Fire
ecology is a branch of ecology that focuses on the origins of wildland fire and
it’s relationship to the environment that surrounds it, both living and
non-living. A wildland fire is
defined as any fire that is burning in a natural environment. Fire ecologists recognize that fire is a natural process, and
that it often operates as an integral part of the ecosystem in which it occurs.
The main factors that are looked at in fire ecology are fire dependence and adaptation of plants and animals, fire history, fire regime and fire effects on ecosystems.
In
the 1930’s, researchers in the southern United States argued against the
negative perspective that has surrounded fire, with the belief that all fire is
bad. It was realized that the
devastating picture painted by huge-scale fires produced fear in the minds of
the public (and in politicians and scientists alike), and that this generated
detrimental results in response to any wildland fires. These researchers
recognized that there are species of plants that rely upon the effects of fire
to make the environment more hospitable for regeneration and growth.
Fire in these environments prepares the soil for seeding by creating an
open seedbed, making nutrients more available for uptake and often killing
plants that are invading into the habitat and competing with native species.
Fire
history deals with how often fires have occurred in a given geographical area.
Through recorded history, we can see into the recent past, but trees are our
source of information on fires in the distant past. Trees record their history through a system of growth rings
that develop on the trees each year. When
a fire goes through an area, the growth rings of that particular tree may be
scarred. On live trees this is
called a fire scar. Fire scars can
also be seen on dead trees. Tree
origin dates (calculated from the total number of rings) can also tell when
fires occurred, in that fires gave way for these new trees to develop.
The study of growth rings is called dendrochronology.
Utilizing dendrochronology, we can determine when fires have occurred in
the past, and sometimes determine their intensity and direction as well as other
information about the weather patterns in that era.
Fire
regime refers to the patterns of fire that occur over long periods of time, and
the immediate effects of fire in the ecosystem in which it occurs. There are
many ways to define a fire regime. Fire regime is a function of the frequency of
fire occurrence, fire intensity and the amount of fuel consumed.
The frequency is determined largely by the ecosystem characteristics, the
duration and character of the weather (whether the season is drier or wetter
than normal, etc.) and ignition sources. The
intensity of a fire is determined by the quantity of fuel available, the
fuel’s combustion rates and existing weather conditions.
Interactions between frequency and intensity are influenced by wind,
topography and fire history. There
are many other factors that can come into play when talking of fire regimes,
though this simple definition will work for most cases.
Approximately
90% of fires in the last decade have been human-caused, either through
negligence, accident or intentional arson.
Some of the fires caused by accidents and negligent acts are through
unattended campfires, sparks, irresponsibly discarded cigarettes and burning
debris. The remaining 10% of fires
are caused by lightning strikes, which are especially prevalent in the Western
United States and Alaska.
Benefits:
The ecological benefits of wildland fires often outweigh their negative effects.
A regular occurrence of fires can reduce the amount of fuel build-up
thereby lowering the likelihood of a potentially large wildland fire.
Fires often remove alien plants that compete with native species for
nutrients and space, and remove undergrowth, which allows sunlight to reach the
forest floor, thereby supporting the growth of native species. The ashes that remain after a fire add nutrients often locked
in older vegetation to the soil for trees and other vegetation. Fires can also
provide a way for controlling insect pests by killing off the older or diseased
trees and leaving the younger, healthier trees. In addition to all of the above-mentioned benefits, burned
trees provide habitat for nesting birds, homes for mammals and a nutrient base
for new plants. When these trees
decay, they return even more nutrients to the soil.
Overall, fire is a catalyst for promoting biological diversity and
healthy ecosystems. It fosters new plant growth and wildlife populations often
expand as a result.
Disadvantages:
Fire can cause soil damage, especially through combustion in the litter layer
and organic material in the soil. This
organic material helps to protect the soil from erosion.
When organic material is removed by an essentially intense fire, erosion
can occur. Heat from intense fires
can also cause soil particles to become hydrophobic.
Rainwater then tends to run off the soil rather than to infiltrate
through the soul. This can also
contribute to erosion. In
actuality, the negative effects of fires on soils are often exaggerated, and
many fairly intense fires in western United States forests cause little soil
damage. There is also the potential
for alien plants to become established after fire in previously uninfested
areas.
Fire suppression/Exclusion Policies
Wildfire
behavior and the effect of fire-exclusion policies on vegetation composition and
structure varies considerably (Smith and Fischer 1997). A substantial amount of
attention has been paid to the development of dense stockings of small trees in
some forests, and the contribution thereof to current fire severity in those
areas. Although this is generally true of drier vegetation ecosystems, this
observation does not always apply to many of the wetter and colder forests that
dominate much of the northern Rockies and the Pacific Northwest. Nor does it
apply to non-forested areas. The frequency of fire historically varies
considerably depending upon the type of vegetation in a given ecosystem. The
ecological effects of wildfire suppression policies instituted in 1911 have also
varied with vegetation type.
In
low-elevation ponderosa pine (Pinus ponderosa) and dry Douglas fir (Pseudotsuga
menziesii) forests, average fire intervals have historically ranged from 5
to 20 years, and low to medium intensity fires were common (Arno 1980, Smith and
Fisher 1997). Fire suppression has been fairly effective in reducing the number
of fire cycles that these low elevation dry coniferous forests have experienced
since the onset of fire suppression (Mutch 1994). This suppression of fire often leads to more intense fires in
these areas when fires do occur, due to the build-up of fuel and conditions that
are conducive to severe fire hazards.
Many dry coniferous
forests have now missed several fire cycles (Mutch 1994). Due to their
accessibility, these forests have also been extensively managed for timber
production and livestock grazing. The ecological consequences of these
management activities have caused a fairly dramatic change in tree density and
forest composition (Smith and Fischer 1997). These changes have often created
stands of dense, small-diameter trees in areas that used to be dominated by
widely spaced old-growth trees. Past management activities have clearly created
a situation in which a greater concentration of fuel is present, and there is a
higher probability of high-intensity fire, should a wildfire spread into or
start in the area.
In
contrast to the dry forests, subalpine forests composed mainly of subalpine fir
(Abies lasiocarpa), lodgepole pine (Pinus contorta var. latifolia), Engelmann
spruce (Picea engelmannii) and whitebark pine (Pinus albicaulis) cover vast
expanses of landscape. These forests are situated at higher elevations, which
are considerably wetter and colder than the dry forests discussed above. Some of
the conifer species present in sub-alpine forests are killed by
moderate-intensity fire (Bradley, et al. 1992). In contrast, lodgepole pine,
which is a dominant species in some sub-alpine forests, often reproduces
prolifically following wildfire (Agee 1993). This is due to the serotiny of its
cones. Some of the lodgepole pine cones are sealed shut by a resinous substance.
These cones often remain on the tree for years at a time, and the seeds are only
released when the resin is melted (at temperatures of 113-140 °F)
during forest fires.
Subalpine forests
typically burn rather infrequently, though often at a much higher intensity than
do dry forests. A few subalpine forest types (e.g. white-bark pine) experience
more frequent fire (Smith and Fisher 1997) but are very limited in distribution
(less than 18% of the sub-alpine forests in the northern Rockies). Historic
fire-return intervals in subalpine forests range from 50 to 300 years (Arno
1980, Smith and Fisher 1997, Agee 1990, Agee 1993). In many cases, historic
fire-return intervals for subalpine forests are longer than the period of time
in which the current fire-exclusion policies have been in effect. Fire exclusion
due to wildfire-suppression activities has not yet measurably altered the
structure and composition of the subalpine forests since they have, in general,
not missed fire cycles like the dry forests have. (Smith and Fisher 1997).
In the northern
Roackies, between the low-elevation dry forests and subalpine forests, a
mid-elevation zone of forest, composed of Douglas-fir (Pseudotsuga menziesii),
grand fir (Abies grandis), subalpine fir (Abies lasiocarpa),
lodgepole pine (Pinus contorta var. latifolia), red cedar (Thuja
plicata), western hemlock (Tsuga heterophylla), western larch (Larix
occidentalis) and other species, is found. The fire regimes and historic
fire-return intervals for these forests vary considerably with location and
forest type (Arno 1980, Bradley, et al. 1992, Smith and Fischer 1997). Historic
fire-return intervals range from 25 years to over 250 years in these stands (Arno
1980, Smith and Fisher 1997). Montane forests, generally, have been
substantially affected by forest-management activities (primarily logging).
These management activities and fire exclusion effects have largely varied
within the diverse regions of montane forests in the western United States.
In some areas, the effects have been subtle and slow to develop, while in
other areas fire exclusion has lead to the development of dense understory
vegetation and changes in forest composition (Smith and Fisher 1997).
Do severe wildfires burn in areas that are not composed of dense forests
resulting from fire exclusion and other land management activities?
In
the western United States, many areas that are not forested or only sparsely
forested often experience severe wildfires. Many of the wildfires that burn each
year burn in non-forested areas or involve substantial acreage of forests with
sparse tree cover. In these areas, forest-thinning programs are inappropriate
(due to a lack of trees) or would have little effect on fire behavior, because
the tree density is already low. In a study performed by the Pacific
Biodiversity Institute during the summer of 2000, seven recent major fires were
examined that clearly illustrate this point: the Kate’s Basin Fire, the Canyon
Ferry Fire Complex, the Mule Dry Creek Fire, the Hanford Fire, the Eastside Fire
Complex, the Maudlow–Toston Fire, and the Maloney Creek Fire. These fires all
burned through areas that included large portions of land that were not composed
of dense forests. This year (2001),
most of the large fires are burning in desert and sparsely forested country.
Examples of this are: the Sheepshead Fire (Oregon), the Lakeview Complex
(Oregon), the Sheep Complex (Nevada) and the Elk Mountain Complex (South Dakota
and Wyoming).
In the northern Rocky
Mountains, there are many areas that regularly experience severe wildfires that
are not in densely forested areas. Persistent seral shrubfields are widespread
in this region, a good example of which, are the “large expanses of
shrub-dominated slopes, where tree regeneration is sparse or lacking, that
characterize many areas in northern Idaho (Smith and Fisher, 1997)." Severe
reburns are the main cause of these persistent shrubfields. Some of these
shrubfields have persisted for 200 years or more, and have a mean fire-return
interval of about 31 years (Barrett 1982).
It is clear that persistent shrubfields are a product of wildfires
burning in an environment where forest thinning would have little benefit.
Could extensive thinning of forests have prevented the current fire
situation?
Silvicultural
thinning (i.e., logging of small-diameter trees to reduce tree densities and/or
underbrush) has been posited as a possible treatment method for reducing
wildfire risk. Although thinning to reduce fuel load has received much media
attention recently, it is controversial among the scientific community and
remains largely untested (Henjum, et al. 1994, DellaSala, et al. 1995, SNEP
1996). There have been few empirical studies looking at the effectiveness of
thinning as a treatment for reducing wildfire hazard (Frost 1999). The studies
that have been conducted have reported highly variable results. Some studies
indicate that thinning treatments designed to reduce fire risk actually increase
the risk and severity of the fires (Huff, et al. 1995, van Wegtendonk 1996,
Weatherspoon 1996). Although these treatments may reduce the flammable biomass
in the area, they also lead to drier forests and higher winds (Countryman 1955,
Agee 1997). Additionally, silvicultural treatments, even when conducted
carefully, can lead to the following adverse conditions (excerpted from Frost
1999):
Fires also burn in US Forest Service Inventoried Roadless Areas and designated Wilderness Areas. Many of the forests in these areas have not been severely altered from their historic fire regimes, and are difficult to access due to steep, rugged topography. Widespread thinning of backcountry areas is likely to be extremely costly, cause extensive environmental damage and create little benefit to society. Thus, the cost involved and the environmental disturbances of applying mechanical treatments over large roadless areas are not justified.
Thinning of
small diameter trees in dense, young forests may be appropriate and result in
reduction of wildfire risk to human communities in certain situations. The most
appropriate place to apply forest thinning is in dry forest types adjacent to
human communities threatened by wildfires. In these areas, it may be appropriate
to thin dense stands of young trees close to homes and community resources. Such
thinning needs to be followed up by a program of regular prescribed burning in
order to be effective. More research is needed on the efficacy of thinning
programs for wildfire risk reduction before there is conclusive evidence to
decide on their benefits or disadvantages.
What kinds of large-scale management practices should be implemented to
reduce wildfire risk? Where should these take place, if they should?
Large-scale
management practices are necessary to control the risk of wildfire in the
interface between forested and rural landscapes. The effects of fire suppression
and the potential for severe wildfire are greatest in these areas. While
rural-forest interfaces occur in many different forest types, they are most
common in dry and montane forests that have been the most altered from their
historic fire regimes by past management activities.
Many
researchers and scientists agree that the best way to reduce wildfire risk in
the rural-forest interface is through the reintroduction of fire to many natural
ecosystems (Walstad, et al. 1990, Mutch 1994, USDA/USDI 1995, Arno 1996, Frost
1999). Prescribed fire appears to be the most effective means for controlling
the rate of spread and severity of wildfire (van Wegtendonk 1996, Stephens
1998). Prescribed fire as a management tool has been increasingly used; however,
more burning is necessary to restore many ecosystems to their historic fire
regimes, thereby preventing the chance of an intense, large-scale wildland fire
(Mutch 1994, UDSA/USDI 1995, Arno 1996, Wright and Bailey 1982). The success of
prescribed fire lies in keeping the fire under control. In some instances,
mechanical treatments (e.g., thinning) may be applied to reduce the fuel loads
to a point at which prescribed fires can be effectively controlled (Mutch 1994).
Agee,
J.K. 1990. The historical role of fire in Pacific Northwest forests. Pp 25-38 In
Walstad, J.
et al. (eds.). Natural and
prescribed fire in Pacific Northwest forests. Oregon State
University Press.
Corvallis, OR.
Agee,
J.K. 1993. Fire ecology of Pacific Northwest forests. Island Press, Washington,
D.C.
Arno,
S.F. 1980. Forest fire history of the northern Rockies. Journal of Forestry
39:726-728.
Arno,
S.F. 1996. The seminal importance of fire in ecosystem management -- impetus for
this
publication. in The Use of
Fire in Forest Restoration. USDA Forest Service General
Technical Report
INT-GTR-341. Intermountain Research Station, Ogden, UT.
Beschta,
R.L. 1978. Long-term patterns of sediment production following
road construction and
logging in the Oregon Coast Range. Water
Resources Research 14:
1011-1016.
Bradley,
A.F., W.C. Fischer, and N.V. Noste. 1992. Forest ecology of the forest habitat
types of
eastern Idaho and western
Wyoming. USDA Forest Service Gen. Tech. Rep. INT-290.
Countryman,
C.M. 1955. Old-growth conversion also converts fire climate. U.S. Forest Service
Fire Control Notes 17(4):
15-19.
DellaSala,
D.A., D.M. Olson and S.L. Crane. 1995. Ecosystem management and biodiversity
conservation: Applications
to inland Pacific Northwest forests. Pp. 139-160 in: R.L.
Everett and D.M.
Baumgartner, eds. Symposium Proceedings: Ecosystem Management
in Western Interior
Forests. May 3-5, 1994, Spokane, WA. Washington State University
Cooperative Extension,
Pullman, WA.
Fahnestock,
G.R. 1968. Fire hazard from pre-commercially thinning ponderosa pine. USDA
Forest Service, Pacific
Northwest Region Station, Research Paper 57. Portland, OR. 16
pp.
Filip,
G.M. 1994. Forest health decline in central Oregon: A 13-year case study.
Northwest
Science 68(4): 233-240.
Forest
Ecosystem Management Assessment Team (FEMAT). 1993. Forest Ecosystem
Management: An ecological,
economic, and social assessment. Report of the Forest
Ecosystem Management
Assessment Team. July 1993. Portland, OR.
Frost,
E. 1999. The scientific basis for managing fire and fuels in national-forest
roadless areas.
World Wildlife Fund. CFR 64
No 201.
Grant,
G.E., and A.L. Wolff. 1991. Long-term patterns of sediment transport after
timber
harvest, western Cascade
Mountains, Oregon, USA. Pages 31-40 in Sediment and stream
water quality in a changing
environment: Trends and explanations. IAHS Publication
203. Proceedings of the
Symposium, 11-24 August 1991, Vienna, Austria.
Hagle,
S., and R. Schmitz. 1993. Managing root disease and bark beetles. Pages 209-228
in T.D.
Schowalter and G.M. Filip
eds. Beetle-Pathogen Interactions in Conifer Forests.
Academic Press, New York.
Harvey,
A.E., J.M. Geist, G.I. McDonald, M.F. Jurgensen, P.H. Cochran, D. Zabowski, and
R.T.
Meurisse. 1994. Biotic and
abiotic processes in Eastside ecosystems: the effects of
management on soil
properties, processes, and productivity. General Technical Report
PNW-GTR-323, U.S.
Forest Service, Pacific Northwest Research Station
Henjum,
M.G., J.R. Karr, D.L. Bottom, D.A. Perry, J.C. Bednarz, S.G. Wright, S.A.
Beckwitt,
and E. Beckwitt. 1994.
Interim protection for late-successional forests, fisheries, and
watersheds: National
forests east of the Cascades crest, Oregon and Washington. The
Wildlife Society Technical
Review 94-2, Bethesda, MD. 245 pp.
Huff,
M.H., R.D. Ottmar, E. Alvarado, R.E. Vihnanek, J.F. Lehmkuhl, P.F. Hessburg, and
R.L.
Everett. 1995. Historical
and current landscapes in eastern Oregon and Washington. Part
II: Linking vegetation
characteristics to potential fire behavior and related smoke
production. USDA Forest
Service Pacific Northwest Forest and Range Experiment
Station, PNW-GTR- 355.
Portland, Oregon.
Megahan,
W.F, L.L. Irwin, and L.L. LaCabe. 1994. Forest roads and forest health. Pages
97-99
in R.L. Everett, ed. Volume
IV: Restoration of stressed sites, and processes. General
Technical Report
PNW-GTR-330, U.S. Forest Service, Pacific Northwest Research
Station.
Meurisse,
R.T. and J.M. Geist. 1994. Conserving soil resources. Pages 50-58 in R.L.
Everett,
ed. Volume IV: Restoration
of stressed sites, and processes. General Technical Report
PNW-GTR-330, U.S. Forest
Service, Pacific Northwest Research Station
Mutch,
R.W. 1994. Fighting fire with prescribed fire – a return to ecosystem health.
Journal of
Forestry 92(11): 31-33
Smith
and Fisher. 1997. Fire Ecology of the Forest Habitat Types of Northern Idaho.
USDA
Forest Service General
Technical Report INT-GTR-363, 142 p.
Sierra
Nevada Ecosystem Project (SNEP). 1996. Status of the Sierra Nevada: Sierra
Nevada
Ecosystem Project, Final
Report to Congress Volume I, Assessment summaries and
management strategies.
Wildland Resources Center Report No. 37. Center for Water and
Wildland Resources.
University of California, Davis, CA.
Stephens,
S.L. 1998. Evaluation of the effects of silvicultural and fuels treatments on
potential
fire behaviour in Sierra
Nevada mixed-conifer forests. Forest Ecology and Management
105: 21-38.
Thomas,
J.W. et al. 1993. Viability assessments and management considerations for
species
associated with
late-successional and old-growth forests of the Pacific Northwest. USDA
Forest Service,
Pacific Northwest Region. Portland, OR.
U.S.
Deparment of Agriculture and U.S. Department of Interior (USDA/USDI). 1995.
Federal
wildland fire management
policy and program review: Final report. U.S. Department of
Agriculture and U.S.
Department of Interior, Washington, D.C.
van
Wagtendonk, J.W. 1996. Use of a deterministic fire growth model to test fuel
treatments. in
Status of the Sierra
Nevada: Sierra Nevada Ecosystem Project Final Report to Congress
Volume II. Wildland
Resources Center Report No. 37. Center for Water and Wildland
Resources. University of
California, Davis.
Walstad,
J.D., S.R. Radosevich and D.V. Sandberg, eds. 1990. Natural and prescribed fire
in
Pacific Northwest forests.
Oregon State University Press, Corvallis, OR
Weatherspoon,
C.P. 1996. Fire-silviculture relationships in Sierra forests. In Status
of the Sierra
Nevada: Sierra Nevada
Ecosystem Project Final Report to Congress Volume II.
Wildland Resources Center
Report No. 37. Center for Water and Wildland Resources.
University of California,
Davis.
Weatherspoon,
C.P. and C.N. Skinner. 1996. Landscape-level strategies for forest fuel
management. Pp. 1471-1492 in
Status of the Sierra Nevada: Sierra Nevada Ecosystem
Project Final Report to
Congress Volume II. Wildland Resources Center Report No. 37.
Center for Water and
Wildland Resources. University of California, Davis.
Wilson,
C.C. and J.D. Dell. 1971. The fuels buildup in American forests: A plan of
action and
research. Journal of
Forestry. August.
Wright,
H.A. and A.W. Bailey. 1982. Fire ecology: United States and southern Canada.
John
Wiley, New York, NY.