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Altered Fire Regimes Within Fire-adapted Ecosystems


by
Gardner W. Ferry
Bureau of Land Management
Robert G. Clark
Roy E. Montgomery
Bureau of Land Management
Robert W. Mutch
U.S. Forest Service
Willard P. Leenhouts
U.S. Fish and Wildlife Service
G. Thomas Zimmerman
National Park Service
Fires ignited by people or through natural causes have interacted over evolutionary time with ecosystems, exerting a significant influence on numerous ecosystem functions (Pyne 1982). Fire recycles nutrients, reduces biomass, influences insect and disease populations, and is the principal change agent affecting vegetative structure, composition, and biological diversity. As humans alter fire frequency and intensity, many plant and animal communities are experiencing a loss of species diversity, site degradation, and increases in the size and severity of wildfires. This article examines the role fire plays in the ecological process around which most North American ecosystems evolved.
The five plant communities selected for study were the sagebrush steppe, juniper woodlands, ponderosa pine forest, lodgepole pine forest, and the southern pineland (Fig. 1). Status and trends of altered fire regimes in fire-adapted ecosystems highlight the role that fire plays in wildland stewardship. Fire regimes are considered as the total pattern of fires over time that is characteristic of a region or ecosystem (Kilgore and Heinselman 1990).
Fig. 1. Range of: a --sagebrush steppe; b -- juniper woodlands; c -- ponderosa pine; d -- lodgepole pine; and e -- southern pineland communities in the United States.

Sagebrush-grass Plant Communities
Greater frequency of fire has seriously affected the sagebrush steppe during the last 50 years (Table). One such community, the semi-arid intermountain sagebrush (Artemisia species) steppe, encompasses about 45 million ha (112 million acres). After repeated fires, non-native European annual grasses such as cheatgrass (Bromus tectorum) and medusahead (Taeniatherum caput-medusae) now dominate the sagebrush steppe (West and Hassan 1985). It is unclear whether cheatgrass invasion, heavy grazing pressure, or shorter fire return intervals initiated the replacement of perennial grasses and shrubs by the non-native annual grasses. It is clear, however, that wildfires aid in replacing native grasses with cheatgrass, as well as causing the loss of the native shrub component (Whisenant 1990). Inventories show that cheatgrass is dominant on about 6.8 million ha (17 million acres) of the sagebrush steppe and that it could expand into an additional 25 million ha (62 million acres) in the sagebrush steppe and the Great Basin sagebrush type (Pellant and Hall 1994). Table. Increase in the number of wildfires and area burned on sagebrush steppe in Idaho (data from the Bureau of Land Management, Idaho State Office, Boise).

  1950-59 1960-69 1970-79 1980-89
Number of wildfires Data incomplete 1,344 1,406 2,334
Area burned (ha) 751,000 663,000 900,000 1,316,000

Western Juniper Woodlands
Juniper woodlands occupy 17 million ha (42 million acres) in the Intermountain region (West 1988). Juniper species common to this region are western juniper (Juniperus occidentalis), Utah juniper (J. osteosperma), single-seeded juniper (J. monosperma), and Rocky Mountain juniper (J. scopulorum). Presettlement juniper woodlands were usually savanna-like or confined to rocky outcrops not typically susceptible to fire (Nichol 1937).
Juniper woodlands began increasing in both density and distribution in the late 1800's (R.F. Miller, Eastern Oregon Agricultural Research Center, unpublished data; Fig. 2) because of climate, grazing, and lack of fire (Miller and Waigand 1994). Warm and wet climate conditions then were ideal for juniper and grass seed production. Fire frequency had decreased because the grazing of domestic livestock had greatly reduced the grasses and shrubs that provided fuel, and relocation of Native Americans eliminated an important source of ignition. Continued grazing and 50 years of attempted fire exclusion have allowed juniper expansion to go unchecked.

Fig. 2. Cumulative establishment of western juniper on Steens Mountain, Oregon (adapted from R.F. Miller, Eastern Oregon Agricultural Research Center, unpublished data).

Ponderosa Pine Forest
Decreases in fire frequency are also seriously affecting ponderosa pine (Pinus ponderosa) forests, a common component on about 16 million ha (40 million acres) in the western United States. Historically, the ponderosa pine ecosystem had frequent, low-intensity, surface fires that perpetuated park-like stands with grassy undergrowth (Barrett 1980). For six decades, humans attempted to exclude fire on these sites (OTA 1993). Fifty years ago, Weaver (1943) stated that complete prevention of forest fires in the ponderosa pine region had undesirable ecological effects and that already-deplorable conditions were becoming increasingly serious. Today, many ponderosa pine forests are overstocked, plagued by epidemics of insects and diseases, and subject to severe stand-destroying fires (Mutch et al. 1993).

Lodgepole Pine Forest
Like ponderosa pine forests, lodgepole pine (Pinus contorta) forests are experiencing a change in structure, distribution, and functioning of natural processes because of fire exclusion and increases in disease. Wildfire may be the most important factor responsible for establishment of existing stands (Wellner 1970). Historical stand-age distributions in lodgepole pine forests indicated an abundance of younger age classes resulting from periodic fires. Fire exclusion, by precluding the initiation of new stands, is responsible for a marked change in distribution of age classes in these forests (Fig. 3).

Fig. 3. Historical and actual age-class distributions of lodgepole pine forest.
Dwarf mistletoe (Arceuthobium americanum), the primary disease of lodgepole pine, also has a profound effect on forest structure and function, although it occurs slowly. Data show that chronic increases of dwarf mistletoe are partly due to the exclusion of fire (Zimmerman and Laven 1984) because fire is the natural control of dwarf mistletoe and has played a major role in the distribution and abundance of current populations and infection intensities (Alexander and Hawksworth 1975). As the frequency and extent of fire have decreased in lodgepole pine stands over the last 200 years, dwarf mistletoe infection intensity and distribution are clearly increasing (Zimmerman and Laven 1984).

Southern Pinelands
In contrast to the juniper, ponderosa pine, and lodgepole pine communities, fire frequencies have not drastically decreased in the 78 million ha (193 million acres) of southern pinelands. These pinelands are composed of diverse plant communities associated with long-leaf (Pinus palustris), slash (P. elliotti), loblolly (P. taeda), and shortleaf pines (P. echinata). Fire has continued on an altered basis as an ecological process in much of the southern pinelands; historically, fire burned 10%-30% of the forest annually (Wright and Bailey 1982); the southern culture never effectively excluded fire from its pinelands (Pyne 1982), although human-ignited fires have partially replaced natural fires. Consequently, the amount of fire has been reduced and the season of burns has changed from predominately growing-season to dormant-season (fall or winter) fires (Robbins and Myers 1992). Altering the burning season and frequency has significantly affected southern pineland community structure, composition, and biological diversity (Fig. 4).

Fig. 4. Understory plant crown coverage after 30 years of burning (Waldrop and Harms 1987).

Implications
The role of fire becomes more complex as it interacts with land management. Maintaining interactions between disturbance processes and ecosystem functions is emphasized in ecosystem management. It is vital for mangers to recognize how society influences fire as an ecological process. In addition, managers must uniformly use information on fire history and fire effects to sustain the health of ecosystems that are both fire-adapted and fire-dependent. Managers must balance the suppression program with a program of prescribed fire applied on a landscape scale if we are to meet our stewardship responsibilities.
For further information:
Gardner W. Ferry
Bureau of Land Management
Division of Fire and Aviation Policy and Management
3905 S. Vista Ave.
Boise, ID 83705
References
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Barrett, S.W. 1980. Indians and fire. Western wildlands 6(3):17-21.

Kilgore, B.M., and M.L. Heinselman. 1990. Pages 297-335 in J.C. Hendee, G.M. Stankey, and R.C. Lucas, eds. Fire in wilderness ecosystems. Wilderness management. 2nd ed. North American Press, Golden, CO.

Miller, R.F., and P.E. Waigand. 1994. Holocene changes in semi-arid pinyon-juniper woodlands: response to climate, fire and human activities in the Great Basin. Biological Science 44 (7). In press.

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Whisenant, S.G. 1990. Changing fire frequencies on Idaho's Snake River Plains: ecological and management implications. Pages 4-10 in E.D. McArthur, E.M. Romney, S.D. Smith, and P.T. Tueller, eds. Proceedings of a Symposium on Cheatgrass Invasion, Shrub Die-off, and Other Aspects of Shrub Biology and Management. U.S. Forest Service Gen. Tech. Rep. INT-276. Intermountain Forest and Range Experiment Station, Ogden, UT.

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Zimmerman, G.T., and R.D. Laven. 1984. Ecological interrelationships of dwarf mistletoe and fire in lodgepole pine forests. Pages 123-131 in F.G. Hawksworth and R.F. Scharpf, technical coordinators. Biology of dwarf mistletoes: Proceedings of the Symposium. U.S. Forest Service Gen. Tech. Rep. RM-111. Rocky Mountain Forest Range Experiment Station, Fort Collins, CO. 131 pp.


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