Effects of Fire on Mexican Spotted Owls in Arizona and New Mexico

Thesis

The Spotted Owl

General Characteristics:  The Mexican spotted owl is a resident raptor species found throughout the mountains and canyons of Arizona, New Mexico, southern Colorado and Utah, and northern and central Mexico. Most of these birds reside in a band of mixed-coniferous and ponderosa pine/Gambel oak (Pinus ponderosa/Quercus gambelii) forest stretching southeast from the southern portion of the Kaibab National Forest in northcentral Arizona down to the Gila National Forest in southwestern New Mexico. There are also substantial subpopulations located in the Sky Island mountain ranges in southern Arizona and in the Sacramento Mountains in southern New Mexico (Ward et al. 1995)

Adult Mexican spotted owls likely have a relatively high survival rate, with the probability of an adult surviving from one year to the next estimated at around 0.8 - 0.9 (White et al. 1995). Juveniles, on the other hand, have a much lower survival rate, ranging from 0.06 - 0.29 (Ganey et al. 1998; Willey 1998; White et al. 1995). There is a great deal of spatial and temporal variation in reproductive output, but one estimate places the general reproductive rate at 1.001 fledglings per pair (White et al. 1995). As is typical for K-selected species (Ricklefs 1990) the owl is long-lived with low reproductive output and generally maintains population densities near carrying capacity. The high survival rate of K-selected species enables them to maintain stable populations over time despite variable recruitment rates (White et al. 1995). The high survival rate of adult Mexican spotted owls may also help them to withstand fluctuations in habitat quality, such as might be expected from periodic fire within their territories.

Mexican spotted owls typically nest and roost in structurally-complex, diverse forests with a variety of age- and/or size-classes, a component of large trees, often with many snags and down logs and relatively high basal areas and canopy closures (Ganey et al. 1999; Gutiérrez 1996; Ganey and Dick 1995). These conditions are typical of old-growth type forests that have generally had minimal human-caused disturbance (Helms 1998). Ganey and Balda (1994), in a study of radio-tagged owls in northern Arizona, found that they did not forage randomly among available habitat types. Rather they tended to be found more often than expected (assuming random habitat selection) in unlogged forests and less often in managed forests, and they were rarely found in non-forested areas.

Mexican spotted owls generally nest in trees, although in the northern part of their range (southern Utah and Colorado) they often nest in caves or cliff ledges in canyons, and seem to prefer shady habitat with steep cliffs and rocky terrain (Willey 1998; Rinkevich et al. 1995). Vegetative components of their habitat vary spatially, with owls in the northern part of their range typically residing in forests dominated by mixed-coniferous species such as Douglas-fir (Pseudotsuga menziesii), white fir (Abies concolor) and ponderosa pine in conjunction with broadleafed species such as Gambel oak, maples (Acer spp.), boxelder (Acer negundo) and New Mexican locust (Robinia neomexicana) (Rinkevich et al. 1995; Ganey et al. 1992). In the southern part of their range (southern Arizona and Mexico), these owls typically occur in Madrean pine/oak forests dominated by Chihuahua pine (Pinus leiophylla var. chihuahuana), Apache pine (Pinus engelmannii) and southwestern white pine (Pinus flexilis var. reflexa) in conjunction with evergreen oaks, Douglas-fir, ponderosa pine and Arizona cypress (Cupressus arizonica) (Rinkevich et al. 1995; Ganey et al. 1992; Kroel and Zwank 1991). A few individuals have been observed to migrate downslope in the winter to Pinyon/Juniper (Pinus edulis/Juniperus spp.) dominated habitats (Ganey et al. 1992).

Within the United States portion of their range, Mexican spotted owl nest and roost sites have primarily been found in mixed-conifer stands, occasionally in pine/oak stands and very rarely in pine, evergreen oak, pinyon/juniper, riparian or other stands. Ganey and Dick (1995), reviewing Mexican spotted owl inventory data gathered between 1990 and 1993 from various parts of the owls’ range, found that 69% - 100% of nest sites were located in mixed-conifer stands, 0% - 28% were in pine/oak stands and 0% - 2% were in pine stands. Roost sites were similarly distributed, with 49% - 99% in mixed-conifer, 0% - 36% in pine/oak and 0% - 2% in pine.

Spotted owls are primarily nocturnal predators. Diurnal observations of Mexican spotted owls in Utah revealed that the owls spent 90% of the observation time roosting quietly (Willey 1998). Although spotted owls primarily hunt at night, they will opportunistically take prey who wander near their roost in the daytime. Nesting owls are significantly more likely to take prey in the daytime than non-nesting owls (Sovern et al. 1994).

They cannot hunt efficiently while flying (Carey et al. 1992), so they generally use a "sit-and-wait" hunting strategy in which they wait for prey animals to come near their location. Structurally complex forests with high vertical diversity are well-suited for this type of hunting strategy because they provide perches from ground level to the upper canopy (Carey et al. 1992). 

Threatened Status: Both the Mexican and the northern spotted owl (S. o. caurina) have been listed as threatened by the US Fish and Wildlife Service, in part due to the historical and ongoing alteration of their habitat from timber harvesting (Block 1994; USDI Fish and Wildlife Service 1995). Restriction of timber harvesting and the ensuing economic, social and political impacts have prompted a great deal of research on this species. Following the listing of the Mexican spotted owl on April 15, 1993, the USDI Fish and Wildlife Service appointed a recovery team and gave them the task of compiling the current knowledge on the owl and developing a plan to recover the owl. This recovery plan was released in 1995 (USDI Fish and Wildlife Service 1995).

Although we may know more about the ecology and status of the spotted owl than we do about any other threatened or endangered species (Gutierrez 1994), much of the research has focused on the northern spotted owl. Comparatively little research has been conducted on the Mexican spotted owl, particularly in regard to this bird’s response to fire (Howe et al. 1992). Habitat use, habitat distribution, and threats differ between the northern and the Mexican subspecies (USDI Fish and Wildlife Service 1995). Based on genetic analysis, Barrowclough and Gutierrez (1990) even suggested that the Mexican spotted owl may be a separate species from the northern and California spotted owls (S. o. occidentalis), pointing out that there has been no gene flow between them for at least 7,000 years.

Stand-maintenance vs. Stand-Replacement Fires: Fires vary in their intensity, duration and size, based on fuel availability, vegetative conditions, topography, climate, temperature, weather conditions and attempts to suppress the fire (Wenger 1984). Given these factors, fire effects on ecosystems can be viewed over a continuum, ranging from small-scale low-intensity fires such as a single lightning-struck snag, to large-scale high-intensity fires such as those that burned a third of Yellowstone National Park in 1988.

Fire effects are often categorized according to the impacts the fire has on the ecosystem. In the interests of drawing practical conclusions from my research, I will follow the fire regime delineation described by Wenger (1984) in which fires are separated into either stand-replacement fires or stand-maintenance fires. Stand-replacement fires, often referred to as "catastrophic" fires, are characterized by moderate- to high-intensity fire activity that kills practically all vegetation within the fire boundary. The dead vegetative material left after the fire often creates an additional fuel hazard, leading to increased fire danger in the future. Stand-maintenance fires include low- to moderate-intensity fire activity which generally burns low to the ground and mainly affects grasses, shrubs, forbs and small trees. This type of fire typically burns off accumulated vegetative debris on the ground without killing larger trees, and thus reduces the danger of future fires without causing major impacts on the current vegetative composition of the area (Wenger 1984).

I have further subdivided stand-maintaining fires into two categories. Canopy-level fire is that which burns into the canopy of some trees but does not cause complete mortality of all the trees in the area. Surface-level fire is that which burns only along the ground and never reaches the tree canopy.

DeBano et al. (1998) and Pyne et al. (1996) differentiate between the term Surface fire, meaning fire that only burns the litter, debris and small plants on the surface of the soil, and Ground fire, meaning fire that actually burns down into the organic material in the upper soil layer. This is a valid distinction to make because the effect on soil, microorganisms and root systems can be radically different between these two types of ground-level fire. Depending upon the depth, density, inorganic content and moisture content of the duff layer, this covering of decomposing plant material can either insulate the mineral soil from the heat of the fire or it can combust in a smoldering reaction that can potentially do great damage to the living material in the soil.

DeBano et al. (1998) and Pyne et al. (1996) and Agee (1993) discuss at length the phenomena of vegetative mortality resulting from exposure to heat, each pointing out that exposure to temperatures above 60° C (140° F) is generally lethal to the plant. Smoldering duff does not produce flames so it does not reach the intense heat of actively flaming vegetative material, but it continues to smolder, and generate heat, for far longer than flaming fires burn. If flaming debris ignites the duff and environmental conditions are such that the smoldering reaction can sustain itself, smoldering duff can raise the temperature of the underlying mineral soil to 300° C (570° F) for several hours at a time, reaching temperatures as high as 600° (1100° F) and producing lethal temperatures down as far as 9-16 cm in dry soil and 40-50 cm in moist soil (DeBano et al. 1998; Pyne et al. 1996). This type of ground fire can cause a great deal of mortality in the soil, including tree mortality when the roots are killed or the tree is girdled at ground level.

Both Ground and Surface fire appear to be stand-maintaining in the mixed-conifer or ponderosa pine-dominated forests used by Mexican spotted owls. Grier (1989) found a general decrease in biomass production as a result of prescribed fire in a ponderosa pine forest in northern Washington, and Swezy and Agee (1991) found individual ponderosa pine mortality caused by fine-root mortality in shallow soils and crown scorch in Oregon, but this mortality was restricted to seedlings, immature trees and weakened senescent groups and actually served to maintain a vigorous overall stand structure. Kalabokidis and Wakimoto (1992), in a Montana study of prescribed fire in ponderosa pine and Douglas-fir forests, found no significant differences in duff depth between burned and unburned sites and little mortality in larger trees following prescribed fires. Ryan and Frandsen (1991) conducted experimental burns on the duff layers surrounding 19 large ponderosa pine trees in Montana and found lethal heating in 45% of cambium samples tested and the subsequent death of 4 of the 19 trees.

In essence, it appears that whether a fire burns into the soil or not does not generally cause differences in stand structure on a scale large enough to make a significant difference in spotted owl presence or reproduction. These studies, plus the fact that I visited some burned territories 3 years after the fire when I could not readily distinguish whether a ground-level fire was a ground or surface fire, led me to consider all ground-level fires as surface fires for the purposes of this study.

Changes in Forest Composition since Presettlement Conditions: Forests have changed drastically over the last 120 years due to management practices such as timber harvesting, grazing and fire suppression (Fulé et al. 1997; USDA Forest Service Southwestern Region 1995; Kolb et al. 1994; Sackett et al. 1994; National Commission on Wildfire Disasters 1994; Covington and Moore 1994b, 1994b; Moody et al. 1992; Harrington and Sackett 1990; Wright 1990; Arno and Brown 1989). This idea is not new. Aldo Leopold, writing in 1924, described the encroachment of Pinyon and Juniper species into grasslands in the Tonto National Forest and the widespread buildup of brush species (oaks, manzanita [Arctostaphylos uva-ursi], mountain mahogany [Cercocarpus spp.] and ceanothus [Ceanothus spp.]) throughout many forests in Arizona, attributing both phenomena to fire suppression and grazing in the 40 years since settlement.

Estimates of presettlement fire frequency in southwestern forests vary somewhat, with the pinyon-juniper type burning approximately every 10-30 years (Wright 1990; Leopold 1924) the ponderosa pine type around every 1.8-12 years (Fulé et al. 1997; Covington and Moore 1994b; Swetnam 1990; Wright 1990; Dieterich 1980a, 1980b; Weaver 1951), and the mixed-conifer type around every 5-22 years (Wright 1990; Ahlstrand 1980). Fires that burn at these frequencies (at around 2-25 yr. intervals) tend to be stand-maintenance fires rather than stand-replacement fires (Wenger 1984), and thus their main impact on the forest is to burn the ground-level fuel (woody debris, grasses, forbs, shrubs and small trees) while harming few of the larger trees.

Since fire suppression, cattle grazing and timber harvest began in the Southwest late in the 19^th century, southwestern ponderosa pine forests have shown an increase in stand density, higher fuel loads, greater canopy closure, increased vertical fuel continuity, decreased vegetative decomposition rates and decreased fire frequency, all of which increase the potential severity and destructiveness of fires (Zwolinski 1990; Covington and Moore 1994a, 1994b). Crown fires, for example, are now common occurrences, and yet they were once almost unknown in the Southwestern ponderosa pine forest type (Covington and Moore 1994b).

Ironically, historic fire suppression has probably had the greatest impact on current fire danger. When fires began to be actively suppressed, ground-level fuel began to accumulate. The dry southwestern climate aided this fuel buildup by inhibiting decomposition (National Commission on Wildfire Disasters 1994), and thus fuel loads have been growing faster than they can decay. The small trees that would normally have been killed in their first few years have instead grown into densely-packed stands of saplings and pole-sized trees (Covington and Moore 1994b; Covington et al. 1994), creating fuel ladders that carry the fire to the crowns of the larger trees.

Heavy grazing throughout the century has also contributed to the problem by reducing the grass and forb layer that would normally carry ground fire (National Commission on Wildfire Disasters 1994; Wright 1990). Elimination of the grass and forb layer also promotes the establishment and growth of tree seedlings by removing potential competition, eventually leading to the development of the dense sapling and pole-sized stands (Sackett et al. 1994).

Prescribed Fire, Prescribed Natural Fire and Wildfire: The origin of the fire may also play a role in how that fire affects the forest. Due to the previously described management activities over the past 120 years, wildfires tend to be far more intense and destructive than they were under presettlement conditions. However, wildfires often have the advantage of occurring during the natural fire season (primarily during the monsoon season between July and September [Fulé et al. 1997; Sackett et al. 1994] and to a lesser extent in late spring between late April and June [Fulé et al. 1997]), and thus burn plants at a time of year in which the plants have likely evolved mechanisms to cope with fire-induced damage. This advantage is, unfortunately, offset by the current unnaturally high fuel loads.

Prescribed burns, on the other hand, are typically conducted under wet or cool conditions when there is little chance that the fire will turn into a large-scale stand-replacing fire. These conditions usually enable the land managers to control the fire and to accomplish specific management goals, and thus prescribed fire has become a very powerful and useful tool. Wildlife managers have found prescribed fire useful for creating diversity in habitat structure by breaking up homogeneous cover types (Severson and Rinne 1990).

The drawback to most prescribed fires, however, is that they rarely occur during the natural fire season. The natural fire season is typically a time at which the forest is in a highly combustible state and forest managers are often reluctant to start fires in areas with both extremely heavy fuel loads and highly combustible conditions. To reduce the threat of losing control of the fire, managers will often conduct prescribed burns under cooler, wetter conditions which generally occur outside of the natural fire season. For example, Harrington and Sackett (1990) recommend that prescribed burns in areas that have not been subjected to fire in decades should be conducted in the fall or early spring when temperatures and humidities are moderate. However, this off-season burning can have significant impacts on vegetative structure and species composition. Zwolinski (1990) points out that the season in which the fire occurs is an important factor in plant survival and reproduction, and Harrington and Sackett (1990) discuss seasonal variation in tree susceptibility to fire. DeBano et al. (1998) describe how moist soils conduct heat better than dry soils, and in cases of long-smoldering duff fires can carry lethal temperatures as deep as 50 cm below the surface. Lethal temperatures in dry soils rarely penetrate deeper than 16 cm (DeBano et al. 1998).

A recent development in fire management is the concept of the "prescribed natural fire." This refers to prescribing a fire for a certain area and waiting for a fire to start and burn there naturally. The area is typically prepared beforehand by prescribed fire or mechanical thinning in order to reduce the chance of catastrophic wildfire, and the prescribed natural fire then has the advantage of burning during the natural fire season while accomplishing specific management goals (W. Block, pers. comm. March 15, 1996).

Few researchers have measured the effects of fire or fire suppression on any aspect of Mexican spotted owls. Some have speculated that spotted owls were not even present in many of their current areas prior to European settlement, and that the owls only moved in after forest management practices altered the landscape (National Commission on Wildfire Disasters 1994). Mexican spotted owls in the Gila National Forest have been observed to return to their territories after prescribed natural fires, provided that the stand structure remained intact (USDA Forest Service Southwestern Region 1995). Some California spotted owls apparently disappeared for several years following a highly destructive fire in 1977 (Elliott 1985). Gaines et al. (1997) describe some impacts of 1994 wildfires on 6 northern spotted owl activity centers in eastern Washington, noting a decrease in the number of reproductive pairs on these sites (although not much below the numbers in previous low years) and an increase in the number of unoccupied sites the year after the fires. Two pairs of radio-tagged northern spotted owls in south-central Washington stayed near their territories after wildfire but shifted their primary activity to lightly burned or unburned areas (Bevis and other 1997). One female owl in this study was found dead in an emaciated condition 2.5 months after the fire, leading to speculation that the fire may have damaged her prey base. Her mate disappeared over the winter and two new owls occupied the territory in 1995.

Effects of Fire on Owl Prey: Fire could affect Mexican spotted owls indirectly through their prey base. Spotted owls may select habitats partially based on prey availability (Ward and Block 1995; Verner et al. 1992), so fire-caused changes in prey populations could potentially alter the quality of the habitat.

The Mexican spotted owl recovery team reviewed a data set of 11,164 prey items collected from 18 geographic areas within the owls’ range (Ward and Block 1995). Ward and Block found that owl diet varied across the owls’ range, and owl reproductive success was not influenced by the presence or abundance of any particular prey species. They hypothesized that owl reproductive success was, therefore, influenced primarily by the total prey biomass consumed rather than the presence or abundance of any particular species. However, unpublished information suggests that the reproductive success of spotted owls in the Sacramento Mountains of southern New Mexico was positively correlated with the abundance of deer mice (Peromyscus maniculatus) (Ward et al. [unpublished], cited in Ward and Block 1995).

Ward and Block found eight prey groups that comprise significant portions of the Mexican spotted owl diet (Table 1). Peromyscid mice, woodrats, microtine voles and birds each represented >= 10% of both the relative frequency and the total biomass of the owls’ diet in at least one of the geographic recovery units delineated by the Mexican spotted owl recovery team. Bats and arthropods were taken in high numbers, but they have little mass and, therefore, did not represent >= 10% of the total biomass. Rabbits and pocket gophers, on the other hand, were taken relatively rarely, but they are larger animals and represented a relatively large proportion of total prey biomass.