
requirement of a local seed source for Douglas-fir regeneration, suggests that portions of the CWH zone must
burn at intervals less than the lifespan of Douglas-fir in order for the species to be maintained on the landscape
(Wong et al. 2004).
Many existing dryer stands in the CWH zone on Vancouver Island and on the southern coast are thought to
have originated after moderate- to high-severity fires (Wong et al. 2004). This hypothesis is supported by
widespread evidence of cohorts of old Douglas-fir, a long-lived species that can establish following fire
(Hermann & Lavender 1990; Klinka et al. 1998). The date of establishment of old Douglas-fir cohorts on
northern Vancouver Island and in the Lower Mainland suggests fires occurred somewhere in the region nearly
every century for the last 1000 years (Parminter 1990; Green et al. 1999); however, Douglas-fir can also
regenerate in large gaps of old stands, thus the simple presence of Douglas-fir in a stand may not indicate a fire
event (Franklin et al. 2002).
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Coastal Douglas-fir Zone
The CDF zone is limited to the areas on the southern Mainland, on several islands in the Georgia Strait, and on
southeastern Vancouver Island generally below 150 m elevation (Meidinger & Pojar [editors] 1991). The zone
lies in the rain shadow of Vancouver Island and the Olympic Mountains. Historically, infrequent standinitiating wildfire was the principal disturbance agent in the CDF. Douglas-fir dominates many of this zone’s
ecosystems because it is well adapted to living with fire (Meidinger & Pojar [editors] 1991). Old Douglas-fir has
thick, fire-resistant bark that protects it from all but the hottest flames. Many large, old trees show areas of
charred bark and fire scars at their base. After a fire, young Douglas-fir seedlings quickly colonize the burned
area. Fires that kill off other less fire-resistant species help to establish and maintain Douglas-fir as the
dominant tree in the area.
Low-severity fires did occur, but these might have been ignited primarily by First Nations people (Wong et al.
2004). Garry oak (Quercus garryana) meadow ecosystems were likely maintained by these low-severity fires.
Fire sizes were typically small, ranging from 1 to 500 ha. Disturbance dynamics likely resemble those in the
adjacent variant of the very-dry maritime subzone of the CWH. Ecosystems of the CDF are classified as NDT2
by the Biodiversity Guidebook (B.C. Ministry of Forests & B.C. Ministry of Environment, Lands and Parks
1995).
Forest Management and Secondary Succession
Forest management activities such as harvesting and reforestation can profoundly impact the structure,
dynamics, and functions of ecosystems. This is especially true in wet coastal ecosystems where large-scale,
stand-initiating disturbances such as fire are rare and the primary mechanism of forest renewal is through gap
dynamics. Historically, clear-cut logging was the main harvesting system used in coastal regions of British
Columbia (Mitchell & Beese 2002). This system, though economically appealing and operationally convenient,
is considered inconsistent with the natural disturbance regime of these coastal ecosystems. Recently, the
variable-retention silvicultural system has replaced clear-cut logging in many moist and wet coastal ecosystems
of the CWH (Mitchell & Beese 2002). Variable retention is a technique for retaining mature trees (singly or in
groups) as key structural elements of a harvested stand, at least until the next harvest rotation or in some cases
extended rotations, in an effort to maintain species and habitat diversity as well as forest-related processes. For
example, variable retention that emulates canopy gaps would create small patches that are dispersed through a
stand and would include multiple entries into the stand over many decades or centuries. Compared to the
conventional system of clear-cut logging, variable retention is focussed on what is retained in order to
maximize the preservation of structural elements and environmental values associated with structurally
complex forests (Franklin et al. 1997; Mitchell & Beese 2002).
Recently, the diversity of managed second-growth forests and their ability to eventually recover to a state
resembling natural old forests has generated much interest and concern. Ryan et al. (2009) studied species
diversity in a series of managed stands using a chronosequence approach on Vancouver Island. They found that
the overall number of species did not change dramatically between successional stages (young, mature, old). In
stands where western hemlock was the dominant species, the regeneration stage supported a greater number of
species than later stages did. They concluded that managed mature and old-growth forests do not necessarily
maintain larger numbers of species than younger successional stages, although the species assemblage and
abundance may differ at each successional stage. However, in their study of harvested coastal rainforests of
northern British Columbia, Banner and LePage (2008) found that species richness doubled from young to old
forests. They observed that vegetation succession following logging disturbance is driven primarily by the predisturbance species composition that survives the disturbance (i.e., most species found in the young forests are
present to some degree in older forests). The higher species richness typically occurring in old-growth forests is
largely due to the recruitment of additional cryptogam and herb species of low cover and constancy. Changes in
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understorey diversity are fairly short-lived following clear-cut logging and slash burning (Halpern & Spies
1995). Populations of most vascular plant species recover to original levels before canopy closure, with the
exception of severely burned sites, where diversity may remain depressed for a long period of time and some
species may experience local extirpation (Halpern & Spies 1995).
To better understand the impact of forest harvesting on tree species diversity across landscapes, the B.C.
Ministry of Forests and Range (2008) compiled tree species data for pre- and post-harvesting stands. The
following statistics are relevant to the Coast and Mountain ecoprovince.
• Provincially, approximately 31% of the forests across timber harvesting land of all regions are considered
a monoculture (i.e., composed of a single tree species) before harvesting (B.C. Ministry of Forests and
Range 2008). This percentage has decreased to 27% for post-harvested stands where free-growing status
has been achieved.
• In the Coast Forest Region, only 7% of forests are considered a monoculture before harvesting. This
percentage has increased to approximately 9% post-harvest at the free-growing stage.
• Approximately 7% of CWH and 20% of CDF forests are considered a monoculture before harvesting.
The percentages have changed to 9% and 26%, respectively, in post-harvested stands at free growing.
The above statistical trends are interesting when compared to the same statistics reported from other regions of
the province. For example, the statistics for the Interior, where clear-cut logging and artificial regeneration is
much more widely practised, show an overall decreasing trend of monoculture for the managed stands at the
free-growing stage. The increased proportion of monoculture in the coastal region might be partially
attributable to vegetation management practices undertaken prior to free-growing status. In moist and wet
ecosystems, vegetation competition flourishes after logging, and vegetation control efforts would
unintentionally suppress the regeneration of secondary tree species.
Climate Change and Forest Secondary Succession
The impact of climate change on forested ecosystems can be direct or indirect (Spittlehouse 2008; Utzig & Holt
2009). The direct impact is related to forest growth, structure, and species composition. Summer drought will
increase the mortality of some susceptible species such as western redcedar (Klinka & Brisco 2009). Even when
current species survive, their growth rates may be affected. Competition from species better suited to the
changed climate will increase. The indirect impact is related to changes in geomorphic processes, disturbance
mechanisms, or the physical components of ecosystems such as soil properties and water availability. Increased
summer temperatures can lead to a higher severity and frequency of fires. Increased frequency of extreme
storm events can lead to amplified ecosystem disturbance through landslides, flooding, and (or) windthrow.
The changed climate may also affect ecosystems by facilitating outbreaks of pests or pathogens, or by favouring
the spread of invasive species.
In the Coast and Mountain ecoprovince, the climate change impacts on successional responses are expected to
be different in the southern dry ecosystems compared to the wet to very wet ecosystems. The climate on the
southern dry ecosystems is expected to affect natural disturbance frequencies, with drought and insect
outbreaks likely to become more significant (Utzig & Holt 2009). These dry ecosystems are expected to become
progressively warmer and drier. The mortality of susceptible species such as western redcedar is expected to
continue. Based on the bioclimatic envelope study of Hamann and Wang (2006), the coastal Douglas-fir
ecosystem has the potential to expand its range northward along the coast and upslope by over 300% between
the years 2071–2100 (by the year of 2085). Decreased moisture and increased temperature will likely result in
higher fire frequency and perhaps more severe fires. It may also result in species such as red alder (Alnus rubra)
spreading upwards as it is very sensitive to drought.
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By contrast, climate change is expected to have a lower impact on forested ecosystems on wet to very wet
climates in the region (Spittlehouse 2008; Utzig & Holt 2009). Climate change will likely increase the number
and intensity of storms (Spittlehouse 2008) in these ecosystems, thereby increasing windthrow disturbances on
high-risk sites and landslide disturbance in wet to very wet ecosystems.
Burton and Cumming (1995) and Cumming and Burton (1996) investigated the effects of climate change
scenarios on species-level phenology and frost effects using a modified version of the ZELIG gap model. The
results suggest that lowland coastal temperate forests could undergo significant losses because some dominant
species, notably Douglas-fir and western hemlock, may not receive sufficient over-winter chilling to induce
cold-hardiness and then suffer serious damage from recurrent frosts that exist in bands and (or) pockets in the
area. If this happens, Garry oak ecosystems and other deciduous forests are expected to increase. Higherelevation coastal forests would generally benefit from longer growing seasons stimulating higher productivity.
More productive species (such as western hemlock) would seed-in from below. Nitschke and Innes (2008)
further refined the ZELIG algorithms into a Tree and Climate Assessment Tool for modelling ecosystem
response to climate change. This tool could be useful in helping to refine the scenarios described above. The
increase in productive species such as western hemlock could leave the MH zone with limited space to move, so
this ecosystem may get replaced over time. This will put species associated with these high-elevation forests at
risk because of the discontinuous nature of the current MH zone, leading to its potential disappearance within
the next 30–50 years.
2.6 SOUTHERN INTERIOR ECOPROVINCE
Ecoprovince Description
This ecoprovince lies east of the crest of the Coast and Cascade ranges and west of the Columbia Mountains
(Demarchi 1996). It is the southernmost part of the Interior Plateau system. This ecoprovince includes the
Thompson Plateau, the Pavilion Ranges, the eastern portion of the Cascade Ranges, and the western margin of
the Shuswap and Okanagan highlands.
This ecoprovince lies in the rain shadow of the Coast and Cascade mountains and contains some of the
warmest and driest areas of the province in the summer. Air moving into the area has already lost most of its
moisture on the windward west-facing slopes of the coastal mountains. In summer, occasional irruptions of hot
and dry air from the Great Basin to the south penetrate the area, bringing clear skies and very warm
temperatures. In winter and early spring, frequent outbreaks of cold, dense Arctic air penetrate to the southern
extremities; however, such events are less frequent than on the plateaus of the more northerly ecoprovinces.
Dominant broad ecosystems include the Interior Douglas-fir (IDF) zone, the Montane Spruce (MS) zone, and
the Engelmann Spruce–Subalpine Fir (ESSF) zone. The IDF zone occurs on plateaus at lower elevations and the
ESSF zone occurs at higher elevations of the plateaus and highlands. The MS zone, a transitional zone between
the IDF and the ESSF, occupies the middle elevations. The Bunch Grass (BG) and Ponderosa Pine (PP) zones
are smaller zones that occupy lowland and valley bottoms of the region. Much of the urban and agricultural
development currently occurring in this ecoprovince is concentrated in the BG and PP zones. These zones are
ecologically important ecosystems for understanding disturbance regimes, particularly in the context of climate
change. The ecoprovince is dominated by dry ecosystems, consistent with the region’s prevailing climate
regime.
Interior Douglas-fir Zone
A diverse and interesting array of ecosystems occurs within the IDF zone. Forests are dominated by Douglas-fir
(Pseudotsuga menziesii var. glauca) trees of all ages and sizes, with a grassy understorey. In the driest parts of
the zone and on hotter and drier sites, grassland and open ponderosa pine (Pinus ponderosa) forests
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predominate. Dense, closed-canopy spruce forests occur on wetter and cooler sites, such as riparian areas.
Lodgepole pine (Pinus contorta var. latifolia) is common at higher elevations of the IDF and where recent fires
have occurred. Wetlands are relatively uncommon and are found in depressions and around the periphery of
open water, such as lakes and ponds.
All IDF ecosystems in British Columbia are classified as NDT4, with frequent stand-maintaining fires and rare
stand-initiating events with a mean return interval of 150–250 years, although intensity and frequency will vary
with topography (B.C. Ministry of Forests & B.C. Ministry of Environment, Lands and Parks 1995). Dominant
agents for natural disturbance include wildfire, bark beetles, root rots, and defoliators such as western spruce
budworm (Choristoneura occidentalis) and Douglas-fir tussock moth (Orgyia pseudotsugata) (Wong et al.
2004). Fire is the primary mechanism influencing stand structure and succession. In grassland and open forests,
wildfire historically played an important role in maintaining grassland communities. The varied intensity and
frequency of wildfires across the landscape has resulted in a mosaic of mostly uneven-aged forests interspersed
with grassy and shrubby openings. Without regular grass fires, trees establish in open grassy areas and, over
time, grasslands can come to be dominated by trees. Frequent grass fires prevent forest encroachment on
grasslands by killing most young trees. There is strong evidence that fire suppression over the past several
decades has lead to forests taking over areas once occupied by grasslands (Gayton 2003).
Most forest fires in this zone are low-intensity surface fires (Meidinger & Pojar 1991). The fires scorch the
forest floor every 10–20 years and fire size is mostly less than 50 ha. The thick bark on old Douglas-fir trees
enables them to survive low-intensity fires, but many young trees and some understorey plants are killed. Over
time, repeated low-intensity fires create a forest made up of Douglas-fir trees of many ages—a multi-aged
(multi-cohort) stand. High-intensity fires occur on average every 150–250 years and frequently burn more than
50 ha at a time. These are stand-destroying fires that burn not only along the ground but also through the forest
canopy, killing trees of all ages in the stand. Following a high-intensity fire, lodgepole pine is often the first tree
to colonize. This results in pure, even-aged pine stands.
Fire intensity and frequency also vary with topography, which influences soil moisture redistribution. This
affects the intensity and the spread of fire, leading to structurally complex forest landscapes composed of multiaged patches with poorly defined stand boundaries (Sandmann & Lertzman 2003). On wetter sites, particularly
in the cooler parts of the IDF, fires are less frequent and perhaps less intense. This enables hybrid white spruce
(Picea glauca x engelmannii), a late successional species that prefers a moister environment, to establish under
the existing structure and grow to become an important part of the mature stand.
Bark beetles, insect defoliators, and root diseases also have significant effects on stand and landscape dynamics
in the IDF zone (Alfaro et al. 1984; Maclauchlan & Brooks 1994; Miller & Maclauchlan 1998). The timber losses
(i.e., its area and value) attributed to these disturbances are much higher than those associated with wildfire
(Parminter 1998). Much research has been undertaken to predict future outbreaks of insects and disease and to
understand the effects of management actions on these outbreaks (Alfaro & Maclauchlan 1992; Shepherd 1994;
Negron 1998).
Montane Spruce Zone
Subalpine fir (Abies lasiocarpa) and hybrid spruce are the main climax tree species of the MS zone, but they
rarely dominate the landscape because of frequent stand-initiating fires and the low ability of these species to
regenerate successfully in large openings (Meidinger & Pojar [editors] 1991). These climax species are relatively
infrequent as major components of the overstorey, except on wet and cooler sites such as riparian areas and on
northern aspect slopes (Lloyd et al. 1990). Lodgepole pine, an early seral “pioneer” species, can aggressively
colonize burnt areas, often with high density, and maintain dominance in the landscape. Lodgepole pine stands
frequently display even-aged and even-sized canopy structure. Subalpine fir and hybrid white spruce
commonly occur as advance regeneration beneath the canopy of lodgepole pine forests in areas where the
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ground remains moist enough to support their regeneration. In wetter climatic areas and on moisture-receiving
sites, maturing stands usually contain a mix of lodgepole pine, hybrid white spruce, and subalpine fir. Douglasfir is an important component of many stands in the dry MS ecosystems, particularly on sites with well-drained
soils and a dry moisture regime (Lloyd et al. 1990). On many steep, south-facing slopes, Douglas-fir can be a
climax species and forms a dominant component of the canopy.
Stand-initiating wildfire and severe mountain pine beetle outbreaks are the two main agents that maintain
forests of lodgepole pine at various structural stages across landscapes in the MS zone. Less severe disturbances
created by low-severity fires, bark beetles, fungal pathogens, and dwarf mistletoe (Arceuthobium americanum)
open the canopy and allow the release of shade-tolerant spruce and subalpine fir (Lloyd et al. 1990). All MS
subzones are designated as NDT3, with frequent stand-initiating events at approximately 150-year intervals
(B.C. Ministry of Forests & B.C. Ministry of Environment, Lands and Parks 1995).
Grassland ecosystems are uncommon in the MS zone (Lloyd et al. 1990). Grasslands occasionally occur on
warm, dry sites such as south-facing upper slopes and on dry ridges. Wetlands occur relatively frequently in the
Okanagan Highlands and on the Thompson Plateau.
Historically, lodgepole pine, the mountain pine beetle, and wildfire formed a complex relationship that can
drive forest renewal and succession towards a variety of outcomes. Lodgepole pine, when old enough and
present in adequate numbers, provides a prime food source for the mountain pine beetle. Following an
extensive outbreak, many dead lodgepole pine can become the major source of fuel loads for wildfire.
Lodgepole pine has developed a strategy that is linked with fire, regenerating after a wildfire via its serotinous
cones. Work done by Axelson et al. (2009) illustrated that this beetle–fuel–fire relationship can play a role in
determining the structure and composition of stands found in this ecosystem. For example, they found that
stand-initiating fires within this ecosystem can result in even-aged seral lodgepole pine stands; however, they
also found that multiple beetle disturbances can also create stands that have variable canopy and cohort
structure. These types of structures can contribute to the succession of non-pine species that are shade tolerant
(e.g., interior spruce and subalpine fir). Mixed-severity fires can also occur in these ecosystems, creating
complex structures in uneven-aged lodgepole pine stands as well. Multiple beetle disturbances maintain this
complex stand structure and can contribute to the succession of pine and non-pine tree species (e.g., trembling
aspen [Populus tremuloides]).
Forest Management and Secondary Succession
As described above, wildfire is the most important disturbance mechanism that drives stand structure,
vegetation dynamics, and landscape pattern of Southern Interior ecosystems. The historical fire regime is of
mixed frequency and severity over much of this region and is believed to be the prime mechanism for creating
and maintaining landscape heterogeneity and stand structural diversity. Without frequent stand-maintaining
fires, IDF forests may develop an understorey composed of shrubs and small-diameter trees, rather than the
grassy understorey that occurred as a result of the historical fire regime. As fire frequency decreases, fire return
intervals increase and burning is postponed. The fuel loads and structure of IDF stands change as a result,
through the accumulation of large quantities of fuel (live, dead, and ladder fuel), making these stands more
susceptible to stand-initiating fires (Arno 1980; Kilgore 1981; Steele et al. 1986). Fire suppression can also result
in higher stand density, forest invasion of grassland, increased incidence of catastrophic fire (e.g., 2003 Kelowna
wildfire), and increased levels of insect infestation (Day 1996). High stand density has been shown to result in
low tree vigour; low stand vigour increases the susceptibility of trees to insect outbreaks, particularly bark
beetles and defoliators (Day 1996).
The social and ecological consequences of severe fires reinforce the need for widespread suppression efforts, yet
suppressing fires without reducing fuels may lead to larger and more intense fires. When large-scale, intensive
fires do occur, they favour species such as lodgepole pine, leading to stands of simplified species composition
26
and structure. Another obvious effect from long-term fire suppression is to unnaturally increase the
proportions of over-mature forest on the landscape.
Harvesting has rapidly become a major form of disturbance in forested ecosystems of the Southern Interior
ecoprovince. Increased use of the silvicultural practice of tree planting, with its decreased reliance on natural
regeneration after logging, has raised many questions about post-harvest stand structure and species
composition at multiple scales. The B.C. Ministry of Forests and Range compiled tree species data from preand post-harvested forests and reported tree species composition by forest regions and biogeoclimatic zones
(2008). These data reveal the following trends in the Southern Interior ecoprovince.
• Approximately 56% of IDF and 43% of MS forests were considered a monoculture before harvesting. The
percentages have decreased to 29% and 14%, respectively, for IDF and MS in post-harvested forests at the
free-growing stage.
• The amount of deciduous forest increased significantly from 500 ha (0.15% of the land base) before
harvesting to 12 537 ha (or 4%) at free-growing for the IDF landscape alone.
The above statistics seem counterintuitive to the general belief that single-species planting in numerous
cutblocks should lead to a higher proportion of tree monocultures in harvested areas. While most postharvested cutblocks were planted with a single tree species, the decrease in the proportion of harvested areas in
single tree species may indicate a significant amount of natural regeneration and in-growth after planting. The
post-harvesting conditions provide a good environment for some tree species to regenerate naturally,
particularly when seed sources are available and seedbed conditions are favourable.
Selective logging (primarily in the form of diameter-limit cutting) was widely practised historically in many
ecosystems of the IDF. The residual stands, although multistoried, are typically composed of trees with smaller
diameters and with less vigour as compared to the original forests (Day 1996). This makes these stands more
prone to disturbance by other biotic agents such as bark beetles and disease.
Climate Change and Forest Secondary Succession
Nitschke and Innes (2008) summarized a complex series of effects from changing climate expected for
ecosystems in the Southern Interior. These included
• primary effects such as an increase in area burned and frequency of fire, and species biological thresholds
(related to moisture and temperature) being exceeded;
• secondary effects such as a decrease in disturbance refugia, loss of regeneration ability for some current
species, and influx of new species; and
• tertiary effects such as decreases and increases in ranges, changes in ecosystem composition, and a
decrease in habitat for some fauna.
They predicted that the PP ecosystem will eventually become similar to the BG ecosystem of today, and much
of the current IDF zone may become more like the PP zone. They did not expect all species and functions to
move as units into the space suitable for “new zones,” as individual species move or adapt at different rates and
scales. The new assemblages of species and the interaction of these species with their new environment and,
potentially, new disturbance agents, may lead to different paths of succession than those of today.
Hamann and Wang (2006) quantified the magnitude of possible ecosystem shifts, and projected a potential
increase in BG zone area of nearly 800% by 2085 (median of years 2071–2100). Potential shifts for other
biogeoclimatic zones include net area gains for the PP (+452%) and IDF (+207%), and a net area loss for the
MS (–68%). The loss for the MS zone indicates that, even after the gains through its expansion into ESSF, much
of the area now identified as MS will come to have conditions more similar to those associated with the current
27
IDF zone. At the stand level, some species may migrate to occupy sites newly within their potential climatic
range; other species will slowly adapt to the changed climate in situ, or become locally extirpated. Other
anticipated impacts are increases in invasive species colonization and dominance and the formation of new
species assemblages (Harding & McCullum 1997; Dukes & Mooney 1999).
At the species level, Hamann and Wang (2006) projected that species currently at their northern range limits in
British Columbia could gain potential habitat at a rate of at least 100 km per decade. Some species, notably
lodgepole pine, are expected to lose significant areas of suitable habitat, whereas the ranges of most common
hardwoods could be largely unaffected. The simulated spatial redistribution of “realized climate space” is
considerable, with the present-day climate in the MS zone projected to disappear particularly rapidly. The
predicted species trend is further supported by more recent work by Flower and Murdock (2009), who focussed
on potential changes in the distribution of Interior Douglas-fir and hybrid spruce.
Burton and Cumming (1995) and Cumming and Burton (1996) investigated effects of climate change scenarios
on species-level phenology and frost effects using a gap model. Their projections suggested that upper-elevation
species such as subalpine fir and spruce will be gradually replaced by Douglas-fir and western larch (Larix
occidentalis); however, Douglas-fir forests were projected to remain substantially unchanged, with possible
increases of drought-tolerant ponderosa pine.
Climate change poses a particular management challenge because of the long lifecycle of tree species. The
climate experienced during the later successional stages is expected to be significantly different from that of the
establishment stage. On the positive side, ecological communities dominated by trees are seldom in equilibrium
with the prevailing climate (Noss 2001), and this lag may allow the opportunity for passive and assisted
adaptation. Personal field observations have revealed that some of the most vigorous and productive Douglasfir plantations often grow outside their climatically suitable range.
2.7 SOUTHERN INTERIOR MOUNTAINS ECOPROVINCE
Ecoprovince Description
This ecoprovince lies east of the Southern Interior plateaus in the southeastern portion of the province
(Demarchi 1996). It consists of four main physiographic systems: (1) the highlands on the western flank, (2) the
Columbia Mountains, (3) the Southern Rocky Mountain Trench, and (4) the Continental Ranges of the Rocky
Mountains on the eastern flank.
The two principal forested biogeoclimatic zones in this ecoprovince are the Interior Cedar–Hemlock (ICH)
zone on the lower to mid-slopes of the Columbia Mountains and wetter localities in the southern Rocky
Mountains, and the ESSF zone on the middle to upper slopes of all mountains in the ecoprovince. Other minor
forested biogeoclimatic units include the PP zone in the Southern Rocky Mountains; the IDF zone in the main
valleys of the Shuswap and Okanagan highlands and Southern Rocky Mountain Trench; the MS zone in the
valleys and lower slopes of the southern Rocky Mountains and eastern Purcell Mountains; and the Sub-Boreal
Spruce (SBS) zone in the upper Fraser River watershed.
Interior Cedar–Hemlock Zone
The climate of the ICH is characterized by a relatively long, warm summer and cool, wet winter. Although
summer is relatively dry in most of this zone, the slow-melting snowpack helps to keep soil moisture levels high
(Meidinger & Pojar [editors] 1991). Commonly referred as the “Interior Welt Belt,” this zone has warm, moist
conditions in the southeast and cooler, wetter conditions in the northwest.
28
With higher tree species diversity than any other forested ecosystems in the province, ICH forests are the most
structurally complex stands in British Columbia. A total of 20 subzone variants range from dry to very wet, with
several tree species capable of occurring on many ecosystems. Western redcedar and western hemlock are the
characteristic climax tree species. Other tree species include ponderosa pine, Douglas-fir, western larch, lodgepole
pine, hybrid spruce, western white pine (Pinus monticola), trembling aspen, and paper birch (Betula papyrifera).
The occurrence and frequency of tree species on any one site depends on site and soil conditions, disturbance
history, and local climate. Subalpine fir and hybrid spruce typically thrive on wetter and cooler sites. Western yew
(Taxus brevifolia) and western larch also occur in the ICH ecosystems but are geographically limited. Wetlands
are relatively uncommon in the zone and are generally confined to valley bottoms.
In wetter parts of the zone where wildfires are rare events, trees grow to great sizes and ages, rivalling the giant
trees of the CWH on the coast. Forests contain many snags and a large accumulation of fallen logs and other
woody debris. These features of old forests provide valuable habitat for a wide variety of life forms, from
seedlings and fungi to birds and large mammals.
The natural disturbance regimes in the ICH zone are as complex as its ecosystems, with biogeoclimatic units
spread across three NDTs (B.C. Ministry of Forests & B.C. Ministry of Environment, Lands and Parks 1995).
All dry and marginally moist ICH units are classified as NDT3, and all moist units are NDT2. The wet to very
wet ICH units are classified as NDT1, with rare and small stand-initiating events occurring at a mean return
interval of approximately 250 years.
In drier parts of the ICH, wildfire can occur frequently. Past wildfires of various sizes and intensities have
created a landscape mosaic filled with young and old forest patches (see Wong et al. 2004). Because fires are less
frequent in the wetter parts of the zone or on wetter sites, these areas are often dominated by large tracts of very
old trees (Arsenault 1997). Pollack et al. (1997) estimated that return intervals of stand-initiating disturbances
range from 101 years in the dry, warm subzone areas, to 129 years in the moist cool, and 226 years in the wet
and cool landscapes. Compared to the coastal rainforest counterpart (i.e., the wet CWH), fire intervals of the
inland ICH rainforest appear to be shorter for any given topographic position (Arsenault 1997). Similar to the
wet CWH, fires occur more frequently on south-facing slopes, creating different seral stage distributions on the
valley floor, lower to toe slopes, and mid-slopes of valleys.
In the absence of fires, and in the period between fires, gap dynamics caused by the interaction between root rot
and windthrow are key components of forest dynamics at the stand level and account for significant areas of
the ICH landscape (Coates & Burton 1997). This is especially true in wet to very wet ICH ecosystems where fire
is typically a rare event. Other important disturbance agents include bark beetles, defoliators such as western
hemlock looper (Lambdina fiscellaria lugubrosa), and pathogens such as armillaria root rot (Armillaria ostoyae)
(Wong et al. 2004).
Engelmann Spruce–Subalpine Fir Zone
The ESSF zone is one of the most geographically extensive and ecologically diverse ecosystems in British
Columbia. Its 50 subzone variants attest to the diverse nature of the zone. The ESSF occurring within the
Southern Interior Mountain ecoprovince ranges from dry to very wet, and from warm to very cold. All of the
dry ESSF ecosystems in this ecoprovince are classified as NDT3, its moist subzones are classified as NDT2, and
the wet and very wet subzones are classified as NDT1 (B.C. Ministry of Forests & B.C. Ministry of
Environment, Lands and Parks 1995).
The ESSF zone occurs at higher elevations immediately below alpine tundra but generally above the MS, ICH,
or SBS zones. Climax species are Engelmann spruce and subalpine fir. Closed canopy forests are common at
lower and middle elevations of the zone. Compared to its northern counterpart, the ESSF zone of the Southern
Interior Mountains ecoprovince typically has a higher diversity of tree species, particularly at lower elevations
29
with a wetter and milder climate. Compared to the ICH and MS at lower elevations, the ESSF has a lower
diversity of tree species. Subalpine parkland, consisting of tree islands interspersed with herb-dominated
meadows, commonly occurs at upper elevations in this zone. Seral pine stands are widespread in subzones or
variants with a frequent fire history. Other tree species such as western hemlock are usually found in the wet
parts of the ESSF, and western redcedar and western white pine are frequently occurring species in warmer
parts of the zone; however, deciduous trees are generally uncommon.
The common occurrence of lodgepole pine in the dry ESSF reflects the greater frequency of wildfire compared
to other wetter ESSF ecosystems. After fires, lodgepole pine is often the first tree species to colonize the burnt
area. With stand development, Engelmann spruce and subalpine fir dominate the tree regeneration underneath
the lodgepole pine canopy. If fire is absent for a sufficiently long period of time, the dominant canopy species
will shift to Engelmann spruce and subalpine fir. Although fire is a major disturbance agent and occurs
relatively frequently in dry ESSF (Wong et al. 2004), the fire-initiated stands are not necessarily even-aged
because post-disturbance regeneration can take decades to establish (Jull 1990; Parish et al. 1999) owing to the
relatively harsh climate for tree regeneration.
Although the fire return intervals in wet parts of ESSF are long, most Engelmann spruce and subalpine fir
forests originated from fires (Antos & Parish 2002). These large, fire-origin stands were subsequently
transformed to take on a fine-scale pattern of all-aged cohorts, reflecting patchy mortality and regeneration
caused by senescence or less severe disturbances, such as budworm, bark beetles, or windthrow (Varga &
Klinka 2001; Antos & Parish 2002; Parish & Antos 2002). These agents create small canopy openings that
enable understorey species, such as subalpine fir and Engelmann spruce, to establish and grow. Some studies
reveal that the small canopy openings generally favour subalpine fir regeneration owing to its seedling
abundance in most ESSF stands and its greater shade tolerance (Parminter 1983; Varga & Klinka 2001). Others
concluded that the two species regenerate equally well and co-exist during stand development (Veblen et al.
1991; Antos & Parish 2002). Despite similarities in stand structure and species composition, age class
distributions are highly variable, suggesting that several site-specific developmental pathways may exist (Parish
2001).
Forest Management and Secondary Succession
A typical ICH forest often contains several tree species at the stand level. Sustained wildfire suppression may
facilitate the shift in tree species composition to late seral species. In the event of major insect outbreaks (e.g.,
mountain pine beetle), the mortality of one component of the canopy will free up growing space for the
residual components. This type of multi-species, complexly structured stand may offer us insights for the
management of resilient ecosystems in the future landscape.
Forest harvesting remains the dominant form of human-induced disturbance in ICH and ESSF ecosystems of
this ecoprovince. Given the tree species richness and productive capacity of ICH ecosystems, demand for a
greater supply of timber-related products has steadily increased. At the same time, calls have also increased to
manage these ecosystems for landscape biological biodiversity, reflecting the rareness of such inland rainforest
globally, whose high diversity and complex ecosystem structure provides critical habitat for a range of wildlife
species. Although several silvicultural systems have been employed in attempts to create landscape patterns that
mimic the natural disturbance regime of these ecosystems, clear-cut logging followed by tree planting has
remained the dominant system on the managed landscape. It is of particular interest to assess the managementinduced changes in tree species composition and diversity at multiple scales. Based on data compiled by B.C.
Ministry of Forests and Range (2008), the following statistical trends are revealed for tree species change in the
managed landscape.
Approximately 15% of ICH forests were considered a monoculture (i.e., composed of a single tree species)
before harvesting. The percentage decreased to 11% post-harvest at the free-growing stage.
30
• Approximately 16% of ESSF forests were considered a monoculture before harvesting. This percentage
increased to 22% post-harvest at the free-growing stage.
• Deciduous species in ICH ecosystems increased to 12 205 ha in managed forests compared to 266 ha of
pre-harvested forests, a near 46-fold increase.
Table 2.1 presents statistics for individual tree species (i.e., pure and leading), resulting from a re-organizing the
data provided in B.C. Ministry of Forests and Range (2008) compilation.
TABLE 2.1 Changes of tree species composition in pre- and post-harvested ICH and ESSF forests of the
Southern Interior Mountain ecoprovince
Species (% landscape harvested)
Biogeoclimatic
zone
Western
redcedar
Douglasfir
Lodgepole
pine
Hybrid
spruce
Western
larch
Hardwoods
Subalpine
fir
ICH (preharvesting)
18.4
22.6
17.7
20.5
4.6
13.4
1.9
ICH (postharvesting)
12.1
21.4
18.3
27.3
2.7
5.0
5.9
ESSF (preharvesting)
< 0.1
1
14.9
66.2
<1
<1
14.0
ESSF (postharvesting)
< 0.1
<1
17.2
37.5
< 0.1
< 0.1
37.5
Since the comparison is based on the percentage of landscape harvested, the composition and changes of
species type outlined in Table 2.1 cannot tell us whether any species diversity shifts have occurred at the stand
and site level. From the broadest scale (i.e., biogeoclimatic zone) at the landscape level, species composition and
diversity in the ICH seem well maintained in post-harvested landscapes. At site and stand levels, multiple
species were observed before crown closure in most ICH stands, even though the majority of harvested
cutblocks were planted with a single species (personal field observations). The potential for natural
regeneration in cutblocks presumably depends on seed sources, seedbed condition, and microclimate (see
LePage et al. 2000). The increased prevalence of monocultures in post-harvesting forests of the ESSF zone may
be partially attributable to delayed natural regeneration in this harsh environment.
The natural regeneration of ICH forests under partial harvesting depends on suitable seedbed substrate, adequate
subcanopy light levels, and seed dispersal distance of parent trees (LePage et al. 2000). Where canopies are
undisturbed, the lack of suitable seedbed substrate appears to be the main factor limiting seedling establishment.
Forest floor disturbance associated with partial cutting creates a diversity of favourable substrates for seedling
establishment, particularly in combination with the larger canopy gaps created by group selection harvesting
methods. Under clear-cut harvesting, the potential for natural seedling establishment can be very limited, even
near the cutblock edge, possibly owing to unfavourable microclimate conditions (LePage et al. 2000).
31
Climate Change and Forest Secondary Succession
Based on projections from ecosystem-based climate envelope studies (Hamann & Wang 2006), the ICH zone is
one that could potentially expand the most through shifts upward and northward by the median year of 2085.
The overall area gain of ICH ecosystems could be 207% by the end of the 21st century. Spatially, much of the
land base currently occupied by the wetter ecosystems of the ESSF, SBS, and MS zones will become suitable for
ICH ecosystems.
Although the ecosystem-based climate space may remain relatively stable, the following dynamics of tree
species in ICH forests are expected to change.
• Ponderosa pine, already a component in some dry and warm climate ecosystems of the ICH zone, will
expand significantly to other dry to moist, but warm, ICH ecosystems. Where local climates are expected
to become wetter and cooler, ponderosa pine may only be suited for well-drained and warm-slope sites.
• Douglas-fir, one of the dominant species currently in dry and moist ICH ecosystems, will expand to wet
and very wet ICH ecosystems, at least on well- to moderately well-drained sites. The good performance
of interior Douglas-fir plantations established decades ago in the ICHvk and ESSFwk shows the
intriguing potential of assisted Douglas-fir migration (personal observations). Based on the Hamann and
Wang (2006) climate envelope study of realized ecosystem space, Douglas-fir also has the potential to
migrate onto a significant portion of the land base currently occupied by the SBS and MS zones.
• Western larch, a species in some ICH ecosystems but with limited geographic distribution, is expected to
undergo a large geographic expansion (Rehfeldt & Jaquish 2010). Initial experience, and the excellent
performance of this species in outplanting trials in the SBS landscape during past decades (Jull 2010),
shows the great potential for expanding this productive and commercially important species. A 508%
gain in potential habitat is predicted based on the Hamann and Wang (2006) study.
• Lodgepole pine, a common species of ICH ecosystems and dry to moist ESSF sites (Table 2.1), is expected
to gain new habitat at upper elevations but lose habitat to ponderosa pine at lower elevations and in warm
ICH areas. Numerous examples of decades-old lodgepole pine plantations in wet and cool or cold ESSF
units demonstrate the adaptability of this species outside the range of its current climate (personal field
observations), although some suffer from diseases such as foliar and stem rust (Woods et al. 2010). In
general, the overall frequency of lodgepole pine will remain relatively stable in the ICH zone.
• Hybrid spruce will continue to be a major species in ICH ecosystems, occurring on the wet sites of drier
and warm climates, and will move upward to gain some habitat lost by Engelmann spruce.
• Deciduous species, such as paper birch and trembling aspen, are an expected component in many ICH
forests, particularly in lower-elevation dry, moist and warm ecosystems. Maintenance of a deciduous
component is generally considered significant for managing biodiversity and resilient ecosystems of
future forests (Campbell et al. 2009).
The ICH ecosystems host a high diversity of tree species with complex stand structures. This trend is expected
to continue under the changing climate. The predicted new climate ecosystem space for these tree species will
not likely be fully realized without assisted migration and adaptation. It is of paramount importance that trees
established today through assisted migration and adaptation must be able to both withstand and thrive in
future climates. This is also relevant for all ecoprovinces discussed in this report.
2.8 CENTRAL INTERIOR ECOPROVINCE
Ecoprovince Description
The Central Interior ecoprovince lies to the east of the Coast Mountains between the Fraser Basin and the
Thompson Plateau (Demarchi 1996). It contains the flat to rolling Chilcotin and Cariboo plateaus and the
southern two-thirds of the Nechako Plateau. It also contains the Chilcotin Ranges, west to the centre of the
32
Pacific Ranges and the Bulkley and Thatsa ranges. Some mountain ranges on the leeward side of the Coast
Mountains are included because of the much drier conditions and, therefore, more interior type of climate. The
area has a typical continental climate, with cold winters, warm summers, and a precipitation maximum in late
spring or early summer. In summer, intense surface heating and convective showers occur, and in the winter
frequent outbreaks of Arctic air penetrate the ecoprovince.
The dominant broad ecosystems include the IDF zone in the south, the Sub-Boreal Pine–Spruce (SBPS) zone in
the centre, and the SBS zone in the north. In addition, the BG zone occurs within the deeply entrenched valleys
of the Fraser and Chilcotin rivers, the MS zone occurs at middle elevations in the Chilcotin Ranges and
southern Chilcotin Plateau, and the ESSF occurs on the middle to upper slopes of all mountains in the
ecoprovince. The SBPS is the most widespread and characteristic ecosystem within this ecoprovince and thus is
the zone on which this synthesis is focussed.
Sub-Boreal Pine–Spruce Zone
The SBPS zone consists of two principal ecosystems, lodgepole pine forests and wetlands (Meidinger & Pojar
1991). Lodgepole pine is the most common tree species in the SBPS, and is the only tree species in many extensive
forest stands of very dry SBPS on the Chilcotin Plateau. In this area, lodgepole pine dominates the forest canopy
and is often the only tree species in the understorey, indicating that it is the climax tree species in these
ecosystems. Stands of hybrid spruce occur on moist and wet sites throughout the zone, but these stands are
usually small and located primarily around the edges of non-forested wetlands and adjacent to streams. In wetter
parts of the zone to the north and east, hybrid spruce is occasionally found in the canopy of mature pine stands, as
well as in the understorey. Trembling aspen is a common seral species throughout the zone but, like spruce, is
usually localized. Douglas-fir, subalpine fir, black spruce (Picea mariana), and black cottonwood (Populus
balsamifera ssp. trichocarpa) occur sporadically on some sites and in some geographically limited areas.
Because of its poorly developed drainage systems, the landscape of the SBPS contains abundant wetlands,
dominated by shrub carr, fens, and marshes. Most wetlands are devoid of trees, and the boundary between
wetland and dry upland forest is often abrupt.
Subzones of the SBPS are designated as NDT3, with frequent stand-initiating events at approximately 100-year
intervals (B.C. Ministry of Forests & B.C. Ministry of Environment, Lands and Parks 1995). Based on this
designation, the dominant disturbance agent is fire of mixed to high severity, even though other agents such as
mountain pine beetle and spruce beetle (Dendroctonus rufipennis) are also capable of causing mortality on a
large scale (Wong et al. 2004). The mountain pine beetle has long been an integral part of SBPS ecosystems.
Historically, endemic populations existed and small outbreaks occurred. Cold winters that killed most of the
overwintering larvae, and pine trees that are sufficiently vigorous to resist the beetle attack, have for the most
part, kept outbreaks in check. When winters are warmer and pine trees are older and less vigorous, the number
of beetles can reach epidemic proportions, such as the current, ongoing outbreak that has killed much of the
mature lodgepole pine forests in the SBPS and beyond. The intensity of the outbreak has resulted in attacks in
some younger stands of lodgepole pine, which are not traditionally attacked (see Maclauchlan 2006).
Fire frequency and severity vary across different ecosystems of the SBPS. In dry to very dry parts of the zone,
such as the Chilcotin Plateau, stand-initiating events are dominated by a mixed-severity fire regime, consisting
of many frequent small- to medium-sized fires punctuated by extremely large fires every 40–100 years (Francis
et al. 2002). Fire is less frequent in the wetter parts of the zone, where fire return intervals range from 91 to
170 years. Lodgepole pine trees are highly susceptible to wildfire but, after a fire, new pine seedlings establish
quickly. As a result, few stands in these forests are more than 120 years old, and most consist of dense pine
trees, all of similar age. The natural thinning process in densely regenerated stands can take a long period of
time, and some of the extremely dense (so-called “dog hair”) stands remain stagnant for extended periods of
33
time. Evidence suggests that some of these dense stands remain repressed during the entire rotational period
(personal field observations).
Forest Management and Secondary Succession
The predominant silvicultural system in the SBPS is even-aged management using clear-cut logging in various
cutblock sizes. After harvesting, natural regeneration with or without site preparation has been the principal
regeneration strategy (Steen & Coupé 1997). The success rate of natural regeneration has generally been high,
particularly in drier parts of the SBPS. The naturally regenerated stands are often unevenly stocked, with dense
patches and small stocking voids. Pre-commercial thinning and fill-planting are often required for stands to
reach a fully stocked and free-growing status. For greater stocking distribution control and shorter regeneration
delay, tree planting has increasingly been the principal method of regeneration.
Partial harvesting systems such as single-tree selection have only been used in stands where Douglas-fir is a
prominent component (Steen & Coupé 1997). Regeneration in these stands is by release of advance
regeneration and ingress of natural Douglas-fir regeneration. In lodgepole stands, canopy openings created by
harvesting, insects, pathogens, and wind are generally restocked by sufficient natural regeneration without site
preparation, particularly in larger openings of over 0.01 ha (Steen & Coupé 1997; Steen et al. 2007). The
decrease in the success of natural regeneration in higher-elevation pine stands (see Steen et al. 2007) indicates
that some form of site preparation may be required on wetter SBPS sites owing to unsuitable seedbed
conditions (i.e., a thicker organic layer on the forest floor).
Even-aged, single-species monocultures were the dominant forest type, with more species found under wetter
climates of the SBPS zone. Before harvest, 76% of stands were considered a monoculture in the SBPS landscape,
with the percentage decreasing to 56% post-harvest at the free-growing stage (B.C. Ministry of Forests and
Range 2008). The significant decrease in the prevalence of monocultures was surprising but may indicate the
substantial amount of natural regeneration of secondary tree species in the managed landscape, particularly in
wetter parts of the SBPS. Planting of Douglas-fir and hybrid spruce has increased in wetter parts of managed
SBPS landscapes in recent years (Ray Coupé, pers. comm.). Regardless of the species planted, lodgepole pine
will be a significant component of the ecosystems because of its natural regeneration capacity in most SBPS
ecosystems. Lodgepole pine was the principal tree species in both pre-harvested (90.5%) and post-harvested
(84.4%) stands at the free-growing stage (B.C. Ministry of Forests and Range 2008). A significant increase in
deciduous species, notably trembling aspen, was also noted, with less than 1% in pre-harvested stands and over
5% in managed stands post-harvest. Other tree species, such as interior Douglas-fir and hybrid spruce,
remained virtually unchanged in proportions (i.e., 5% for Douglas-fir and 3% for hybrid spruce).
Climate Change and Forest Secondary Succession
For SBPS ecosystems under a changing climate, Hamann and Wang (2006) projected the potential for a large
decrease in lodgepole pine and increased frequency of interior Douglas-fir. By the end of the century, the SBPS
ecosystem as presently defined could disappear almost entirely and largely be replaced by a more IDF-like
ecosystem. Some wetter SBPS ecosystems near the east and northern boundary of the present range may be
replaced by an ICH-like ecosystem. This potential shift in species and ecosystems may not be unexpected since
interior Douglas-fir is already a thriving species on many sloped sites of the SBPS landscape, where air drainage
is fairly good, particularly on warm, south-facing aspects.
This predicted shift in ecosystem state is further supported by independent climate modelling completed in the
Cariboo-Chilcotin area. Using both global and regional climate models, Dawson et al. (2008) analyzed regional
climate data for the past century and projected future climate change trends. Their results showed a climate
warming of 2.0–2.5°C by the year 2050. To put this into perspective, a 2.5°C warming in mean annual
temperature for Quesnel would make this region equivalent to the 1960–1990 baseline temperature for
34
Kelowna in the province’s Southern Interior. The results also showed that the temperature increases for the
region will be even greater after the year 2070. The Regional Climate Model projects some spatial differences in
climate change by 2050 for the Cariboo-Chilcotin region—summer and annual temperature increases could be
uniformly distributed across the region, while in the winter, the northern part is projected to warm more than
the south. Projected changes in annual and summer precipitation are predicted to be greatest in the central
Chilcotin Plateau and in a north–south band east of the Fraser River. Winter precipitation is projected to
increase over most of the region, especially in the northwest and north-central areas.
British Columbia’s mountain pine beetle infestation is now widely cited as the most destructive insect pest
outbreak in the province’s history, as well as a concrete example of a climate change impact (International
Panel of Climate Change 2007). Because of the predominance of mature and overmature lodgepole pine in
the SBPS, no other forested ecosystem in the province bears the same magnitude of impact from this
infestation. Many younger pine forests, including plantations, have also been killed (Maclauchlan 2006; B.C.
Ministry of Forests and Range 2007). As the pine stands die off, landscapes are transformed from living
forests to vast areas of standing dead trees. The majority of killed trees are still merchantable for a significant
period of time, and this has resulted in a dramatic increase in large-scale salvage harvesting in the beetleaffected areas. With or without management intervention, lodgepole pine will be able to grow back over
much of its previous area through natural or artificial means of regeneration (Burton 2010). Nevertheless, if
the climate is changing as the evidence now indicates, should we restore the forest to a composition and
structure that arose from conditions that prevailed more than a century ago (Burton 2010)? Managing the
post-beetle landscape of the SBPS is a serious challenge but an exciting opportunity for the region’s forest
managers, whose current forest-management decisions may be some of the most important in the province’s
history. The species or species mix selected for planting post-salvage SBPS stands will determine the
succession pathways of future forests in this ecoprovince.
2.9 SUB-BOREAL INTERIOR ECOPROVINCE
Ecoprovince Description
This ecoprovince lies north of the Central Interior ecoprovince, east of the Coast Mountains and west of the
Interior Plains, in the north-central part of British Columbia (Demarchi 1996). It consists of several physiographic
systems: the low-lying plateau area of the Nechako Lowlands, the northern portion of the Nechako Plateau, and
the southern portion of the Northern Rocky Mountain Trench. The mountains to the north and west include the
southern Skeena and Omineca mountains, while those to the east include the Hart Ranges and associated
foothills, the southern Muskwa Ranges and associated foothills, and the McGregor Plateau.
Prevailing westerly winds bring air from the Pacific Ocean to the area over the Coast Mountains by way of the low
Kitimat Ranges or the higher Boundary Ranges (Demarchi 1996). Much of this area is in a rain shadow, and
coastal air has low moisture content when it arrives. Moisture does enter the area when a southwest flow occurs
over the low Kitimat Ranges. Summer surface heating leads to convective showers, and winter frontal systems
result in precipitation that is evenly distributed throughout the year. Outbreaks of air from the Arctic are frequent.
The southern edge of the ecoprovince is near the typical southern extent of the Arctic air mass in January.
Dominant ecosystems include the SBS zone, which occupies much of the landscape on the Nechako Plateau,
Nechako Lowlands, Northern Rocky Mountain Trench, and many of the valleys; the ESSF zone, which occurs on
the middle to upper slopes of all mountains; and the ICH zone, which occurs in the wetter valleys of the Skeena
Mountains. This synthesis focusses on the two broad ecosystems represented by the SBS and ESSF zones.
35
Sub-Boreal Spruce Zone
The typical SBS forest is found on a vast, gently rolling landscape of glacial till covered with dense stands of
conifers (Meidinger & Pojar [editors] 1991). The zone’s upland forests contain a distinctive combination of tree
species. The climax species are hybrid white spruce and subalpine fir, which often dominate the coniferous
forests of the SBS landscape, particularly in wet sites and (or) the wetter subzones. Black spruce occasionally
occupies cold air sites with moist to wet soil moisture regimes. In drier parts of the zone, lodgepole pine is a
dominant species with hybrid spruce and subalpine fir in the understorey. Douglas-fir also occurs on welldrained, dry and warm sites and is mixed with other tree species on most mesic sites in the southern portion of
the ecoprovince.
The combination of a moist climate, deep snow, and higher precipitation results in highly productive forests,
often with a lush understorey. Hybrid white spruce is a particularly productive species because it combines the
straight, clean growing habit of the white spruce with the hardier qualities of the Engelmann spruce. Many
natural forests in this zone have a complex structure and mixed ages. In fire-origin stands, an even-aged
overstorey of lodgepole pine often has a multi-aged understorey owing to delayed regeneration of hybrid
spruce and subalpine fir. In wetter landscapes where fire is less common, insect and disease attacks are often
species- and age-specific, leading to open multi-aged stands. Deciduous tree species, such as trembling aspen
and paper birch, are common components of pioneer stands on dry to moist upland sites after wildfire and
dominate in areas of land clearing and prescribed fire. As forest stands succeed from earlier to later successional
stages, vascular plant diversity decreases, whereas canopy structure becomes more complex as gap dynamics
develop (Clark et al. 2003). Although the SBS forests contain few tree species, successional changes are
pronounced, with structure changing more than composition in the course of succession.
Given the relatively flat topography of a typical SBS landscape, wetlands are common ecosystems, which
include fens, bogs, swamps, and marshes. Natural grassland and shrub land are generally rare.
The process of forest secondary succession is often triggered by fires. Historically, most fires were initiated by
lightning strikes (Meidinger & Pojar [editors] 1991). In most places, lodgepole pine re-seeds and dominates an
area directly after fire. Lodgepole pine is particularly adapted to regenerate following a fire owing to the
serotinous nature of its cones. If the fire is not too severe, deciduous trees can sprout from underground roots
(aspen), the base of the trunk (birch), or can regenerate from seeds that disperse to the site within a year or two
after the fire. Depending on pre-existing vegetation and seed source availability, the first plants to grow back
after fire are often herbs and shrubs. Deciduous trees may dominate the ecosystem for a long time, but conifers
eventually prevail. The frequent fires in upland forests leave irregular patches that grow back at different rates,
creating a mosaic of forests of various types and ages.
Ecosystems of the SBS zone range from dry and warm to very wet and cool. Subzones are grouped into two
NDTs: all dry and moist SBS subzones are classified as NDT3, having frequent stand-initiating events with a
mean return interval of approximately 150 years; and all wet and very wet SBS subzones are classified as NDT2,
having infrequent stand-initiating events with a mean return interval of approximately 200 years (B.C. Ministry
of Forests & B.C. Ministry of Environment, Lands and Parks 1995). In recognizing the substantial variation in
natural fire events, such as the frequency, intensity, and size of remnants, several natural disturbance units were
further delineated within this ecoprovince and the SBS zone (DeLong 2010). Based on documented differences
in disturbance processes, stand development, and the temporal and spatial landscape pattern, these units are
thought to better reflect important elements that were not dealt with sufficiently under the NDT classification
system. The natural disturbance unit concept is currently used as an important part of forest management in
this ecoprovince (DeLong 2010).
Fire is an important disturbance mechanism for the natural regeneration of the forest, particularly in dry and
moist ecosystems of the SBS zone. Significant variations in fire frequency and severity occur in different SBS
36
ecosystems (see Wong et al. 2004). In the dry parts of the SBS, fire frequency ranges from 100 years to
170 years, while in its moist ecosystems, the range is from 100 years to 270 years. For wet to very wet SBS
ecosystems, the fire return interval ranges from 500 years to several thousand years. Thus, variation in the
frequency of these stand-initiating disturbances appear to follow a gradient in precipitation (DeLong 1998).
Fires of moderate to high severity dominate the natural disturbance regimes of most forests in the SBS zone
(Parminter 1990; DeLong & Tanner 1996; Delong 1998, 2010); however, regional precipitation patterns and
topography can influence disturbance size. For example, in montane topography, average patch size decreases
along a moisture gradient across SBS subzones and variants, and patch size distributions were significantly
different between the moist, wet, and very wet SBS ecosystems (DeLong 1998, 2010). In the very wet SBS
ecosystem on the McGregor Plateau, for example, over half of the examined patches were less than 10 ha and
less than 1% were greater than 1000 ha (Hawkes et al. 1997).
DeLong & Kessler (2000) investigated the ecological characteristics of mature forest remnants left by wildfire in
sub-boreal landscapes near Prince George, B.C., and found some remnants had an uneven-aged, episodic
pattern of lodgepole pine regeneration. In the absence of stand-initiating disturbances, gap disturbances drive
stand dynamics of the SBS ecosystems (Kneeshaw & Burton 1997). Even-aged patches are modified by standmaintaining disturbances (Lewis & Lindgren 2000) and patchy mortality within mature stands can be caused by
a wide range of disturbance agents, such as root rots, bark beetles, spruce beetle, western hemlock looper, forest
tent caterpillar (Malacosoma disstria), western balsam bark beetle (Dryocoetes confusus), and wind (see Wong et
al. 2004). Widespread and high-severity mountain pine beetle outbreaks, such as the ongoing infestation, can
also cause lodgepole pine tree mortality over a large area.
Patch mortality caused by insects and pathogens and associated gap dynamics can generate a fundamentally
different successional pathway than that of stand-initiating fires or harvesting. The gap dynamics generally
favour a succession dominated by late seral species, usually shade-tolerant species growing in the understorey,
whereas fire and logging remove this component and favour the re-establishment of early seral, shadeintolerant species (Sinton et al. 2000).
Engelmann Spruce–Subalpine Fir Zone
The ESSF zone occurs at higher elevations immediately below alpine tundra, and above the SBS or ICH of the
Sub-Boreal Interior ecoprovince. Engelmann spruce and subalpine fir are the climax tree species, with
lodgepole pine frequently occurring or dominating in drier parts of the ESSF ecosystems. Compared to more
southern ecoprovinces, the ESSF of Sub-boreal Interior ecoprovince typically has fewer tree species, even
though other structural attributes are as complex as those found in the southern ESSF.
The common occurrence of lodgepole pine in the dry ESSF reflects the greater frequency of wildfire compared
to other wetter ESSF ecosystems. All moist ESSF ecosystems are classified as NDT2, while all wet ESSF
ecosystems are classified as NDT1 (B.C. Ministry of Forests & B.C. Ministry of Environment, Lands and Parks
1995). Underneath the lodgepole pine canopy, Engelmann spruce and subalpine fir dominate tree regeneration.
In the absence of fire for a sufficiently long period of time, canopy dominance shifts to Engelmann spruce and
subalpine fir.
In wet ESSF forests, where large stand-initiating disturbances are infrequent or rare, small-scale disturbances
associated with the mortality and replacement of individual trees are a primary source of heterogeneity in forest
composition and structure (see Wong et al. 2004). Mean disturbance return intervals range from 219 to 794 years
based on the results of several studies (Hawkes et al. 1997; Steventon 1997; DeLong 1998; Steventon 2001).
Disturbance rates are correlated with climatic conditions, with wetter and more snow-prone ESSF ecosystems
exhibiting longer return intervals (e.g., Steventon 1997). Because of rare stand-initiating events and long lifecycles,
particularly for Engelmann spruce, the majority of the ESSF landscape is dominated by old forests.
37
Forest Management and Secondary Succession
Forest harvesting, salvage logging of forests attacked by mountain pine beetle, and fire suppression are all
important management regimes that potentially affect the successional pathways of future forest ecosystems of
the SBS landscape. Although harvesting is designed to simulate disturbances caused by fires, it cannot
completely duplicate the effects of a fire (DeLong 2010). Harvesting may approximate certain characteristics of
fire such as size, shape, and fire skips. Some other characteristics such as variable disturbance intensity can be
partially managed for.
Under the unmanaged condition, dominant natural agents such as insects and forest fires drive the pathways of
succession to the next generation of forests. Insects may proceed indiscriminately (i.e., kill mature trees) and
provide a favourable environment for surviving trees and advance regeneration to grow (mostly late seral tree
species such as subalpine fir). In contrast, stand-initiating fires may be caused by lightning in forests with or
without insect infestation. This process generally favours development of early seral and shade-intolerant tree
species such as lodgepole pine. As a result of these processes, the SBS landscape typically is heterogeneous with
variable patch sizes, species composition, and age class distribution.
The mountain pine beetle is currently the dominant agent driving disturbance on the moist to dry portions of
this ecoprovince. In addition to high overstorey mortality in beetle-affected stands, widespread salvage logging
and large fires in beetle-killed forests now account for most of the disturbance. In wetter hybrid spruce and
subalpine fir stands, the dominant disturbance remains clear-cut harvesting. Artificial regeneration using tree
planting is the common method of restocking these stands. Tree planting is generally believed to create forests
with lower tree species diversity and simple structure, particularly in stands where only a single tree species is
planted and vegetation control is used as a management intervention to achieve free-growing status. Concerns
have been raised that this type of homogenization through intensive silviculture can increase overall landscape
susceptibility to fires, insects, and diseases (Franklin & Forman 1987; Turner & Romme 1994; Bergeron et al.
1998). On the other hand, large infrequent disturbances such as harvesting may not undermine the
mechanisms determining species composition (Turner et al. 1997), with changes in species diversity expected
to be short lived following harvesting (e.g., Halpern & Spies 1995). Nevertheless, this may not happen if some
ecosystem components are damaged to the extent that natural repair cannot occur, such as with soil-destroying
fires or hydrology-changing logging operations. Therefore, only time (and monitoring) will tell whether SBS
ecosystems are capable of recovering to a pre-harvested state, even though we know that succession can take
different pathways, depending on the disturbance type.
At the landscape level, the B.C. Ministry of Forests and Range (2008) reported tree species composition and
diversity in pre- and post-harvested forests. The following statistics are relevant to the SBS and ESSF
ecosystems in this ecoprovince.
• In the SBS, monoculture forests have increased from 27% pre-harvesting, to 32% post-harvest at the freegrowing stage.
• In the ESSF, monoculture forests have increased from 14% pre-harvesting, to 27% post-harvest at the
free-growing stage.
• Deciduous types (pure and leading) have increased from 0.3% pre-harvesting in the SBS to 7.3% postharvest at the free-growing stage. A similar rate of increase was also reported in ESSF ecosystem.
38
Table 2.2 shows the species type changes for the pre- and post-harvested forests of both the SBS and ESSF zones
(B.C. Ministry of Forests and Range 2008).
TABLE 2.2 Changes in species in pre- and post-harvested forests of the SBS and ESSF zones of the Sub-Boreal
Interior ecoprovince
Species (% landscape harvested)
SBS
ESSF
Lodgepole
pine
Hybrid
spruce
Subalpine
fir
Lodgepole
pine
Engelmann
spruce
Subalpine fir
Pre-harvesting
46.9
48.1
2.0
21.9
55.3
16.2
Post-harvesting
47.1
39.3
3.9
26.6
47.9
20.3
Although the overall composition of the pine type did not change in the post-harvested forests of the SBS zone,
the specific composition change within the pine type is significant. For example, the lodgepole pine–hybrid
spruce proportion decreased from 60% to 45% in pre- to post-harvested stands, whereas lodgepole pine–
trembling aspen and lodgepole pine–subalpine fir types increased significantly in the post-harvested forests.
Likewise, the hybrid spruce–lodgepole pine proportion decreased from 37% to 28% in pre- to post-harvested
stands at the free-growing stage.
The massive outbreak of the mountain pine beetle in the pine forest of the SBS landscape poses a serious
challenge for forest managers and decision makers. The lodgepole pine type constitutes approximately 47% of
all forest types (Table 2.2) in the SBS ecosystems and the majority of lodgepole pine have been killed by the
beetle infestations. Management responses to the affected forests and follow-up actions have major
implications for the successional pathways of future forests. Burton (2010) has illustrated three potential
pathways for beetle-infested forests under various natural development and management regimes (Figure 2.2).
Figure 2.2 shows the alternative recovery trajectories and associated adaptive cycles in the wake of a mountain
pine beetle outbreak in lodgepole pine forests. Forest stands can develop into a more resilient mixed-species
forest (pathway A) if abundant advance regeneration of spruce, fir, and other tree species is present, or if postlogging silviculture establishes seedlings of diverse species and origins. Re-establishment of a new pine forest
(pathway B) requires forest fire or logging, followed by the establishment of lodgepole pine from scattered
cones or planted seedlings. Forest regeneration can be indefinitely deferred or inhibited by shrub growth
(pathway C) in situations where little or no advance regeneration is present, and where no fire, harvesting, or
silviculture occurs (Burton 2010).
Initial response to the mountain pine beetle outbreak was to control further infestations through sanitation
harvesting and selective harvesting of infested trees. The operation quickly shifted to salvage harvesting to
recover commercial timber before value deterioration. Through the increased allowable annual cut allocation in
the beetle-affected timber supply areas, salvage logging remains a major form of harvesting in the SBS
landscape today and perhaps for years to come. For reasons of economic and silvicultural efficiency, as well as
safety concerns, clear-cut logging remains the standard harvesting system employed for salvage logging, even
though it is unusual for all trees in an insect-attacked stand to be dead or damaged (Burton 2010). This system
of salvage operations would alter the post-disturbance successional pathway from one favouring late-seral,
shade-tolerant species, such as hybrid spruce and subalpine fir, to one favouring more early-seral, shade39
intolerant species, such as lodgepole pole and trembling aspen (e.g., Stadt 2001). Although the successional
responses to the beetle infestation are different in different ecosystems (Dykstra & Braumandl 2006; Hawkins &
Rakochy 2007), many studies on previously infested forests have indicated that most forests are capable of
being fully restocked with no management intervention—given sufficient time (Hawkes et al. 2003; Coates et al.
2006; Dykstra & Braumandl 2006; Forest Practices Board 2007; Burton 2010).
FIGURE 2.2 Alternative recovery trajectories and associated adaptive cycles in the wake of a mountain
pine beetle outbreak in lodgepole pine forests (from Burton 2010).
Climate Change and Forest Secondary Succession
The Prince George region has experienced an average warming trend of 1.3°C over the last century (Picketts et
al. 2009). Future warming is predicted to be greater in northern British Columbia than in the southern part of
the province and more pronounced in the winter than in the summer, particularly when looking at daily
minimum temperatures (Spittlehouse 2008). Warming is predicted to be least in coastal areas, where it is
moderated by the oceans. For the future climate of Prince George and area, the annual temperature is projected
to increase an average of 1.6–2.5°C by the middle of the 21st century (Picketts et al. 2009). Precipitation is
projected to increase by 3–10%, primarily in winter with possible decreases in summer. Although important
uncertainties exist in the specifics of these climate projections, the warming trend is consistent and is predicted
to intensify over time.
The ecosystem-based climate envelope study (Hamann & Wang 2006) projected the potential for a 69% loss of
SBS ecosystem distribution by the 2050s and 85% by the 2080s. Much of the moist and wet SBS area may evolve
to ecosystems that are ICH-like, while the drier SBS areas will be suitable for IDF-like ecosystems. At the
species level, hybrid spruce and subalpine fir are expected to continue as important species on the landscape.
Douglas-fir, however, is predicted to eventually dominate the landscape because of the lower frequency of early
40
frosts, the reduced severity of growing-season frosts, and a significant increase in the length of the growing
season (Cumming & Burton 1996). Good performance of Douglas-fir plantations established in the 1970s and
1980s on a range of sites on the wet SBS, seem to support the projections (personal field observations). Other
species trials of western redcedar and western larch in the SBS landscape (Jull 2010) provide valuable data from
the past decades on the adaptability of these species to the new climate and ecosystem space. Although
uncertainties are associated with species distribution modelling (Thuiller 2004), the trend of a warming climate
is consistent and many species, such as Douglas-fir and to a lesser extent western redcedar and western
hemlock, are already a component of the SBS ecosystems in certain limited geographic areas. The risk for the
assisted migration of these species on the managed SBS landscape would seem to be low and the opportunity in
the post-beetle landscape high to test these species within the changing environmental envelope.
The mountain pine beetle infestation on the vast SBS landscape provides an unprecedented opportunity for
forest managers and decision makers to strategically select species or seed sources that will be suitable for
reforesting harvested lands under both current and expected future conditions; such forests can be expected to
have an advantage in maintaining forest cover that is better adapted to a new climate (Burton 2010). Indeed,
the beetle infestation and focussed salvage logging may serve to break the “biological inertia” (Von Holle et al.
2003) of mature pine trees that have been holding sites for a century or more, and that might otherwise be
expected to delay the establishment of trees better adapted to current and future climates.
Under the changing climate, significant retreat of the ESSF ecosystem type is projected (Hamann & Wang
2006). Although Engelmann spruce and subalpine fir are expected to continue in their dominance of the
landscape, other lower-elevation species (e.g., lodgepole pine, western redcedar, western hemlock, and
deciduous species) will increase in importance through northward migration, particularly on disturbed sites.
The success or failure of existing lodgepole pine plantations in the high-elevation wet ESSF landscape needs to
be investigated for better understanding of the successional pathways of the anticipated forest types.
2.10 BOREAL PLAINS ECOPROVINCE
Ecoprovince Description
The Boreal Plains ecoprovince lies east of the Rocky Mountains and south of the Fort Nelson Lowlands
(Demarchi 1996). It occurs on the Alberta Plateau, and consists of plateaus, plains, prairies, and lowlands and is
generally of low relief apart from deeply incised river beds. It extends eastward, across northern Alberta,
Saskatchewan, and Manitoba, and the southern Northwest Territories.
The climate of this ecoprovince typically is continental in nature, where cold air masses infiltrate during winter
and warm air masses form in the summer under conditions of a high sun and long days. Moist Pacific air
masses most often have dried while crossing successive ranges of mountains before reaching this area. In
warmer months, rain is largely the result of surface heating, which leads to convective showers. Winters are
cold because irruptions of Arctic air meet no barriers.
The dominant ecosystems in this ecoprovince include the Boreal White and Black Spruce (BWBS) zone, with
the moist and warm subzone at lower elevations and the wet and cool subzone at middle to upper elevations.
The ESSF zone occurs west of the Rocky Mountain foothills on ridge summits south of the Peace River, and the
Spruce–Willow–Birch (SWB) zone occurs on ridge summits north of Halfway River.
Boreal White and Black Spruce Zone
A combination of fire history and extensive cultural disturbance, in the form of land clearing and prescribed
fire, has resulted in trembling aspen dominating about half of the forests in this moist and warm subzone of the
BWBS (DeLong et al. 2010). Balsam poplar (Populus balsamifera ssp. balsamifera) is common on lower slopes
41
and along stream and river courses. White spruce (Picea glauca) dominates moister sites where disturbance has
been limited. Lodgepole pine is present as a seral species on drier and poorer sites. Black spruce forests, often
with a minor component of tamarack (Larix laricina), are common on organic soils. Black spruce also occurs
mixed with lodgepole pine on upland sites with cold soils or limited rooting depth. Tamarack occurs to a
limited extent as pure stands on very wet, rich sites and rarely on some upland sites.
Muskeg, a peatland combination of bogs and nutrient-poor fens, is a common ecosystem throughout the
BWBS landscape (Meidinger & Pojar 1991). The most common trees in this ecosystem are stunted black spruce
and tamarack. Muskeg occurs over poorly drained, deep layers of peat. Other minor but important ecosystems
include boreal grassland and scrub communities occupying steep, south-facing slopes; and small but productive
marsh and shallow lake ecosystems throughout the BWBS.
Fire is the key stand-initiating disturbance agent operating in the BWBS zone. Flooding is a significant
disturbance agent along broad fluvial terraces adjacent to larger rivers (DeLong et al. 2010). The disturbance
rate from fire is generally estimated at about 1% of the total forested area per year (i.e., a fire cycle of 100 years)
but will vary from area to area depending on climatic and topographic factors. Historically, large wildfires
(> 1000 ha) dominated the landscape and upland sites were regenerated quickly by dense trembling aspen,
mixed trembling aspen and spruce, or lodgepole pine, resulting in large patches of relatively even-aged forests.
On wetlands, black spruce, tamarack, Alaska paper birch (Betula neoalaskana), and occasionally white spruce
regenerate after fire. In small areas where past fires were intense, stands may regenerate to willow or alder.
Almost all regeneration occurs within a few years of disturbance; however, as white spruce and black spruce
increase in size, these species become more obvious in many stands originally dominated by trembling aspen or
lodgepole pine (Kabzems & DeLong 2011). This shift occurs more rapidly and these species become more
dominant in the canopy on wetter sites. Post-fire stands on upland sites are often initially very dense and selfthin over time.
Tomentosus root disease is thought to be a key disturbance agent affecting white spruce and in some localized
areas may cause conversion from spruce-dominated stands to aspen-dominated stands over the course of 20–
40 years (DeLong et al. 2010). Eastern spruce budworm (Choristoneura fumiferana) may also cause significant
mortality of mature or immature spruce, especially in floodplain forests (e.g., Alfaro et al. 2001) and lead to
conversion of mixed to almost pure aspen stands.
Forest Management and Secondary Succession
In the absence of fire, the natural stand dynamics of boreal mixedwood forests are characterized by a gradual
shift from broadleaf-dominated to conifer-dominated stands (Lecomte et al. 2010). As a result, unmanaged
stands 0–80 years old vary in composition from “pure” broadleaf to “pure” conifer, and the presence of coniferdominated stands on the landscape increases with age (time since disturbance > 80 years).
Trembling aspen and white spruce are the two most important commercial tree species in the BWBS
ecosystem, and the two often grow in mixedwood conditions. Conventional forest management practice
promotes either a deciduous or coniferous component after harvesting even where the original stand was
considered mixed (Lieffers & Beck 1994). Current management practice promotes a managed species
composition based on the “pre-industry” forest conditions as a guide. These management objectives fail to
recognize the dynamic nature of boreal mixedwood succession and species composition and the concept that
forest type proportions change over time (Lecomte et al. 2010).
Rapid early growth rates of aspen compared to slower initial growth of white spruce make it difficult to balance
growing space requirements when both species are regenerated at the same time. The amount of white spruce
natural regeneration occurring after harvesting has generally been irregular in its distribution and highly
variable in its densities (Peters et al. 2002; Comeau et al. 2005; Kabzems & DeLong 2011). The proportion of
42
white spruce natural regeneration after harvest is similar to what is present in mature aspen-dominated stands,
but its distribution is more irregular after harvest. Stands that were previously mixedwoods have often become
dominated by deciduous species after harvesting (Ball & Walker 1997), and the vigorous and plentiful aspen
regeneration following clear-cut logging is usually a direct result of the promotion of root suckering by full
sunlight exposure, soil warming, and the cutting of the parent tree (Peterson & Peterson 1995).
Mixedwood forest proportions appear to have increased in managed stands at the free-growing stage compared
to pre-harvested forests (B.C. Ministry of Forests and Range 2008). Mixedwood forests accounted for 54% in
the pre-harvested landscape and this percentage increased to 85% post-harvest at the free-growing stage. The
largest increase, which occurred in deciduous-leading mixedwood types, may be partially attributed to the
definition of “mixedwood” used and the assumptions made in collecting this information for reporting
purposes. According to local experts, the area of “mixedwood” stands harvested (depending on the definition)
has been far less than the “pure” end of the spectrum over the past 25 years (R. Kabzems, pers. comm.).
The poor regeneration of the white spruce component of mixedwood stands has led to numerous silvicultural
systems research trials. Silvicultural systems that produce both coniferous and deciduous volume on the same
area have been promoted (DeLong 1991; Lieffers & Beck 1994). Temporal separation of the two species has
been the most common approach, using practices such as retaining white spruce advance regeneration (Brace
& Bella 1988), or underplanting aspen stands with white spruce (DeLong 2000). Establishing white spruce by
planting it under established aspen stands has substantial potential as a technique for regenerating boreal
mixedwood stands. The presence of an aspen overstorey serves to ameliorate frost and winter injury problems
and suppresses understorey vegetation that may compete with white spruce (Comeau et al. 2009), and can
further reduce attack by white pine weevil, Pissodes strobi (DeLong 2002).
Climate Change and Forest Secondary Succession
Much of the warmer BWBS that occurs within the Boreal Plains ecoprovince could become climate space for
IDF-like ecosystems by 2050 and PP ecosystems by the median year of 2085, according to the projections of an
ecosystem-based climate envelope study (Hamann & Wang 2006). Such a potential shift in ecosystem state is
truly remarkable compared to other possible ecosystem shifts in northern British Columbia. Although the
climate space could become IDF-like, growing season frost may become a major limiting factor for Douglas-fir
establishment and survival, if this species is considered as part of any adaptation strategy. While the climate of
the BWBS ecosystem will certainly become progressively warmer and drier, the ecosystem that may replace it
under a changed climate is debatable, particularly in the eastern parts of the zone (Utzig & Holt 2009). Fire
frequency and severity are expected to increase, possibly doubling in annual area burned as a result of warming
combined with decreased summer precipitation (Krawchuk et al. 2009).
Future climate change is projected to have major effects on forest structure and composition at the landscape
level, most notably by altering tree species distributions and frequency of stand-initiating disturbance regimes
(Utzig & Holt 2009; Brassard & Chen 2010). Boreal forests may experience a greater frequency of standinitiating fires through a reduction in annual precipitation, resulting in a greater proportion of grassland and
younger forest on the landscape (Overpeck et al. 1990; Thompson et al. 1998; Amiro et al. 2001).
Since climate change is expected to dramatically alter the structure and composition dynamics of the boreal
forest, management guidelines need to be adapted to better respond to the predicted shift in ecosystem state.
Mixedwood tree species such as trembling aspen and white spruce are expected to continue as prominent
components of future forests.
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2.11 TAIGA PLAINS ECOPROVINCE
Ecoprovince Description
This ecoprovince lies to the east of the northern Rocky Mountains in the northeastern portion of British
Columbia. It is characterized as a large lowland dissected below the Alberta Plateau surface by the Liard River
and its tributaries, namely the Fort Nelson and Petitot rivers. It extends into the upper Mackenzie River Basin
in the Northwest Territories.
The climate is continental and cold; dense Arctic air flowing unimpeded from the north can easily blanket the
area in winter and spring. The long sub-Arctic winters are generally dark with little heating by solar radiation.
In summer, its location between the Arctic and Pacific air masses gives it long periods of cloud cover and
unstable weather.
The BWBS zone dominates this ecoprovince with the moist and cool subzone occupying lower-elevation flat
areas and the wet and cool subzone dominating raised hills and ridges. This zone is characterized by long, cold
winters and short growing seasons (Meidinger & Pojar 1991).
The growing season climate of the BWBS in the Taiga Plains is similar to that of the Boreal Plains but is colder
in the winter (DeLong et al. 2010). Aspen and white spruce mixedwood forests dominate the better-drained
sites. On the extensive poorly drained sites, black spruce forests dominate along with a mix of tamarack on the
slightly richer sites. Black spruce also occurs mixed with lodgepole pine on poorer upland sites and on steeper
cool aspects where permafrost is present. Lodgepole pine is relatively common, especially on drier sites in
combination with black spruce or on very well-drained sites with coarse soils. Lodgepole pine is also more
common on all sites in the Liard River drainage. Paper birch occurs sporadically on moist richer sites.
Fire is one of the key stand-initiating disturbance agents along with flooding, which occurs along broad fluvial
terraces adjacent to large rivers. The disturbance rate from fire is generally estimated at about 1% of the total
forested area per year (i.e., a fire cycle of 100 years) but will vary from area to area depending on climatic and
topographic factors.
Historically, large fires (> 1000 ha) dominated the landscape, and upland sites were regenerated quickly by
dense trembling aspen, mixed trembling aspen and spruce, or lodgepole pine, resulting in large patches of
relatively even-aged forests. On wetlands, black spruce, tamarack, Alaska paper birch, and occasionally white
spruce would regenerate after a fire. Almost all regeneration occurs within a few years of disturbance; however,
as white spruce and black spruce increase in size, these species become more obvious in many stands originally
dominated by trembling aspen or lodgepole pine. On wetter sites, these shifts occur more rapidly and these
species become more dominant in the canopy. Post-fire stands on upland sites are often initially very dense and
self-thin over time.
Forest Management and Secondary Succession
The dynamics of trembling aspen–white spruce stands in the ecoprovince are distinguished by the large size of
individual trees, longevity, and the low occurrence of internal decay in trembling aspen. After disturbance, two
patterns of species establishment are evident. In codominant stands, recruitment periods for trembling aspen
and white spruce overlap, with white spruce recruited over a 29–58 year period behind the aspen. This lag
seems to indicate a dominant recruitment episode rather than a continuous recruitment process. In addition,
white spruce trees in codominant stands do not go through a period of suppression and then release, a pattern
often associated with stand-level trembling aspen mortality and commonly described in other boreal
mixedwoods (Kabzems & Garcia 2004). See Kabzems and Garcia (2004) for the description of mixedwood
succession in this ecoprovince.
44
Climate Change and Forest Secondary Succession
Predicted climate change shifts to BWBS ecosystems in the Taiga Plains ecoprovince are similar to those
predicted for the Boreal Plains ecoprovince (i.e., become more IDF-like, see previous section), but these
changes will take longer to occur (see Hamann & Wang 2006). By the median year of 2055 (2040–2070), BWBS
ecosystems will likely remain similar to their present state, although disturbances such as wildfire are expected
to be more frequent, severe, and intense owing to increased temperature and decreased precipitation (see
Krawchuk et al. 2009). By the median year of 2085, the ecosystem climate space will be similar to current IDF
and PP climates, according to the predictions of climate impact models (see Hamann & Wang 2006). This delay
in ecosystem state response is largely attributable to the colder climate in the Taiga Plains, where the prevailing
climate regime is influenced by cold and dense Arctic air.
2.12 NORTHERN BOREAL MOUNTAINS ECOPROVINCE
Ecoprovince Description
The Northern Boreal Mountains ecoprovince lies east of the Boundary Ranges of the Coast Mountains, west of
the interior plains, and south of the Yukon Territory, in the north-central portion of British Columbia
(Demarchi 1996). The general character of this ecoprovince is one of mountains and plateaus separated by wide
valleys and lowlands. This area encompasses the Teslin, Taku, Tanzilla, and Stikine plateaus; the Cassiar
Mountains; the Liard Plain and Liard Ranges; the northern portion of the Alsek Ranges; the Skeena and
Omineca mountains; the northern Rocky Mountain Trench; and the Muskwa Ranges and associated foothills.
Prevailing westerly winds bring Pacific air to the area over the high St. Elias Mountains and Boundary Ranges.
Coastal air is greatly reduced in moisture when it reaches the area, and this ecoprovince is characterized by
rain-shadow effects, resulting in some areas being very dry. Summer surface heating leads to convective
showers that, together with winter frontal systems, result in precipitation amounts evenly distributed
throughout the year. Outbreaks of Arctic air are frequent during the winter and spring. The rugged relief leads
to a complex pattern of surface heating and cold air drainage in the valleys.
Major forested ecosystems in this ecoprovince include the BWBS zone and the Spruce–Willow–Birch (SWB)
zone. The BWBS occurs throughout the valley bottoms and extensive plains, and SWB occurs throughout the
high valleys and middle slopes of the mountains. A treeless Alpine Tundra (AT) ecosystem occurs at elevations
above the SWB.
Spruce–Willow–Birch Zone
The SWB zone is the most northerly subalpine zone in the province and is found at elevations above the BWBS
zone (Meidinger & Pojar 1991). Winters are long and cold and summers are short and cool. Temperatures
average more than 10°C for only one month and mean annual precipitation is 460–700 mm. Lower elevations
(except for valley bottoms with prevailing cold air) of the SWB zone are generally forested with white spruce
and trembling aspen. At higher elevations, shrub/parkland dominates the landscape with deciduous shrubs
such as scrub birch and willow. Seral stands of lodgepole pine are relatively uncommon, indicating that fire
disturbance intervals are less frequent compared to adjacent low-elevation ecosystems (Meidinger & Pojar
1991). The disturbance regime is classified as NDT2, with a mean return interval for stand-initiating events of
approximately 200 years (B.C. Ministry of Forests & B.C. Ministry of Environment, Lands and Parks 1995).
Stands are often sparsely treed owing to extensive cold air drainage and cold temperatures (DeLong et al. 2010).
Older forests have short, large-diameter white spruce with variable amounts of subalpine fir. Extensive prescribed
burning in many valleys has resulted in widespread seral trembling aspen forests, particularly on warm-aspect
45
slopes. Black spruce is common on upland sites, often with lodgepole pine, on cooler-aspect slopes and in
wetlands. Balsam poplar occurs along streams and rivers and is often associated with white spruce.
Forest Management and Secondary Succession
The SWB has the harshest climate of all the forested ecosystems in the province. Very few forestry and
agriculture activities have occurred in the SWB landscape. Habitat management uses extensive prescribed
burning in many valleys, which have created large seral trembling aspen forests (DeLong et al. 2010) and many
grassy slopes (Ray Coupé, pers. comm.). In the absence of repeated stand-initiating events, those aspen forests
will gradually shift to mixedwood and eventually be replaced by a spruce-leading forest type.
Climate Change and Forest Secondary Succession
Climate impact models (Hamann & Wang 2006) suggest that the SWB ecosystem climate space will decrease
69% by 2025 and 93% by 2055. This predicted shrinking of the SWB ecosystem is the greatest projected loss
of all biogeoclimatic zones in British Columbia. The present SWB landscape is predicted to have a climate
similar to the current ESSF climate. Although many of the existing tree species, such as white spruce,
subalpine fir, lodgepole pine, and trembling aspen, will likely remain on the landscape under the changed
climate, the structure and density of the forests will undergo significant shifts. Forests will become denser
and multistoried, with significantly better productivity under increased temperatures and a longer growing
season. The increased frequency and severity of natural fires will also play an important role in realizing this
ecosystem climate space change.
2.13 CONCLUSIONS
For many decades, forest managers have relied on paradigms of ecological stability and the natural range of
historical variability as guidance for forest management decision making. An underlying premise is that by
maintaining forest conditions within the range of pre-industry conditions, these forests are likely to be
sustainably maintained into the future. Under a changing climate and other environmental conditions, forest
practices that rely on historical conditions and natural variability are not only costly but also ineffective because
these practices may create forests that are ill-adapted to current conditions and more susceptible to undesirable
changes (Millar et al. 2007). Acknowledging that the future will be different from both the past and the present
is a prerequisite if forest managers are to accept new approaches to forest management.
British Columbia will have greater warming and changes in precipitation than the global average (Spittlehouse
2008). The pattern of the changing climate will vary from region to region, with northern areas experiencing
greater warming than southern and coastal areas. Minimum temperatures are expected to increase faster than
maximum temperatures. While summer precipitation is projected to increase in the northern and coastal
regions, southern regions are expected to be drier. Extreme temperature and precipitation events are projected
to be of greater magnitude and more frequent, increasing the intensity and duration of heat waves, drought,
and storm severity. Concurrent with a changing climate are changes in the frequency and intensity of
disturbance by fire, insects, and disease (Dale et al. 2001; Volney & Hirsh 2005; Woods et al. 2010).
The potential ranges of species will move northward and upward in elevation, and it is likely that new
assemblages of species will develop in space and time (Cumming & Burton 1996; Hamann & Wang 2006;
Spittlehouse 2008). Across many regions of British Columbia, numerous examples exist of species outplanted in
areas that exceed their natural range of distribution and historical variability. Personal field observations reveal
that where these species are planted above and north of their normal ranges, they have performed substantially
better than those species planted below their normal ranges in southern or lower elevational range limits. Since
many of these plantations were established in 1970s and early 1980s when commercial timber value was an
important guide for species selection, field investigation and monitoring of these sites may provide us with
46
valuable information on patterns of survival, forest health, growth, and productivity. Without migration, many
species will be able to survive in their current location under a changing climate, but their growth rates will be
affected and competition will increase from other species or genotypes more suited to the new climate
(Spittlehouse 2008). Species may be unable to move into areas where the climate is suitable because of barriers
to movement, slow migration rates, unsuitable growing substrate, or lack of overall environmental habitat
(Stewart et al. 1998; Gray 2005). Human-assisted migration and breeding/selection for species adaptation will
be an essential component of the climate adaption strategy.
Predictions regarding the general trend of climate change are relatively consistent from various modelling
assumptions. A great uncertainty, however, is related to the frequency and intensity of extreme weather events,
which may ultimately determine the success of individual tree species under the newly gained habitat. It was
concluded with high confidence that the frequency and severity of wildfire, insect outbreaks, drought, and
extreme weather events will increase in North America in coming decades as a result of climate change
(International Panel of Climate Change 2007). Any adaptation strategy to this changing climate must ensure
that any negative impacts will be less severe than if no adaptation occurred (see Spittlehouse 2008).
In this synthesis, we largely based the direction and magnitude of tree species and ecosystem state shifts on the
projections made by Hamann & Wang (2006), who used average climate conditions of a given biogeoclimatic
zone. It is expected that the successional response of tree species and ecosystems to the changing climate will
vary by site conditions. The realized ecosystem-climate space may develop earlier on drier and warmer sites
and later on wetter and cooler sites than predicted.
Even where species composition and diversity after harvesting are maintained at the landscape level, other
important structural attributes exist, including species and structural diversity at the stand level, the patch size
mosaic of forest and openings, and seral stage distributions, all of which are expected to be different in
managed forests when compared to the forests of natural origin. The number of large green trees, snags, and
abundant downed wood, which are characteristic of old-growth forests, will decline in managed forests unless
special practices are implemented to maintain these features (Swanson & Franklin 1992). Other structural
changes are likely to occur in managed stands for attributes such as downed wood, thick forest floor litter, and
terrestrial moss and lichens, which provide important habitat for other species.
Promoting the resilience of forested ecosystems is the most commonly suggested adaptive option in a climate
change context (Spittlehouse & Stewart 2003; Millar et al. 2007; Campbell et al. 2009). Managing for resilient
forests and ecosystems in British Columbia represents a profound paradigm shift in forest management and
poses major challenges to existing forest practices and policies (Campbell et al. 2009). Resilient forests often
possess such traits as greater structural and compositional diversity and ecological complexity (Drever et al.
2006; Campbell et al. 2009). The recent mountain pine beetle infestation is a good example of the vulnerability
of the province’s forests under direct impact of climate change and the presence of a single, dominant tree
species over extensive geographic areas. Learning from this experience, and iteratively incorporating lessons
into future plans, is the necessary lens through which forest management must be conducted (Spittlehouse &
Stewart 2003; Burton 2010). Management practices such as assisting species migrations, or increasing diversity
in genetic and species planting mixes are considered appropriate (Millar et al. 2007; Campbell et al. 2009).
British Columbia’s Biogeoclimatic Ecosystem Classification system (Pojar et al. 1987; Meidinger & Pojar 1991)
and mapping provide a scientific foundation, as well as practical solutions to species management under a
changing climate. Forest managers have long been accustomed to using this system to guide their management
practices, including species and site selection for outplanting, stocking standards, and vegetation management.
To many forest managers, adaptation under a changing climate may be regarded as managing species and
similar ecosystems in new locations by using past experience. The largest challenges facing forest managers will
involve setting landscape-level objectives for desired future forest conditions and making the necessary policy
adjustments needed to implement resilience-based management approaches (Campbell et al. 2009).
47
The provincial government has several initiatives focussed on policy review and adjustments based on the best
emerging science and through incorporating past experience. The Chief Forester’s “Guidance on Tree Species
Composition at the Stand and Landscape Level”1 is a good example of policy alteration. Other initiatives
include the Future Forest Ecosystem Initiative,2 which provides scientific understanding of issues related to
climate change; and the Kamloops Future Forest Strategy,3 which focusses on a feasibility study of management
practices for species and habitat under a range of ecological conditions and future climate scenarios. Under a
rapidly changing climate, we cannot rely solely on historical forest conditions to provide us with adequate
targets for current and future forest conditions. Climate variability, both naturally occurring and
anthropogenic, as well as modern land use practices and stressors, create novel environmental conditions never
before experienced by forested ecosystems (Millar et al. 2007; Campbell et al. 2009).
2.14 ACKNOWLEDGEMENTS
The preparation and publishing of this report was supported by the British Columbia Ministry of Forests,
Lands and Natural Resource Operations through the Future Forest Ecosystem Scientific Council. The authors
are grateful for constructive comments made by Ray Coupé and Craig DeLong, both from the Ministry of
Forests, Lands and Natural Resource Operations, and Phil Burton from the Canadian Forest Service. We would
also like to thank the blind reviewers who provided additional comments and feedback to this document.
Notes
1. Available at:
http://www.for.gov.bc.ca/hfp/silviculture/TreeSpeciesSelection/Chief%20Forester%20Guidance%20on%20Tree%2
0Species%20Composition.pdf
2. For more information on the Future Forest Ecosystem Initiative, see:
http://www.for.gov.bc.ca/hts/Future_Forests/
3. For more information on the Kamloops Future Forestry Strategy, see:
http://www.for.gov.bc.ca/hcp/ffs/KamloopsFFS.htm
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Author information
Shikun Ran, RPF, Ecora Resource Group Ltd., 218–1884 Spall Road, Kelowna, BC V1Y 4R1. Email:
[email protected]
Kathie Swift, RPF, FORREX Forum for Research and Extension in Natural Resources, 360–1855 Kirschner
Road, Kelowna, BC V1Y 4N7. Email: [email protected]
57