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Ecology |
Department of Biology, University of Mississippi, University, Mississippi 38677 USA
Received for publication June 27, 2000. Accepted for publication December 22, 2000.
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ABSTRACT |
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 TOP ABSTRACT INTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONLITERATURE CITED |
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Key Words: bog • life history • matrix models • pitcher plants • population dynamics • sensitivity analysis
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INTRODUCTION |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONLITERATURE CITED |
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; Norquist, 1984
; Gilliam and Christensen, 1986
; Waldrop, White, and Jones, 1992
; Brockway and Lewis, 1997
; Maliakal, Menges, and Denslow, 2000
). Carnivorous plants are thought to be particularly vulnerable to competitive displacement in unburned seepage bogs and wet pine savannas of the southeastern coastal plain of the United States (Wells, 1928
; Wells and Skunk, 1928
; Eleuterius, 1968
; McDaniel, 1971
; Schnell, 1976
; Weiss, 1980
; Folkerts, 1982
; Gibson, 1983
; Barker and Williamson, 1988
). Previous studies have shown that foliage of pitcher plants (Sarracenia spp.) and seedling densities of pitcher plants and sundews (Drosera capillaris) increase dramatically following winter fires (Eleuterius, 1968
; Weiss, 1980
; Barker and Williamson 1988
; Brewer, 1999a, b
). A cost–benefit model (Givnish et al., 1984
) predicts that carnivorous plants would be at a competitive disadvantage in low-light environments as aboveground biomass and litter increase during prolonged intervals between fires (Givnish et al., 1984
). The benefit of producing carnivorous structures declines as growth becomes limited more by light than by the availability of nutrients (Givnish et al., 1984
). Thus, both theory and empirical observations would appear to support the idea that carnivorous plants benefit from frequent fires.
As important as frequent fires may be in maintaining species diversity in pine savannas, short-term increases in foliage and seedling densities following fire do not provide adequate proof that frequent fires are necessary to maintain populations of carnivorous plants. Fire-stimulated increases in foliage may indicate shifts in allocation rather than increases in growth or density (Brewer, 1999a
). Furthermore, just because individual fires increase seedling establishment does not mean that frequent fires are necessary to maintain viable populations. The viability and duration of dormant or semi-dormant stages may enable carnivorous plants to escape competition during fire-free intervals. Small, short-lived species such as Drosera capillaris and Utricularia spp. apparently escape competition by producing a persistent seed bank (Brewer, 1999b, c
; Maliakal, Menges, and Denslow, 2000
; J. S. Brewer, unpublished data). Long-lived perennial plants such as Sarracenia spp. may avoid competition by reducing allocation to costly pitchers during years without fire (Brewer, 1999a
). There are no studies of the real impact of fire-mediated changes in seedling establishment or pitcher production on pitcher-plant populations. And yet, numerous authors cite declines in the density of pitchers and seedlings with fire suppression as evidence that frequent fires are necessary to maintain populations of carnivorous plants (Eleuterius, 1968
; Weiss, 1980
; Folkerts, 1982
; Barker and Williamson, 1988
; Givnish, 1989
).
Ideally, one would use long-term controlled experiments to test the effects of fire frequency on the growth and viability of carnivorous-plant populations. Unfortunately, such data are not currently available. Another approach is to model long-term population dynamics using responses to one or a few individual fires. Although this approach has been used to model the effects of fire regime on pine and oak populations in xeric longleaf pine systems (e.g., Rebertus, Williamson, and Moser, 1989
), such an approach has not been used in hydric pine–savanna systems (e.g., pitcher-plant bogs). Matrix models can provide reasonable projections of population dynamics into the future, given a number of contingencies within realistic bounds (Caswell, 1989
). The validity of model results could be tested and the models refined as experimental data become available.
In this study, I examined the potential impact of fire frequency on population dynamics of Sarracenia alata (Wood) Wood (yellow pitcher plant, hereafter Sarracenia) in pitcher plant bogs in southeastern Mississippi, USA. I chose to study this species for three reasons. First, Sarracenia exhibits fire-stimulated seedling establishment (Barker and Williamson, 1988
; J. S. Brewer, Laws, and Mozingo, unpublished data). Second, others have argued that populations of Sarracenia decline in the absence of fire (Eleuterius, 1968
; Folkerts, 1982
; Barker and Williamson, 1988
). Third, Sarracenia alata is a common species that exhibits a morphology and life history similar to rare congeneric species of pitcher plants (e.g., S. leucophylla, S. oreophila). It therefore provides a model to guide studies of the ecology and natural history of rare species of pitcher plants.
The objectives of the current study were threefold. First, I summarized demographic data collected in the current and previous studies (Barker and Williamson, 1988
; Brewer, 1999a
) in the form of stage-structured Lefkovitch matrices. Second, I examined the effects of fire-stimulated increases in seedling establishment on population growth. Third, I performed elasticity and stochastic sensitivity analyses of vital rates of stage-structured populations subjected to different hypothetical fire regimes. In particular, I examined the impact of changes in adult persistence, juvenile persistence, and juvenile growth on population growth and viability.
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MATERIALS AND METHODS |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONLITERATURE CITED |
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). The stages of its life cycle I recognized as important in the current study are presented in a life-cycle diagram (Fig. 1). These stages were based on size, specifically the diameter of the largest pitcher (see Brewer, 1999a
; Fig. 1). I used biological differences in choosing these size classes. Seedlings were considered recent germinants or small 1-yr-old juveniles. Seedlings were entire genets whose largest pitcher was <4 mm in diameter at the keel of the lip. Juveniles were entire genets whose largest pitcher was 
4 mm but <10 mm in diameter. A fundamental difference between seedlings and juveniles is that exclusion of prey from pitchers of juveniles potentially reduces growth rate, whereas prey exclusion has no significant effect on seedlings (J. S. Brewer, unpublished data). Small adults were ramets whose largest pitcher was 10 mm but <20 mm in diameter (or <30 cm in length). This minimum threshold corresponded to the smallest ramets I had observed flowering. Large adults were all ramets whose largest pitcher was 20 mm in diameter or 30 cm in length. Large adult ramets also were observed to produce daughter rhizomes, whereas small adult ramets typically did not.
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). Because rates of vegetative reproduction were low, I had few demographic data on vegetative recruits. For this reason they were lumped together with small adult ramets. As more data become available vegetative recruits could be treated as distinct from small adults. All vital rates are shown in Table 1.
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; J. S. Brewer, Laws, and Mozingo, unpublished data), fires apparently do not increase rates of germination (Ellison, 2001
; J. S. Brewer, Laws, and Mozingo, unpublished data).
Sources of data for matrices
I generated stage-structured matrices (Lefkovitch, 1965
) from averages of vital rates based on 2, or in the case of survivorship, 3 yr of demographic data from a population that was burned, on average, once every 3 yr. This fire-return interval is typical of populations at the University of Mississippi Forest Lands (hereafter, UMFL), most of which have received regular prescribed burning since the early 1980s. I used data collected in the current study and previous studies (Barker and Williamson, 1988
; Brewer, 1999a
). A previous experiment showed that individual winter fires had no significant effect on means or variances of growth, survivorship, or sexual or vegetative reproduction of transplants beyond the seedling stage (Brewer, 1999a
). I used demographic data obtained from Brewer (1999a)
to calculate vital rates associated with vegetative reproduction and survivorship of seedlings, juveniles, and adults in two consecutive years. I used data obtained from Barker and Williamson's (1988)
study of the same species in a similar habitat to calculate vital rates associated with fecundity.
In the current study, I collected a third year of data on growth and survivorship of seedlings and post-seedling juveniles and adults. I marked and monitored 80 plants each of seedlings, juveniles, small adults, and large adults from May 1998 to May 1999 (see Brewer, 1999a
, for details on study sites and censusing procedure).
The transition from adult ramet to seedling (i.e., fecundity) was obtained from data presented in Barker and Williamson (1988)
. I used their data because they collected fecundity data from replicate burned and unburned plots (8 x 8 m), whereas I only have data from unburned sites. Barker and Williamson's data on fecundity from their unburned control plots were similar to what I estimated at my field sites. I calculated the transitions from established ramets ("buds" cf. Barker and Williamson, 1988
) to seedlings in burned and unburned plots. These averages of fecundity were then used to calculate stage-specific fecundities for small and large adult ramets.
Fecundity was assumed to vary from year to year according to the proportion of ramets that flowered in the year preceding seedling emergence (see Barker and Williamson, 1988
). Seedling densities were highly correlated with the incidence of flowering in the previous year (Barker and Williamson, 1988
; J. S. Brewer, unpublished data). Year-to-year variation in flowering percentages of adults was obtained from Barker and Williamson (1988)
. These authors found the incidence of flowering varied considerably among years (32.5%). This magnitude of variation is comparable to what I have found at UMFL.
Deterministic model analysis
For each relevant stage (small adults and large adults), I calculated fecundities for years with and without fire. Assuming deterministic geometric growth, I estimated the effect of the frequency of fire-stimulated seedling establishment on population growth by calculating the finite rate of increase, 
. Matrices contained geometric means of fecundity, which varied according to the proportion of years with fire (Table 2). When evaluating the effects of fire frequency on population dynamics, the means of all vital rates other than fecundity were held constant. Using RAMAS EcoLab, was calculated by finding the maximum nonnegative eigenvalue (Caswell, 1989
) for each of five matrices: Annual fires, No fire, and one for each of three matrices with average fire-return intervals of 3, 7, and 20 yr. The 3-yr fire-return interval is typical of the population studied and of most populations at UMFL. In addition to , 95% confidence intervals were calculated for each population. These were calculated using standard deviations derived from 500 bootstrapped samples. Expected s derived from bootstrapping did not differ appreciably from those generated by RAMAS. Therefore, only s produced by RAMAS were presented.
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), I calculated elasticities (proportional sensitivities) for each vital rate (see de Kroon et al., 1986
). In addition to elasticities, I performed a loop analysis for each of the matrices (van Groenendael et al., 1994
) . This was done to examine the impact of changes in unique life-cycle pathways (e.g., early vs. late reproduction) on growth rate.
Stochastic model analysis
To examine the effect of fire-stimulated seedling establishment on population increases and decline, I incorporated environmental stochasticity in vital rates into the projections. Year-to-year stochasticity was produced by randomly sampling from a normally distributed range of vital rates using standard deviations derived from 2–3 yr of empirical data (see Akçakaya, Burgman, and Ginzburg, 1999
).
Five hundred replications of 20-yr projections were run for each stochastic simulation. I quantified the decline of each population by calculating the proportion of replicate populations that declined to <50% of the original population size in 20 yr. Because I did not incorporate demographic stochasticity in the simulations (i.e., random variation among individuals in the population), results provided underestimates of decline probabilities. Nevertheless, previous work with perennial plants has shown that environmental stochasticity is often much more important than demographic stochasticity, especially in large populations (Menges, 1992
). In addition to the probability of decline, I examined the probability that a population doubled in size in 20 yr. Initial population sizes were set at 450 individuals (50 seedlings, 125 juveniles, 50 small adults, and 225 large adults, approximating a stable distribution expected for a population experiencing a 3-yr fire-return interval). I assumed no density dependence (see Menges, 1992
). I did not permit survival rates (i.e., Pi + Gi) to increase above 1 or fall below 0 in any year. However, P4 + V (large adult survival + vegetative reproduction) was allowed to exceed 1, and P4 was truncated at 1. All bootstrapping and simulations were performed on a spreadsheet (Microsoft Excel 98 for Macintosh).
In the demographic analyses, I assumed that fire frequency had no effect on juvenile or adult survivorship or growth (see Brewer, 1999a
). I recognize that long-term fire suppression that enables increased colonization and recruitment of woody species could have a significant negative effect on juvenile and adult survivorship of long-lived perennial herbs (Wells, 1928
; Wells and Skunk, 1928
; Brewer, 1998a
; Maliakal, Menges, and Denslow, 2000
). However, the empirical data in support of woody encroachment of bogs following fire suppression are equivocal (Streng and Harcombe, 1982). There is no convincing experimental evidence that a single fire has an immediate positive or negative effect on the growth or survivorship of juveniles or adults of this species in habitats burned at the frequency currently prescribed at UMFL (see Brewer, 1999a
). At present, there are no published data quantifying the effect of fire frequency on annual rates of adult survivorship in this species. However, as these data become available, they could be incorporated easily into the models.
To examine the sensitivity of population dynamics to proportional changes in the different vital rates, I simulated 38% reductions in all vital rates. This percentage equals the geometric mean reduction in fecundity expected from reducing fire frequency from once every 3 yr to complete fire suppression. In addition, I evaluated the sensitivity of populations to 7% reductions in adult persistence, juvenile persistence, and juvenile growth. Choosing this percentage was an a posteriori decision. Reductions of more than 7% in adult persistence resulted in declining populations. Thus, I used 7% reductions to evaluate the relative importance of adult persistence, juvenile persistence, and juvenile growth on population growth and viability. Statistically significant differences among simulations in the probability of decline or increase were determined for selected comparisons using a chi-square test of independence. Significance values of multiple comparisons were adjusted using Sîdak's multiplicative inequality (Sokal and Rohlf, 1981
).
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RESULTS |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONLITERATURE CITED |
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= 1.044), assuming fixed vital rates and a fire-return interval of 3 yr (Fig. 2). Increasing fecundity by increasing fire frequency to annual burning increased to 1.104 (Fig. 2). Reducing geometric mean fecundity by 38% to levels associated with fire suppression still produced a growing population ( = 1.028). Populations were projected to grow at rates of 3.4 and 3% per year at fire-return intervals of 7 and 20 yr, respectively (Fig. 2). Confidence intervals associated with all fire frequencies except annual burning contained a stationary population ( = 1; Fig. 2). Thus, the finite growth rates of these populations were not significantly different from unity.
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) and much less sensitive to changes in fecundity or growth by small adults or vegetative reproduction by large adults (Table 3). Furthermore, population growth was more sensitive to sexual reproduction by large adults than sexual reproduction by small adults, suggesting that delayed reproduction may be favored in this species (Table 3).
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Stable proportions and reproductive values of stages varied according to fire frequency. As expected, stable stage distributions contained a larger proportion of adults and fewer seedlings as fire frequency decreased (Fig. 4). Distributions in the field were not stable, in large part, because of recurring fires. Seedling densities are higher in years following fire than in other years (Barker and Williamson, 1988
; J. S. Brewer, Laws, and Mozingo, unpublished data). Among post-seedling plants, however, I found that 16% were juveniles, 25% were small adults, and 59% were large adults. This distribution differed somewhat from the stable stage distribution, which contained a greater fraction of juveniles than small adults (see 3-yr fire-return interval in Fig. 4). However, the observed preponderance of large adults is consistent with what would be expected in a population with a stable stage distribution.
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2 = 115.92, P < 0.0001, df = 1). An increase in mean fire-return interval from 7 to 20 did not produce a significant decline in the probability of doubling (from 23.6 to 21.4%, Table 4, 2 = 0.69, adjusted P = 0.97, df = 1). The probability of decline below 50% in 20 yr was <1 in 500, regardless of fire frequency (Table 4).
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Reducing fecundity (by 38%) by removing fire-stimulated increases in seedling establishment had a greater impact on population increase than did reducing juvenile growth by 38% (Table 4; compare no burn to 38% reduction in G2; 20 vs. 36% doubling probability, respectively; 2 = 33.07, P < 0.0001, df = 1). Conversely, a 38% drop in juvenile persistence had a much greater effect on population increase than did a similar reduction in fecundity (Table 4; 8.6 vs. 20%, respectively; 2 = 27.29, P < 0.0001, df = 1).
Juvenile persistence had a much greater effect on population growth than did juvenile growth (Table 4). Reducing the juvenile persistence by 7% significantly reduced the probability of doubling in 20 yr from 57 to 39.6% (Table 4; 2 = 28.94, P < 0.0001, df = 1). In contrast, reducing juvenile growth by 7% did not significantly reduce the probability of doubling in 20 yr (57 vs. 53.2%, Table 4; 2 = 1.46, adjusted P = 0.83, df = 1). The probability of doubling was affected more by reductions in juvenile persistence than by reductions in juvenile growth for large (38%) reductions as well (Table 4).
A 38% reduction in the persistence of small adults produced similar results to an equivalent reduction in juvenile persistence (Table 4). Population growth was not particularly sensitive to 38% reductions in either seedling persistence or vegetative reproduction (Table 4).
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DISCUSSION |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONLITERATURE CITED |
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), I found that the growth and survival of established juveniles and adults are relatively unresponsive to individual fires. Sarracenia exhibited considerable morphological plasticity in response to fire. Individual ramets reduced allocation to pitchers while storing biomass in rhizomes and roots during years between fires, when aboveground biomass associated with neighboring vegetation was greater. This species increased allocation to pitchers, however, immediately following a fire. Bud dormancy and morphological plasticity in allocation to carnivorous structures may enable established ramets of pitcher plants to respond effectively to changes in light availability associated with modest changes in fire regimes (Givnish et al., 1984
; Zamora, Gómez, and Hódar, 1998
; Brewer, 1999a
). These different studies of Sarracenia (Barker and Williamson, 1988
; Brewer, 1999a
) illustrate that examining short-term responses of only one stage of the life cycle of plants to fire provides an incomplete or even misleading picture of the role of fire in maintaining populations of fire-tolerant plants (Bergelson, 1990
; Brewer, 1999b
). Demographic analyses, such as this one, and long-term experiments are necessary to understand the responses of long-lived plants to fire.
The modest effect of fire-stimulated seedling establishment on the maintenance of Sarracenia populations is entirely consistent with the life history of this species. Demographic patterns in Sarracenia indicate a "stress-tolerator" strategy (Grime, 1977
). A comparative demographic analysis of 66 plant species from a wide variety of habitats revealed that population growth of stress-tolerators was more sensitive to changes in survival and persistence within stages than to changes in growth or fecundity (Silvertown et al., 1993
). Given the greater sensitivity of population dynamics of Sarracenia to changes in ramet survival than any other vital rate, more work is needed to investigate potential causes of adult and post-seedling juvenile mortality in this species.
The responses of Sarracenia to fire contrast with those of a short-lived carnivorous species, Drosera capillaris (Brewer, 1999b
). Drosera capillaris shows much higher rates of juvenile and adult mortality than does Sarracenia, owing in part to the former's smaller size and greater vulnerability to burial by shifting sediment (Brewer, 1999b
). In addition, not only do fires increase seedling establishment in Drosera, they also create conditions that increase rates of growth to maturity (Brewer, 1999b
). Thus, the impact of fire on carnivorous plants depends on the life history of the species in question. Nevertheless, despite being relatively short-lived, Drosera capillaris may persist within the habitat during extended periods without fire by lying dormant in a persistent seed bank in the soil (Brewer, 1999b
; J. S. Brewer, unpublished data).
The effects of fire suppression on pitcher-plant populations remain an open question. Open parts of bogs recover quickly from individual fires (in terms of aboveground biomass), and there does not appear to be a significant buildup of aboveground biomass in these areas 
2 yr after a fire until trees become established (Brewer, 1999a
). I hypothesize that pitcher-plant populations are negatively affected by increases in fire-return intervals only insofar as such increases permit increased establishment of trees. The extent to which fire suppression permits tree establishment in pitcher-plant bogs is not known. Streng and Harcombe (1982) argued that edaphic conditions, alone, were sufficient to restrict the establishment of trees in a wet grass–sedge bog in east Texas. I have found that fire suppression of up to 15 yr does not necessarily permit increased establishment of woody species in pitcher plant bogs in Mississippi (J. S. Brewer, unpublished data). Bog sites differ in the rate at which trees become established, due to differences in edaphic conditions (J. S. Brewer, unpublished data; see also Glitzenstein et al., 1995). Pine seedling and sapling mortality are higher in more poorly drained bogs (even in the absence of fire; J. S. Brewer, unpublished data). The rate at which trees become established, in turn, influences the rate at which shrubs and vines become established. Trees provide safe perches for frugivorous birds carrying seeds of long-lived species of shrubs and vines (e.g., Ilex glabra, I. vomitoria, I. coriacea, Gaylussacia mosieri, Smilax laurifolia; Brewer, 1998a
; Ashley and Brewer, 1999
). Larger perches become the sites of dense, tree-centered thickets (Brewer, 1998a
), which may persist even after the tree dies (J. S. Brewer, personal observations). Species richness of carnivorous plants and the frequency of pitcher plants are significantly lower within these thickets (Brewer, 1998a
). I suggest that this is partly because of the negative effects of woody plants on the survival of established carnivorous plants (Brewer, 1998b
).
The optimal prescribed-fire frequency necessary to maintain carnivorous plant diversity may depend on the productivity of the bog, which in turn is likely to be a function of the hydrologic regime. I suggest that frequent fires (i.e., those occurring more than once a decade) may not be required to reduce tree establishment and thus maintain carnivorous plant communities in the most poorly drained bogs. Nevertheless, attempts at complete fire suppression in these bogs would be risky, unwarranted, and in the long term, impossible (see Streng and Harcombe, 1982). Once woody species get a foothold in a bog, recovery of carnivorous plant communities may require active intervention, involving perhaps mechanical removal of woody species or, at the very least, repeated growing-season fires (Olson and Platt, 1995
).
Is fire-stimulated seedling establishment relevant to the conservation of rare pitcher plants?
If fire-stimulated seedling establishment has relatively small effects on the maintenance of populations of common species of pitcher plants, as this study indicates, then what role should prescribed burning play in the conservation of rare pitcher plants? To answer this question, one must go beyond the simple demographic analysis presented here and investigate metapopulation dynamics of rare species (see Menges, 1990
; Schemske et al., 1994
; Thomas, 1994
; Hanski and Simberloff, 1997
). Most pitcher plants and other carnivorous plants have fairly restrictive habitat requirements (e.g., wet, sunny, nutrient-poor bogs and savannas; Givnish et al., 1984
). Thus, suitable habitat patches may be scattered across the landscape and isolated, with dispersal limitation being potentially severe. Because of low rates of seedling establishment and the prohibitive expense and difficulty of transplanting large numbers of adult ramets, reintroductions will produce very small founder populations that are vulnerable to local extinction. Consequently, one must understand which variables increase rates of seedling establishment following introduction to or colonization of suitable habitat patches.
If fire increases the establishment of seedlings in rare species, as it does in Sarracenia alata, then small reductions in fire-return intervals (from 7 to 3 yr) could significantly promote colonization of suitable habitat patches by increasing population growth rates and buffering small populations from extinction (Huston, 1994
). Furthermore, some rare species of Sarracenia appear to be predominantly outcrossing plants (Godt and Hamrick, 1996
). Genetic diversity in pitcher plant populations declines sharply in response to bottlenecks and founder effects in some species (Godt and Hamrick, 1996
). By adding new genets, fire-stimulated increases in seedling establishment could increase genetic diversity in recovering populations of pitcher plants. Combining the transplantation of a small but genetically diverse sample of adult ramets with frequent prescribed winter burning may provide a relatively cost-effective means of establishing a genetically diverse population of pitcher plants. If the absence of pitcher plants from bogs is the result of habitat deterioration or collection by enthusiasts rather than dispersal limitation, however, then habitat restoration and protection must precede reintroduction efforts (Thomas, 1994
).
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FOOTNOTES |
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2 Author for reprint requests (Fax: 662-915-5144, jbrewer{at}olemiss.edu
).
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LITERATURE CITED |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONLITERATURE CITED |
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pd < 10 mm). Stage 3 is small adults (10 
