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Division of Biological Sciences, University of Montana, Missoula, Montana 59812
Received for publication June 30, 1998. Accepted for publication December 1, 1998.
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ABSTRACT |
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 TOP ABSTRACT INTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONLITERATURE CITED |
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Key Words: Caryophyllaceae • demography • fire effects • population dynamics • prolonged dormancy • Silene spaldingii.
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INTRODUCTION |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONLITERATURE CITED |
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; Pickett and White, 1985
; Huston, 1994
). Disturbances often remove dominant competitors, temporarily providing adequate water, nutrients, and light for establishment (Grubb, 1977
; Grime, 1979
). Many plant species are adapted to regular disturbance regimes and may even require disturbances such as fire, windthrow, or grazing to persist (Denslow, 1980
; Pickett, 1980
; Watt, 1981
; Sousa, 1984
). Disturbance also plays an important role in the persistence of many rare or endangered species (Parsons and Browne, 1982
; Bowles et al., 1990
).
Fire is an important disturbance in many grassland, shrubland, and forest ecosystems in North America and throughout the world (Kozlowski and Ahlgren, 1974
), and its effects vary among these systems. Fire can remove much or all aboveground biomass, increasing light penetration to the soil surface and often raising surface temperatures (Hulbert, 1969
, 1988
; Peet, Anderson, and Adams, 1975
). Combustion of live and dead fuels or removal of litter layers may result in enhanced nutrient availability (Raison, 1979
; Knapp and Seastedt, 1986
; Dudley and Lajtha, 1993
; Brewer, 1995
). Fire may also stimulate germination of species adapted to recently fire-disturbed sites (Daubenmire, 1968
; Glenn-Lewin et al., 1990
). The effects of fire vary among plant species, some responding positively and others negatively (Daubenmire, 1968
; Vogl, 1974
; Howe, 1995
).
Numerous plant responses to fire in grasslands have been documented. These include increased flowering, enhanced vegetative vigor and productivity, increased as well as decreased mortality, and increased germination and recruitment (Daubenmire, 1968
; Vogl, 1974
; Glenn-Lewin et al., 1990
). However, usually only one or two important life history parameters have been studied in the same species at the same time. Only rarely have the demographic consequences of fire been studied by following marked individuals (Hartnett and Richardson, 1989
; Silva et al., 1991
; Paige, 1992
; Menges and Dolan, 1998
), and most such studies were carried on for only 1 or 2 yr. However, a knowledge of the effects of disturbances such as fire on plant demography is required to understand population-scale phenomena that lead to changes in community structure (Hartnett and Richardson, 1989
).
The season in which fire occurs may affect how plant species and communities respond (Abrahamson, 1984
; Biondini, Steuter, and Grygiel, 1989
; Howe, 1995
; Cobb, Gori, and Whitham, 1996
). In the tallgrass prairie region late-season fires retard the vigor of warm-season species, allowing cool-season species to increase, while early-season fires have the reverse effect (Howe, 1995
). Early and late dormant-season fires are commonly used in natural areas management (Steuter, Grygiel, and Biondini, 1990
) and may have different effects on vegetation. The demographic effects of fire season on individual species have been little studied, and seasonal differences in fire effects are not known for grasslands of intermountain western North America.
Fire is a frequently used tool for managing grassland natural area preserves (Loucks, 1968
; Cole, Klick, and Pavlovic, 1992
; Howe, 1994
), and may also be prescribed for managing populations of rare plants (Bowles et al., 1990
; Hessl and Spackman, 1995
; Grigore and Tramer, 1996
). However, not all plant species respond favorably to fire, so fire effects on rare or endangered species must be understood before fire management plans for natural areas are implemented (Jacobson, 1991
). Silene spaldingii Wats. is an endangered plant endemic to intermountain grasslands of the Pacific Northwest. The largest known population of S. spaldingii occurs in northwest Montana on a preserve owned by The Nature Conservancy. Prescribed fire is one of the management options on this preserve. The purpose of my study was to determine the effect of fire and season of fire on recruitment, survivorship, and reproduction of Silene spaldingii.
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MATERIALS AND METHODS |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONLITERATURE CITED |
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27 km south and 75 m higher, mean annual precipitation was 438 mm for 1950–1980. Mean July maximum and January minimum temperatures were 27.9° and -11.4°C, respectively. Climate on the preserve is warmer and drier. The preserve currently consists of 600 acres of fescue-wheatgrass-needlegrass grasslands developed on rolling glacial topography. Grasslands of western Montana are transitional between Palouse Prairie typical of eastern Washington and Oregon and grasslands of the Northern Great Plains (Antos, McCune, and Bara, 1983
). Silene spaldingii occurs in the bottom of shallow swales and on cool slope exposures with relatively deep soil. Common grasses in these habitats are Festuca scabrella, F. idahoensis, and Poa pratensis (Lesica, 1997
) Of these, F. scabrella is the most sensitive to fire (Mueggler and Stewart, 1980
). Antennaria microphylla, Arnica sororia, and Hieracium cynoglossoides are common forbs (nomenclature follows Hitchcock and Cronquist, 1973
).
Fires ignited by lightning as well as Native Americans played an important role in structuring the vegetation of the Tobacco Plains by preventing the establishment of woody vegetation (Koterba and Habeck, 1971
; Dorey, 1979
; Barrett and Arno, 1982
). Mean presettlement fire-return interval for western Montana valleys was 9 yr (Barrett and Arno, 1982
) and was estimated to be 6.4 yr (range of 2–13 yr) for the north end of the Tobacco Valley in southeast British Columbia (Dorey, 1979
).
Species description
Silene spaldingii Wats. (Caryophyllaceae) is a perennial, geophytic herb, 20–40 cm tall, arising from a simple or branched caudex surmounting a long, slender taproot. Rhizomes or other means of vegetative propagation are lacking (Hitchcock and Maguire, 1947
; Lesica, personal observation). Flowers, borne in a branched, terminal inflorescence, bloom in July and set seed in August. Seeds will germinate with as little as 4 wk of cold treatment (Lesica, 1993
), so germination likely occurs in fall as well as spring. Rosettes are formed the first year, after which vegetative stems are produced. New recruits may appear as vegetative stems in their first summer if they germinated the previous fall (Lesica, personal observation). Existing plants send up new vegetation in mid-May and become senescent by early September (Lesica, 1997
). Plants may become reproductive in their second season, but most plants flower for the first time when 2 yr or older (Lesica, 1997
). Seeds are dispersed by being shaken from an orifice on the top of the mature ovary.
Silene spaldingii plants may go undetected for one to several years but reappear in subsequent years (Lesica and Steele, 1994
; Lesica, 1997
). At Dancing Prairie 10% of plants recorded as summer dormant produced small leaves that senesce and disappear by early July, while the remainder produced no aboveground vegetation during the entire growing season (Lesica, 1997
). Prolonged dormancy can be inferred by following the fate of marked or mapped individuals for numerous years. The presence of summer-dormant plants makes exact estimation of many demographic parameters difficult. At Dancing Prairie a mean of 41% of S. spaldingii plants exhibited prolonged dormancy each year in 1989–1994. Of the 193 episodes of prolonged dormancy recorded during this period, 75% were 1 yr in duration and 90% were either 1 or 2 yr long (Lesica, 1997
). Thus, 41% of S. spaldingii plants could not be detected in the first and last years of the study (1991, 1996); 10% of plants could not be detected in the second and second to last years (1992, 1995); and only 4% (10% x 41%) of plants were undetected in 1993–1994 (Lesica, 1997
).
Silene spaldingii is endemic to the Palouse region of southeast Washington and adjacent Oregon and Idaho and is disjunct in northwest Montana (Hitchcock and Maguire, 1947
). Much of the habitat of S. spaldingii has been lost to agricultural development. Although once widespread in the Palouse region, S. spaldingii is now known from mainly isolated sites on the periphery of its former range. Most remaining populations are small and threatened by exotic weed encroachment, livestock grazing, and herbicide treatment. Silene spaldingii is listed as threatened or endangered in all four states in which it occurs (Lesica and Shelly, 1991
) and is currently being considered for listing under the Federal Endangered Species Act.
Field methods
In July 1991 I established 30 113-m2 (6 m radius) circular plots in the north end of the preserve where S. spaldingii plants are most common. Plot centers were permanently marked by driving a metal rod into the ground and affixing a tag with the plot number. I located plots so that each would contain at least ten plants of S. spaldingii; otherwise plot location was haphazard. In each plot I recorded the location of each plant of S. spaldingii by measuring the distance from the center post to the nearest 1 cm and the compass bearing from the plant to the post to the nearest degree. For each annual census I recorded the growth form, reproductive status, and fecundity of each S. spaldingii plant using the following classification: rosette—lacking any visible stem elongation; dormant—no aboveground parts observed; vegetative—one or more stems present but no flowers; and reproductive—one or more fertile stems present. The number of flowers was recorded for each reproductive plant.
I employed a stratified-random design to assign the plots to the three treatments: fall burn, spring burn, and control. I ordered the plots according to the number of S. spaldingii plants they contained and then divided the 30 plots into ten consecutive groups of three. One plot from each group was assigned to each of the three treatments using a random numbers table.
I divided each plot into four equal cells along the four principal compass bearings for the purpose of mapping and censusing. When several S. spaldingii plants occurred in close proximity, making individual identification uncertain, I eliminated the cell in which they occurred from analysis of that plot for all years. I eliminated 11 such cells: five fall burn, four spring burn, and two control. Eliminating 9% of the cells with high density of S. spaldingii plants may introduce a small bias if plants behave differently in these areas; however, this bias is minimal because cells were eliminated from all three treatments with approximately equal frequency.
The fall burn treatment was carried out mid-September 1991, and the spring burn was performed in mid-April 1992. These are the two times of year when prescribed fire is feasible on the preserve. A border 1 m wide was mowed around each plot, and the debris was raked clear. Fires were started with a drip torch and extinguished after the entire plot had been burned. Experimental plots were small compared to the size of natural or prescribed fires. However, plot size was large compared to the presumed mean distance of passive dispersal (van der Pijl, 1982
), suggesting that demographic parameters in burn plots should approximate those of larger burns.
Annual censusing of Silene spaldingii plants was conducted in mid-July 1991–1996, when the plants were flowering. At the same time in 1992 I also recorded observations on the severity of burns in the treatment plots based on the amount of litter remaining. I used these observations to place burns into one of two severity classes: high and medium. A high-severity burn left little litter; mineral soil was apparent throughout most of the plot. In a medium-severity burn thin layers of fine litter and ash were apparent over at least half of the plot. There were no low-severity burns with coarse fuels remaining in any of the treatment plots. In 1996 I measured the depth of plant litter to the nearest 1 cm at four points, 3 m from plot center along primary compass bearings, in each plot.
Data analysis
Repeated-measures analysis of variance was used to test hypotheses regarding recruitment, survivorship, proportion of reproductive plants, flowering, proportion of dormant plants, and total density of Silene spaldingii. Each analysis for multiyear post-burn treatment effect was followed by contrast tests for a difference between burn and control treatments and between the two burn treatments. For each demographic parameter I performed one-way analysis of variance (ANOVA) for separate years in order to determine in which years the effect was present. Count and proportion variables were square-root and arcsine transformed, respectively, to conform to the assumptions of the tests (Sokal and Rohlf, 1981
, pp. 421–428). Plot means for flowers per reproductive plant did not require transformation prior to analysis. The 1991 preburn recorded plot densities, unbiased with respect to treatment, were used as covariates in analysis of recruitment, survivorship and total plant density. The 1991 plot means were used as a covariate in the analysis of flowering. A probability level of P = 0.05 was used to assign statistical significance and was not adjusted for multiple tests (Stewart-Oaten, 1995
). Untransformed plot means for litter depth were analyzed with ANOVA.
With rare exceptions, all rosette S. spaldingii plants are new recruits based on results of a long-term demography study begun in 1987 (Lesica, 1997
). Plants in the vegetative size class may be new recruits as well, but they may also be older (Lesica, 1997
). For example, a vegetative-size plant first appearing in 1993 may have been a new recruit, but it could also be an older plant that was dormant in 1991–1992. On average, only 4% of S. spaldingii plants remained undetected after two years of recording at the site (Lesica, 1997
). Thus, after 1993 96% of all plants recorded for the first time were new recruits. Only rosette plants were counted as recruits in 1991–1993, while all vegetative plants recorded for the first time were also considered recruits in 1994–1996.
Annual estimates of total plant density and proportion of dormant S. spaldingii plants were confounded by the inability to detect all plants in all years (see above). By eliminating the initial and final 2 yr of data, acceptable levels of accuracy were achieved; 1993–1994 estimates were 96% accurate (Lesica, 1997
). The proportion of summer-dormant plants was measured as the number of dormant plants divided by the total minus the number of recruits.
Only plants that failed to reappear for at least three consecutive years were assumed dead because 96% of summer-dormant plants on average are detected in the 3rd yr of recording (Lesica, 1997
). Number of surviving S. spaldingii plants observed in 1991 was analyzed for 1992–1994. Survival of cohorts recruited in 1992 and 1993 was analyzed through 1994. Survivorship of combined plot totals for the 1991 sample population and the two cohorts was also analyzed using the nonparametric logrank test (Pyke and Thompson, 1986
; Hutchings, Booth, and Waite, 1991
). Statistical analyses were performed on a microcomputer using SYSTAT (Wilkinson, 1986
).
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RESULTS |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONLITERATURE CITED |
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Depth of litter in 1996 was 8.7, 9.6, and 11.1 cm in fall- and spring-burn plots and control plots, respectively, and there were no differences between treatments and control (F1,27 = 1.9, P = 0.18) or between fall- and spring-burn treatments (F1,27 = 0.3, P = 0.56).
Recruitment
Recruitment of Silene spaldingii was sporadic. No recruitment (i.e., rosette plants) was detected in treatment or control plots in 1991, the year preceding the burn treatment. Recruitment was high only in 1993 in control plots, while in burn plots it was high in both 1992 and 1993 (Fig. 1A). During 1992–1996 mean number of recruits per plot was significantly higher in burn plots compared to controls (P < 0.001, Table 1). Burn plots averaged more recruits than controls in 1992, 1993, and 1994, and differences were significant in 1992 and 1994 and marginally significant (P = 0.08) in 1993 (Fig. 1A). There were no differences between treatment and control plots in 1995 and 1996, the final 2 yr of the study.
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Survivorship
Survivorship could be accurately estimated only through 1994. Survivorship of adult plants was high with 80% survival between 1991 and 1994. There was no difference in survival of the S. spaldingii plants recorded in 1991 between burn plots and controls in 1992–1994 (F3,78 = 0.7, P = 0.49; logrank test 
2 = 0.31, P = 0.58; Fig. 2A). Survivorship of new recruits was appreciably lower than for adults, with nearly 50% mortality in the first 2 yr. There appeared to be a tendency for higher 2-yr survival of the cohort recruited in 1992 in control plots (Fig. 2A); however, this difference was not significant (F2,52 = 0.56, P = 0.58; 2 = 0.22, P = 0.64), possibly due to the small number of control recruits. There was no difference in 1-yr survival of the 1993 cohort between treatment and control plots (F1,26 = 0.1, P = 0.74).
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Prolonged dormancy
I was able to accurately estimate the proportion of summer-dormant S. spaldingii plants per plot only in 1993 and 1994. The proportion of dormant plants varied widely between these 2 yr: 13–30% in 1993 and 46–79% in 1994. Control plots had significantly more dormant plants than burn plots in 1993–1994 (F2,54 = 15.9, P < 0.001). However, there were significantly more dormant plants in burn plots in 1993 (F1,26 = 7.7, P = 0.01), but significantly more in control plots in 1994 (F1,26 = 13.8, P = 0.001; Fig. 1B).
There were no differences in summer dormancy between spring and fall burns over these 2 yr (P = 0.55).
Reproduction
The mean proportion of recorded S. spaldingii plants in reproductive condition varied between 0.45 and 0.52 in 1992 and between 0.79 and 0.82 in 1995 (Fig. 1C). The mean proportions of reproductive plants in treatment and control plots were nearly equal in 1991, before the burns (F1,27 = 0.1, P = 0.77), but there were more reproductive plants in control plots in 1992–1996, although the difference was only marginally significant (F5,130 = 2.2, P = 0.055). The difference was primarily due to a 20% higher proportion of reproductive plants in control plots in 1993 (F1,26 = 4.25, P = 0.049; Fig. 1C). Treatment and controls were not different in any other year following the burn (P > 0.47, Fig. 1C).
The mean proportion of reproductive plants was not different between fall- and spring-burn treatments in 1991 (F1,27 = 1.3, P = 0.26) or in 1992–1996 (F5,130 = 0.5, P = 0.76).
The mean number of flowers per reproductive S. spaldingii plant per plot in 1991 was not significantly different among treatments (F1,27 = 1.27, P = 0.27). Following the burn in 1992–1996, burn plots had significantly more flowers per plant than controls (P = 0.004, Table 2). The trend for greater flowering in burn plots was marginally significant in 1992 and 1993 (P = 0.09) and was no longer apparent after 1994 (Fig. 1E). There was no difference in flowering between fall and spring burns in 1992–1996 (P = 0.76).
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DISCUSSION |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONLITERATURE CITED |
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). Interference from litter accumulation has been demonstrated in other species and is thought to be the cause of reduced diversity in some grassland systems (Tilman, 1993
; see reviews in Vogl, 1974
; Facelli and Pickett, 1991
). For some species only germination was inhibited (Werner, 1975
), while for others litter interfered with both germination and seedling survival (Goldberg and Werner, 1983
; Hamrick and Lee, 1987
; Bergelson, 1990
). Reduction of the litter layer by grazing or fire is expected to result in increased seedling germination and/or establishment and is the most likely cause of the higher rates of recruitment observed in burn plots in this study. Grassland fires may also result in warmer soil temperatures and increased available nutrients (Raison, 1979
; Knapp and Seastedt, 1986
; Hulbert, 1988
) which could cause enhanced germination and recruitment (Daubenmire, 1968
; Glenn-Lewin et al., 1990
).
The effect of fire on the reproductive capacity of S. spaldingii was equivocal. Over the period of 1992–1996, the proportion of reproductive plants in spring burn and control plots was nearly equal, but was lower following the fall-burn treatment. The number of flowers per reproductive plant was greater in burn plots, but this trend was apparent for only 2–3 yr following the treatments. It is not known whether the increase in flowering translated into an increase in seed production. These effects of fire on reproductive capacity did not likely have an effect on recruitment in 1992–1993. However, an increased seed set might affect recruitment in the future, depending on seed bank dynamics, disturbance regime, and weather. Increased flowering and seed production have often, but not always been reported for grassland species following fire (Daubenmire, 1968
; Vogl, 1974
). Silene regia, a congener of eastern North American grasslands, responds to fire with increased density of reproductive stems and number of flowers per stem (Menges, 1995
; Menges and Dolan, 1998
).
Fire had no detectable effect on the survival of adults or recruits of Silene spaldingii. Some studies in grassland systems have found that fire can increase size and survival of grasses and forbs (Hartnett and Richardson, 1989
; Silva et al., 1991
), while others have found that survivorship decreased (Paige, 1992
; Cobb, Gori, and Whitham, 1996
). The effect of fire on a species often depends on the season in which the burn occurs (Howe, 1995
), and this may explain some of the variation in results reported in the literature. In my study, neither burn treatment had a significant effect on survivorship, probably because they occurred before and after the plant's active growing season. In northwest Montana, lightning-caused fires would have occurred most frequently during July and August (Barrows, 1951
), although some probably occurred in spring and fall as well (S. Arno, U.S. Forest Service, Intermountain Research Station; personal communication). Summer fires would likely have a more adverse effect on S. spaldingii because plants are actively growing and flowering at that time.
Experimental demography studies may be necessary for fully understanding population dynamics. I was able to pinpoint the significant effects of fire on Silene spaldingii by following the fate of individual plants for five summers following burn treatments. Studies in which plots are simply censused following treatments cannot distinguish between changes in survival and recruitment. Furthermore, studies documenting enhanced recruitment following fire without determining recruit survivorship may not accurately predict the ultimate effect of the treatment if survival is low. Herbaceous geophytes that demonstrate summer subterranean dormancy present a particularly vexing problem when studying the effects of fire. Changes in the density of detectable plants may or may not indicate real changes in population size, especially since fire may affect the proportion of summer dormant plants in complex ways.
Effects of fire season
Season of fire can be important to the effects of burn treatments on herbaceous plants (Howe, 1995
; Cobb, Gori, and Whitham, 1996
), and my study demonstrated a between-fire season effect on recruitment in S. spaldingii. A greater amount of residual fine litter in spring- compared to fall-burn plots indicated more severe fires for the fall burns, probably due to lower fuel moisture at that time of year. Fall-burn plots had lower Silene spaldingii recruitment in 1992, resulting in lower recruitment for the entire study period. The thin layer of fine litter remaining after the cooler spring burn may have provided better safe sites for germination and survival. Whatever the cause, differences between the two treatments did not ultimately translate into a significant difference in mean density of S. spaldingii.
Prolonged dormancy
The proportion of summer-dormant plants was lower for burn (0.38) compared to control (0.46) plots in 1993–1994. However, the number of dormant plants was higher for burn plots in 1993 but higher in controls in 1994 (Fig. 1B). In addition, the between-year variation in prolonged dormancy was smaller in burn plots compared to controls. At Dancing Prairie Preserve prolonged dormancy of S. spaldingii demonstrated biennial cycles in 1989–1994, with a higher proportion of dormant plants occurring in even-numbered years (Lesica, 1997
). It appears that fire had a damping effect on this cycle, raising the low proportion of dormant plants in 1993 and lowering the higher proportion in 1994.
Understanding the mechanism responsible for fire-induced change in rates of prolonged dormancy is problematic because the causes of prolonged summer dormancy are not understood. Environmental stresses such as drought or flooding have been invoked to explain high levels of prolonged dormancy (Lesica and Steele, 1994
), but evidence is generally anecdotal. Winter temperature and amount of summer precipitation were positively correlated with the proportion of summer dormant S. spaldingii plants the following year at Dancing Prairie in 1989–1994 (Lesica, 1997
). However, it is difficult to imagine a mechanism to explain these correlations. The biennial cycling of prolonged dormancy of S. spaldingii plants suggests that an endogenous, physiological rhythm synchronized by an environmental cue could play a role (Lesica, 1997
). A high rate of summer dormancy for the flush of 1992 recruits in burn plots might partly account for higher treatment dormancy in 1993; however, the proportion of plants dormant in 1993 was nearly equal for 1992 recruits (38%) and nonrecruits (36%; 2 = 0.28, P = 0.60). Perhaps the warmer early spring ground temperatures or enhanced nutrient availability following fire induce a significant proportion of plants to shift their cycles by 1 yr.
Implications for conservation
Disturbances such as fire or grazing that reduce the quantity of litter in highly productive grasslands are likely to be important to the long-term persistence of Silene spaldingii. Grazing by large ungulates is thought to have been of minor importance in the evolution of grassland ecosystems in most of intermountain western North America (Mack and Thompson, 1982
), although wapiti (Cervus canadensis) may have had some impact in northwest Montana. On the other hand, fire was a common occurrence in these intermountain grasslands (Barrett and Arno, 1982
) and undoubtedly played a role in shaping presettlement plant communities (Koterba and Habeck, 1971
). Fire is a preferred management tool at Dancing Prairie Preserve for ecological and logistical reasons (B. Martin, The Nature Conservancy; personal communication). In the absence of fire and grazing, significant recruitment of S. spaldingii is episodic (Lesica, 1997
). My study has shown that fire can enhance recruitment and population density of S. spaldingii at Dancing Prairie Preserve. In the absence of grazing, fire may be required to reduce litter and allow adequate recruitment of S. spaldingii and other ecologically similar species. In eastern Washington S. spaldingii occurs in less productive grasslands in which substantial accumulations of litter do not occur. Fire may not demonstrate positive effects on S. spaldingii at these sites.
This study suggests that burning productive grasslands will have a positive effect on S. spaldingii whether prescribed in spring or fall. Silene regia, an eastern North American prairie plant, has also been shown to respond positively to fire (Menges and Dolan, 1998
). With regard to conservation of S. spaldingii, there is little reason to choose fall rather than spring burning, although recruitment may be somewhat higher following a spring burn.
Enhanced recruitment of S. spaldingii persisted for 3 yr following burn treatments. Depth of litter accumulation in burn plots reached control levels 5 yr after the treatments, although litter density still appeared higher in controls after this length of time. Presettlement fire return intervals were 6–9 yr (Dorey, 1979
; Barrett and Arno, 1982
). These results suggest that a fire interval of 5–10 yr would be appropriate for the preserve and would provide optimum growth rates of S. spaldingii populations in these habitats. Shorter intervals may not allow accumulation of sufficient fuel loads to carry a fire, and there would be little benefit to recruitment rates.
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FOOTNOTES |
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LITERATURE CITED |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONLITERATURE CITED |
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Grime, J. P. 1979 Plant strategies and vegetation processes. John Wiley & Sons, Chichester.
Grubb, P. J. 1977 The maintenance of species-richness in plant communities: the importance of the regeneration niche. Biological Review 52: 107–145.
Hamrick, J. L., and J. M. Lee. 1987 Effect of soil surface topography and litter cover on the germination, survival and growth of musk thistle (Carduus nutans). American Journal of Botany 74: 451–457.[CrossRef][ISI]
Hartnett, D. C., and D. R. Richardson. 1989 Population biology of Bonamia grandiflora (Convolvulaceae): effects of fire on plant and seed dynamics. American Journal of Botany 76: 361–369.[CrossRef][ISI]
Hessl, A., and S. Spackman. 1995 Effects of fire on threatened and endangered plants: an annotated bibliography. U.S. Department of Interior National Biological Service Information and Technology Report 2, Washington DC.
Hitchcock, C. L., and A. Cronquist. 1973 Flora of the Pacific Northwest. University of Washington Press, Seattle, WA.
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