Reversal of human-induced vegetation changes in Sequoia National Park, California

Reversal of human-induced vegetation changesin Sequoia National Park, California

D. Graham Roy and John L. Vankat

Abstract: We resampled 76 permanent plots that had been established in the woodlands and forests of SequoiaNational Park in 1969. Our objectives were to describe vegetation changes in the tree and shrub layers and determinethe effects of prescribed burning that began in the 1960s. We compared changes in species importance and tree sizeclass distributions between sample dates and between burned and unburned plots. Species composition had remainedsimilar in all nine vegetation types sampled except in the ponderosa pine forest wherePinus ponderosaDougl. ex P. &C. Laws. decreased in importance from 28 to 15% andAbies concolor(Gord. & Glend.) Lindl. increased from 18 to31%. Structural changes were more common, as tree density decreased in the blue oak woodland (19%), and live oakwoodlands (15%), as well as in ponderosa pine forest (41%), white fir forest (5%), giant sequoia groves (39%), andred fir forest (24%). Decreases in density were greater in burned plots but occurred in unburned plots as well,indicating that prescribed fire and self-thinning contributed to decreases in density. Tree density was unchanged in thelodgepole and subalpine forests, but increased in the Jeffrey pine forest (58%). The decreases in tree density representa reversal of earlier trends.

Résumé: Les auteurs ont échantillonné de nouveau 76 places échantillons permanentes établies en 1969 dans lesforêts claires et les forêts fermées du parc national Séquoia. Leur objectif était de décrire les changements dans lesstrates arborescente et arbustive de la végétation et déterminer les effets du brûlage dirigé, qui a commencé dans lesannées 1960. Ils ont comparé les changements concernant l’importance des espèces et la distribution des classes dedimension des arbres entre les différentes dates d’échantillonnage et entre les places échantillons brûlées et nonbrûlées. La composition spécifique est restée semblable dans les neuf types de végétation échantillonnés, excepté dansla forêt de pin ponderosa où l’importance duPinus ponderosaDougl. ex P. & C. Laws. a diminué de 28 à 15% etcelle d’Abies concolor(Gord. & Glend.) Lindl. s’est accrue de 18 à 31%. Les changements dans la structure étaientplus fréquents à mesure que la densité diminuait dans la forêt claire de chêne bleu (19%) et celle de chêne del’intérieur (15%), ainsi que dans la forêt de pin ponderosa (41%), la forêt de sapin concolore (5%), la futaie deséquoiadendron géant (39%) et la forêt de sapin rouge (24%). La diminution de la densité a été plus grande dans lesplaces échantillons brûlées, mais elle était également présente dans les places échantillons non brûlées, indiquant par làque le brûlage dirigé et l’éclaircie naturelle avaient contribué à la diminution de la densité. La densité des arbres étaitinchangée dans la forêt de pin tordu latifolié et dans la forêt subalpine, par contre elle s’est accrue dans la forêt de pinde Jeffrey (58%). La diminution de la densité des arbres représente le renversement d’une tendance antérieure.

[Traduit par la rédaction] Roy and Vankat 412

Vegetation changes must be interpreted within the contextof natural and human influences. In the conifer forests ofwestern North America, naturally occurring fires have beenimportant disturbances influencing their composition andstructure. Humans have altered natural forest structure byexcluding fire and introducing new disturbances, which leadto species replacement and modified vegetation structure.

Knowledge of vegetation change is particularly importantin national parks where land managers are responsible formaintaining natural ecological processes and vegetation.

Fire suppression and livestock grazing in the national parksof the western United States and Canada represent humaninfluences that have altered the structure and composition ofnatural vegetation. In dry forests of western North America,frequent, low-intensity fires maintained an open forest struc-ture because fires killed saplings and recycled nutrients toherbaceous species that compete with tree seedlings. Firesuppression has allowed tree seedlings to survive, producingdense stands of subcanopy-height trees that decrease the ho-mogeneity of forest patches (Tande 1979; Barrett et al.1991; Swetnam 1993).

Another change in the natural disturbance regime of manywestern national parks occurred when livestock grazing be-gan between the mid-1800s and early 1900s. Grazing alsoled to increases in density of small trees as livestock re-moved grasses and herbaceous plants, releasing tree seed-lings from competition (Belsky and Blumenthal 1997).When fuel levels decreased on the forest floor, fires de-creased in frequency as well, further contributing to in-creases in tree density. When grazing ended after NationalParks were established in western North America, tree

Can. J. For. Res.29: 399–412 (1999) © 1999 NRC Canada

399

Received March 31, 1998. Accepted January 4, 1999.

D.G. Roy1 and J.L. Vankat. Department of Botany, MiamiUniversity, Oxford, OH 45056, U.S.A.

1Author to whom all correspondence should be addressed.Present address: Department of Biology, P.O. Box 3001,Department 3AF, New Mexico State University, Las Cruces,NM 88003, U.S.A. e-mail: [email protected]

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recruitment either returned to pregrazing levels (Madany andWest 1983), or remained high because seedlings had littlecompetition from grasses and forbs that recovered slowly(Vankat and Major 1978; Vale 1981, 1987).

The effects of past fire suppression and grazing in Se-quoia National Park, in the southern Sierra Nevada of Cali-fornia, are similar to other parks of western North America.Estimates of fire frequencies prior to Euro-American influ-ence suggest that fires burned every 8–20 years in themidelevation conifer forests (Kilgore and Taylor 1979;Swetnam 1993), and an average of two lightning fires oc-curred annually in the shrublands and woodlands of thepark’s foothills (Parsons 1981). At least in the woodlandsand forests, frequent fires were generally low in intensity,maintaining an open structure by burning shrubs and sap-lings (Vankat 1970; Rundel et al. 1977; Vankat and Major1978; Swetnam 1993). Beginning in the 1860s, however, thefire regime abruptly changed when human ignitions de-creased with the decline of Native American populationsand cattle and sheep grazing reduced fuel loads (Vankat1977; Parsons 1981; Swetnam 1993). After Congress estab-lished Sequoia National Park in 1890, the U.S. Governmentalso initiated aggressive control of wildfires with the intentof protecting natural resources (Vankat 1977).

Several studies have examined the effects of fire suppres-sion on vegetation in Sequoia National Park (Vankat 1970;Kilgore and Briggs 1972; Parsons and DeBenedetti 1979;Vankat and Major 1978; Bonnicksen and Stone 1982; Bar-bour 1988). Fire suppression in conifer forests has led to in-creased densities of shade-tolerant species (Barbour 1988)and has also reduced the abundance of shrub species that re-quire fire for reproduction (Vankat and Major 1978; Barbour1988).

Past livestock grazing has also resulted in increased densi-ties of trees in some forests and woodlands in Sequoia Na-tional Park (Vankat 1970; Vankat and Major 1978). Duringthe mid-1800s up to 100 000 sheep grazed in the Park dur-ing the summer, but widespread grazing ended after Con-gress established the Park. In oak (Quercusspp.) woodlandsand lodgepole pine (Pinus contortaDougl. ex Loud.) forests,a flush of tree reproduction followed the end to grazing, ap-parently because domestic sheep left exposed mineral soil,ideal conditions for seed germination, and no longer tram-pled and ate seedlings (Vankat and Major 1978). The end tograzing has only led to an increased density of dominanttrees rather than a shift in species composition, which oc-curred with fire suppression.

With evidence that fire suppression led to increased treedensity, the National Park Service began prescribed burningin Sequoia National Park during the 1960s to restore foreststo the densities that existed before fire suppression began.Prescribed burning represented a major shift in policy de-signed to influence the direction of vegetation change in thepark (Pyne 1982, 1984). The goal of the program was to re-duce the density of understory trees and fuel loads, restoringthe forests to conditions that existed when the park was es-tablished (Bancroft et al. 1985; Butts 1985). Several studieshave examined the direct effects of prescribed fire on treemortality and species composition in areas that have burned(Kilgore 1973a, 1973b; Keifer and Stanzler 1995). In un-burned areas, however, it is unknown whether previous in-

creases in density are continuing as in the northern SierraNevada (Ansley and Battles 1998) or have reversed, as pre-dicted for the conifer forests of the San Bernadino Moun-tains in southern California (Minnich et al. 1995).

In 1969, the second author established 110 permanentplots in Sequoia National Park to quantitatively describe thewoodlands and forests and to qualitatively describe vegeta-tion changes resulting from fire suppression and grazing(Vankat 1970; Vankat and Major 1978). We resampled 76 ofthe permanent plots to determine how the vegetation hadchanged since then. Although 27 years is a short time periodrelative to the life-spans of woodland and forest trees, it en-compasses the period of prescribed burning in the Park, andit is sufficient to answer two specific questions. (i) Havepreviously documented changes in vegetation continued?(ii ) What have been the effects of prescribed burning onwoodland and forest composition and structure?

Physical settingSequoia National Park is bordered by Kings Canyon National

Park to the north and national forests on all other sides. It islocated between 36°1′ and 36°42′N and between 118°14′ and118°55′W. The park has great topographic variation and ranges inelevation from 450 to 4417 m. The Kaweah and Kern rivers are thetwo major river drainages and are situated on the west and eastsides of the Great Western Divide, respectively. The climate ischaracterized as Mediterranean, with most precipitation fallingfrom October through May and drought conditions in the summer.Precipitation varies with elevation, and the Kern drainage receivesless precipitation than the Kaweah because of the rain shadow ofthe Great Western Divide (Stephenson 1988).

VegetationDifferences in slope, aspect, and climate within the park pro-

duce many vegetation types (Vankat and Major 1978; Vankat 1982;Stephenson 1988) along topographic and moisture gradients(Vankat 1982). The following descriptions include only the vegeta-tion types examined in this study and are based on Vankat andMajor (1978) and Vankat (1982), although differences withStephenson (1988) are minor.

Low-elevation woodlandsThe blue oak woodland is at 390–700 m and is dominated by

Quercus douglasiiHook. & Arn. (blue oak),Aesculus californica(Spach) Nutt. (California buckeye), andQuercus wislizeniiA. DC.(California live oak). Tree cover is generally at the lower end of30–85%. Herbaceous cover is up to 80%.

The lowland live oak woodland occurs in riparian areas at390–1000 m. Tree densities are higher than in the blue oak sub-type, and canopy cover is generally at the high end of 30–80%.Quercus wislizeniiis the dominant tree, butAesculus californicaand Q. douglasiiare also important.

Midelevation forestsThe ponderosa pine forest occurs only in the Kaweah River

drainage at elevations of 1300–2200 m. Xeric sites at the lowerend of the elevation range have an open canopy of scattered, ma-ture Pinus ponderosaDougl. ex P. & C. Laws. and a denseunderstory ofQuercus kelloggiiNewb. (black oak) andCalocedrusdecurrens(Torr.) Florin (incense cedar). Mesic sites have a moreclosed canopy and are codominated byCalocedrus decurrens,Pinus ponderosa, and Q. kelloggii. These sites havePinus

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400 Can. J. For. Res. Vol. 29, 1999



lambertiana Dougl. (sugar pine) as a subdominant andAbiesconcolor (Gord. & Glend.) Lindl. (white fir) as a frequent under-story tree. Both xeric and mesic areas haveChamaebatia foliosaBenth. as the dominant shrub, andArctostaphylos viscidaC. Parryis also common.

The Jeffrey pine forest occurs in both the Kaweah and Kerndrainages and at higher elevations than the ponderosa pine forest(2100–2800 m). The canopy is generally open and dominated byPinus jeffreyi Grev. & Balf. (Jeffrey pine), with Calocedrusdecurrens and Abies concoloras subdominants.Arctostaphylospatula E. Greene andChamaebatia foliosaare the most commonshrubs.

With more mesic conditions, the ponderosa pine forest gradesinto a white fir forest with an elevation range of 1600–2300 m.Tree cover is over 75% and is dominated byAbies concolor. Com-mon subdominants areCalocedrus decurrensand Pinus lamber-tiana; Q. kelloggii may also occur in more xeric stands. Shrubcover is low, but where presentRosa spithameaS. Watson is themost common species. Groves ofSequoiadendron giganteum(Lindl.) Buchholz (giant sequoia) contain scattered individuals ofthe species in forests that are otherwise similar in composition tothe white fir forest. Nevertheless, the groves are usually treated asa separate vegetation type because of their importance in the land-scape (Vankat and Major 1978; Vankat 1982; Stephenson 1988)and to provide information for managing the groves (Harvey 1980;Stephenson 1996).

High-elevation forestsThe red fir forest ranges in elevation from 2300 to 2800 m and

contains mostly pure stands ofAbies magnificaA. Murray (red fir)with approximately 80% canopy cover.Abies concolormay becodominant at lower elevations, andPinus contortais present athigher elevations near the transition with the lodgepole pine forest.Shrubs are generally absent.

The lodgepole pine forest ranges from 2500 to 3100 m and con-tains relatively homogenous stands ofPinus contorta. Abies mag-nifica is a subdominant in some low elevation stands.Ribesmontegeumis the most common shrub.

The subalpine forest occurs at 3000–3500 m and contains eitherPinus balfourianaGrev. & Balf. (foxtail pine), Pinus albicaulisEngelm. (whitebark pine), orPinus monticolaDougl. (westernwhite pine) in an open canopy with about 25% cover. Some standsof Pinus contortamay also be considered subalpine. Herb andshrub cover is low.

Vankat (1970) placed 110 permanent plots (50 × 2 m belttransects marked with steel stakes at both ends) in areas of SequoiaNational Park accessible by road or trail. Except for six placedacross forest–meadow ecotones, all plots were located in relativelyhomogeneous areas judged to be representative of the major vege-tation types of the park. We used original site photographs, maps,and written directions to relocate the stakes for 76 of the 110 plotsfrom June to August of 1996. We resampled the plots following theprocedures of Vankat (1970), measuring the diameter at breastheight (DBH) of trees≥2.5 cm, mapping their position in thetransect, and projecting the canopies of woody plants onto thetransect to calculate cover using the line-intercept method(Andresen and McCormick 1962). Nomenclature follows Hickman(1993). Tree boles that overlapped the edges of plots were counted,resulting in a slight overestimate of basal area, especially forSe-quoiadendron giganteum.

To determine which plots had been prescribed burned, we com-pared the locations of plots with a Park Service database contain-ing the location, area, and dates of fires. We confirmed which plotshad burned using evidence of fire in the plots such as charred bark

or wood. The first fires occurred in 1969 and the last in 1994, butwe could not date all of the fires because some areas that obvi-ously burned were not recorded in the park’s fire database.

In our analysis, we first combined all plots to examine generalchanges in species composition between sample dates. Relativedensity for each species in a vegetation type was calculated as theratio of the number of stems of a species to total stems for all spe-cies in the vegetation type. Relative values for cover and basal areawere calculated with the same method, and the average of relativedensity, basal area, and cover was calculated as the importancepercentage (IP) for each tree species in the vegetation types.

We placed trees from both sample dates into 10-cm size classes(2.5–9, 10–19, 20–29 cm, etc.) and tested for changes in the sizeclass distributions of tree species with >20 individuals during bothsample dates. We used a chi-square test for large samples, butwhen greater than 25% of the cells in the contingency table con-tained expected values <5, we used a Fisher’s exact test (SASInstitute Inc. 1990; Sokal and Rohlf 1995). When significant dif-ferences were found with either test, the contingency table was col-lapsed to a 2 × 2table to compare the number of stems in each sizeclass to the combined total of all others. We performed aBonferroni’s adjustment, dividingα = 0.05 by the number of 2 × 2table comparisons to calculate a critical value for the tests. To in-sure more accurate tests, the actual number of trees in each vegeta-tion type was used, but to facilitate further comparisons the dataare presented as the number of trees per hectare. (See Roy (1997)for the number of trees in each size class.)

We compared individual plots between sample dates usingdetrended correspondence analysis (DCA) from PC-ORD version2.0 (McCune and Mefford 1995) with relative tree and shrub coverdata for individual plots. Large differences in species compositionnecessitated performing ordinations on individual vegetation typesseparately, and ordinations of the live oak woodland, lodgepolepine forest, and subalpine forest had too few stands or species tobe meaningful. Variation that was accounted for by individual axesof each ordination was expressed as a percentage of the sum of theeigenvalues of the first three DCA axes. Plots from both sampledates were analyzed on the same ordination, so in several instancesplot position did not change when the relative cover values amongspecies remained the same. When the same plot had different ordi-nation positions for 1969 and 1996, however, we examinedcompositional data for the plot to determine what accounted forshifts in position.

We separated unburned and burned plots within each vegetationtype to evaluate the effects of prescribed fire. We used the DCA or-dinations to detect shifts in species composition in individualburned and unburned plots. We also compared unburned andburned plots for changes in tree basal area, density, mortality, andingrowth (trees that were≥2.5 cm in 1996 but not present in 1969)between the sampling dates when at least three plots within a vege-tation type had burned. The Jeffrey pine, red fir, and lodgepolepine forests each had only one burned plot.

Low-elevation woodlands

Blue oak woodlandWe relocated nine plots in the blue oak woodland, and

they indicated that total tree density decreased 19% (111trees/ha; Table 1). Importance percentages indicated thatQ. douglasii remained the dominant tree, but its absolutedensity decreased 12% (78 trees/ha).Arctostaphylos viscidamaintained only 1% shrub cover in both 1969 and 1996(Table 2). The position of most plots did not change on theDCA ordinations (Fig. 1A), at least in part because they

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402 Can. J. For. Res. Vol. 29, 1999

contained onlyQ. douglasii.The size class distribution ofQ. douglasiialso changed from 1969 to 1996 (Table 3) witha statistically significant decrease in the proportion of trees

in the 10–19 cm DBH size class (Fig. 2A), where mortalitywas 15%.

Comparisons of three unburned and six burned plots

Density Basal area Cover IP

Vegetation type Species 1969 1996 1969 1996 1969 1996 1969 1996

Blue oak woodland Quercus douglasii 678 600 21 28 55 54 96 98(900 m2, 9 plots) Aesculus californica 33 0 1 0 2 3 3 1

Quercus wislizenii 0 0 0 0 1 1 1 1Total 711 600 22 28 59 58 100 100

Live oak woodland Quercus wislizenii 400 250 17 21 73 72 55 50(200 m2, 2 plots) Quercus douglasii 100 100 20 22 18 14 26 25

Aesculus californica 150 150 3 7 32 25 19 21Rhamnussp. 0 50 0 1 0 0 0 4Total 650 550 40 51 123 111 100 100

Ponderosa pine forest Abies concolor 466 444 14 25 24 30 18 31(900 m2, 9 plots) Calocedrus decurrens 655 244 33 43 36 21 28 26

Quercus kelloggii 422 267 11 6 42 36 20 21Pinus ponderosa 256 66 69 32 29 12 28 15Pinus lambertiana 178 144 6 1 10 5 6 7Total 1977 1165 133 107 141 104 100 100

Jeffrey pine forest Pinus jeffreyi 200 178 63 69 27 18 55 43(450 m2, 5 plots) Calocedrus decurrens 200 267 23 27 8 9 27 26

Abies concolor 89 356 1 4 7 9 11 22Populus balsamifera 22 22 3 4 3 5 5 6Pinus contorta 22 22 1 1 1 1 2 3Total 533 844 91 105 46 42 100 100

White fir forest Abies concolor 600 507 130 124 67 54 74 72(1400 m2, 14 plots) Calocedrus decurrens 64 86 12 6 13 10 9 9

Pinus lambertiana 50 79 14 11 12 5 9 8Cornus nuttallii 29 29 6 0 10 18 5 8Abies magnifica 7 14 1 0 2 3 2 2Quercus kelloggii 7 7 0 0 3 2 1 1Pinus jeffreyi 0 0 0 0 1 0 <0.5 0Quercus chrysolepis 0 0 0 0 <0.5 0 <0.5 0Umbellularia californica 0 0 0 0 <0.5 0 <0.5 0Total 757 722 163 141 108 92 100 100

Giant sequoia groves Abies concolor 791 473 94 65 59 50 46 50(1100 m2, 11 plots) Sequoiadendron giganteum 45 45 819 267 26 20 37 34

Pinus lambertiana 82 73 37 29 16 14 8 12Abies magnifica 55 18 5 6 10 3 5 3Calocedrus decurrens 36 9 0 0 1 0 1 1Cornus nuttallii 0 0 0 0 9 1 3 0Total 1009 618 955 367 121 88 100 100

Red fir forest Abies magnifica 878 689 174 181 66 62 86 90(900 m2, 9 plots) Abies concolor 122 89 16 15 8 6 10 9

Pinus lambertiana 11 11 0 0 1 1 1 1Pinus contorta 22 0 5 0 3 1 3 0Total 1033 789 195 196 78 70 100 100

Lodgepole pine forest Pinus contorta 980 920 153 156 55 48 93 88(500 m2, 5 plots) Abies magnifica 60 120 10 14 4 10 7 12

Pinus lambertiana 0 0 0 0 <.5 0 <.5 0Total 1040 1040 163 170 59 58 100 100

Subalpine forest Pinus balfouriana 275 275 106 113 16 17 68 66(400 m2, 4 plots) Pinus contorta 100 100 51 56 9 12 32 34

Total 375 375 157 169 25 29 100 100

Note: The area and number of plots in each woodland or forest are given in parentheses.

Table 1. Absolute density (trees/ha), basal area (m2/ha), cover (%), and importance percentages (IP) of tree species in the woodlandsand forests of Sequoia National Park in 1969 and 1996.



indicated that changes in structure were similar in both plottypes. Basal area, for example, decreased 36% (10 m2/ha)and 27% (5 m2/ha) and density decreased 15% (85 trees/ha)and 11% (65 trees/ha) in unburned and burned plots, respec-tively (Fig. 3). Although stand structure was unaffected,burning did influence the species composition in threestands as illustrated in the ordination of tree and shrub cover(Fig. 1A). The shifts in burned plots reflect increased coverof Arctostaphylos viscida. The effects of burning should notbe overestimated, however, because four other burned plots

(with Q. douglasiias the only tree species) did not change incomposition.

In the blue oak woodland, we found that there was noingrowth of Q. douglasii, but little evidence that the speciesis affected by fire.Quercus douglasiiregeneration has beenabsent for many decades, with the last major episode of treerecruitment occurring in the late 1800s and possibly corre-sponding to the end of cattle and sheep grazing in the park(Vankat and Major 1978). Fires maintained the oak wood-lands before Euro-American settlement (McClaran and

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Roy and Vankat 403

Absolute cover Relative cover

Vegetation type Species 1969 1996 1969 1996

Blue oak woodland Arctostaphylos viscida 1 1 100 100(900 m2, 9 plots) Rhamnus ilicifolia <0.5 0 0 0

Total 1 1 100 100Live oak woodland Toxicodendron diversilobum 1 <0.5 100 100(200 m2, 2 plots) Total 1 0 100 100Ponderosa pine forest Chamaebatia foliolosa 20 15 76 49(900 m2, 9 plots) Ceanothus integerimus <0.5 7 <0.5 24

Arctostaphylos patula 4 4 15 13Arctostaphylos viscida 2 2 6 7Rosa spithamea 1 2 3 7Toxicodendron diversilobum <0.5 <0.5 <0.5 <0.5Arctostaphylossp. <0.5 <0.5 <0.5 <0.5Total 26 30 100 100

Jeffrey pine forest Chamaebatia foliolosa 12 11 35 53(450 m2, 5 plots) Arctostaphylos patula 21 9 62 43

Artemisia tridentata 1 1 3 5Total 34 21 100 100

White fir forest Corylus cornuta <0.5 10 0 47(1400 m2, 14 plots) Symphoricarpossp. 2 7 100 33

Chrysolepis sempervirens 0 2 0 12Ceanothus integerrimus 0 2 0 8Ceanothus parvifolius 0 <0.5 0 0Rubussp. 0 <0.5 0 0Ribes nevadense 0 <0.5 0 0Ribes roezlii 0 <0.5 0 0Rosa spithmea <0.5 <0.5 0 0Rubus leucodermis 0 <0.5 0 0Sambucus mexicana 0 <0.5 0 0Total 2 21 100 100

Giant sequoia groves Rosa spithamea 0 2 0 72(1100 m2, 11 plots) Ribes roezlii 0 1 0 28

Corylus cornuta 0 <0.5 0 0Ribessp. 0 <0.5 0 0Symphoricarpos albus 0 <0.5 0 0Chrysolepis sempervirens 2 0 100 0Total 2 3 100 100

Red fir forest Symphoricarpossp. <0.5 1 0 100(900 m2, 9 plots) Total <0.5 1 0 100Lodgepole pine forest Ribes montigenum 2 2 68 100(500 m2, 4 plots) Phyllodoce breweri 1 0 32 0

Salix orestera <0.5 0 0 0Total 3 2 100 100

Note: The area and number of plots in each woodland or forest are given in parentheses.

Table 2. Absolute (%) and relative cover of shrub species in the woodlands and forests of Sequoia National Park in1969 and 1996.



Bartolome 1989; Mensing 1992), and are also known tocause sprout development inQ. douglasii (Mensing 1992;Haggerty 1994). However, we recorded no ingrowth in theburned plots, so other factors such as climate, grazing, andsprout competition with annual grasses may also influencerecruitment (Gordon et al. 1989). Without active manage-ment to produce appropriate conditions for regeneration,trends of decreasedQ. douglasiidensity may continue in thenext few decades.

The similarity of change in both burned and unburnedplots indicates that established individuals ofQ. douglasiiare resistant to fire. Haggerty (1994) also documented resis-tance to fire, finding that 93% of theQ. douglasiiindividu-als survived an arson fire above the park headquarters at

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404 Can. J. For. Res. Vol. 29, 1999

*

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Fig. 1. Detrended correspondence analysis (DCA) ordinations ofplots in the vegetation types based on tree and shrub cover in1969 and 1996. Eigenvalues were used to calculate thepercentage of total variation in the first three DCA axes thataxes 1 and 2 describe. B, Burned plots. W, Windthrow. (A) Blueoak woodland. Axis 1 = 96% and axis 2 = 3%. *Four burnedand two unburned plots are superimposed. (B) Ponderosa pineforest. Axis 1 = 65% and axis 2 = 25%. (C) Jeffrey pine forest.Axis 1 = 66% and axis 2 = 26%. (D) White fir forest. Axis 1 =52% and axis 2 = 33%. *One burned and two unburned plotsare superimposed. (E) Giant sequoia groves. Axis 1 = 55% andaxis 2 = 33%. (F) Red fir forest. Axis 1 = 75% and axis 2 =19%. *Three unburned plots are superimposed.

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1969 Samples

1996 Samples

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Size Classes (cm)

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>40

Fig. 2. Size-class distributions of tree species in 1969 and 1996.Percent mortality in each size class is provided above thedistributions. Note difference in scales. (A)Quercus douglasiiinthe blue oak woodland. An asterisk indicates a significantdifference between sample dates (α = 0.01, χ2 834= . , df = 1,p < 0.005). (B)Calocedrus decurrensin the ponderosa pineForest. An asterisk indicates a significant difference betweensample dates (α = 0.02, χ2 87= . , df = 1, p < 0.005).(C) Quercus kelloggiiin the ponderosa pine forest. An asteriskindicates a significant difference between sample dates (2.5- to9-cm size class:α = 0.02, χ2 187= . , df = 1, p < 0.001;10–19 cm size class:α = 0.02, χ2 198= . , df = 1, p < 0.005).(D) Abies concolorin the white fir forest. (E)Abies concolorinthe giant sequoia groves. (F)Abies magnificain the red firforest. (G)Pinus contortain the lodgepole pine forest.Differences in size classes between sample dates were notstatistically significant for Figs 2D–2F (Table 3).



Ash Mountain. She also found that basal area remained es-sentially the same because mortality was restricted to smallsize classes, but we found that mortality was absent in thesmallest size class.

Live oak woodlandWe relocated only two plots in the live oak woodland

where tree density decreased 15% (100 trees/ha; Table 1).Importance percentages ofQ. wislizenii decreased slightlyfrom 55 to 50%, as its density decreased 38% (150 trees/ha),but no ingrowth occurred (Fig. 3). Importance percentagesfor the two subdominant species,Q. douglasiiandAesculuscalifornica, remained similar (Table 1).Toxicodendrondiversilobum(Torr. & A. Gray) E. Greene remained the onlyshrub, with its absolute cover decreasing from 1 to <0.5%(Table 2).

Based on old photographs, Vankat and Major (1978) re-ported thatQ. wislizeniihad increased in density because offire suppression. Our findings show that this trend has re-versed, with decreases in total density. However, we cannotdetermine if densities have decreased to levels before firesuppression began.

Midelevation forests

Ponderosa pine forestWe relocated nine plots in the ponderosa pine forest, and

the forest structure has become more open in these plots. To-tal tree density decreased 41% (812 trees/ha) and corre-sponding decreases occurred in basal area and cover(Table 1). In addition there have been substantial shifts inthe importance percentages of tree species (Table 1). Theimportance percentage ofPinus ponderosadecreased from28 to 15% because of substantial decreases in density (74%;190 trees/ha), basal area (54%; 37 m2/ha), and cover (59%;7% absolute cover).Abies concolorincreased in importancepercentage (18 to 31%) because of increases in basal area(78%; 11 m2/ha) and cover (25%; 6% absolute cover) anddespite a slight decrease in absolute density (5%, 22 trees/ha).Cover in the shrub layer increased 15% (4% absolute cover),and Ceanothus integerrimusHook & Arn. increased from<0.5 to 7% (Table 2). The DCA ordinations reflected thesechanges in composition because all plots changed, generallymoving away from the center of the plot as heterogeneity in-creased from 1969 to 1996. Examination of compositionchanges in each plot showed that decreases inPinus ponder-

osa and increases in the importance percentages of eitherAbies concoloror Calocedrus decurrensaccounted for theshifts in plot position.

Diameter growth and decreases in tree density appear tohave affected the forest structure, as the proportion of treesin different size classes of bothCalocedrus decurrensandQ. kelloggii changed (Table 3).Calocedrus decurrensde-creased in the smallest (2.5–9) cm size class (Fig. 2B) wheremortality was 56%. ForQ. kelloggii ingrowth produced anincrease in the 2.5- to 9-cm size class despite mortality of100%, but the 10- to 19-cm size class decreased with 44%mortality (Fig. 2C).

Four plots remained unburned while five burned, andcomparisons between these plots indicated that changeswere not always consistent. Most species had greater de-creases in density and higher mortality with burning, how-ever (Fig. 4). The density ofAbies concolordecreased 11%(50 trees/ha) and 50% (160 trees/ha) in unburned and burnedplots, respectively, with higher mortality and no ingrowth inburned plots.Calocedrus decurrensdecreased in density31% (125 trees/ha) and 52% (320 trees/ha) in unburned andburned plots, respectively, reflecting higher mortality andlower ingrowth in burned plots. The basal area ofPinus pon-derosadecreased 41% (40 m2/ha) and 74% (35m2/ha), whiledensity decreased 63% (125 trees/ha) and 79% (220trees/ha) in unburned and burned plots, respectively. Highermortality in the burned plots contributed to the proportion-ately larger changes in those plots; no ingrowth was re-corded in either plot type. Changes in burned plotscorrespond to increased cover ofCeanothus integerrimusonthe DCA ordination (Fig. 1B).

Both the lack ofPinus ponderosaestablishment and mor-tality indicate major changes in this forest type. There areseveral explanations for the lack of establishment in the park(L. Mutch, personal communication).Pinus ponderosarequires exposed sites for germination, so recent distur-bances may not have created suitable conditions for estab-lishment. Prescribed fires in particular may not have burnedhot enough to form gaps in the crown, although a crown fireoccurred in one plot without ingrowth. In addition, tree es-tablishment is spatially heterogeneous, so it could be occur-ring in areas we did not sample.

Tree mortality is generally the cumulative effect of sev-eral stresses (Elliot and Swank 1994), and factors implicatedin Pinus ponderosamortality in southern California includedrought in the mid-1970s, attack by bark beetles

© 1999 NRC Canada

Roy and Vankat 405

Vegetation type Species χ2 df PFisher’sexact test

Blue oak woodland Quercus douglasii 0.001Ponderosa pine forest Abies concolor 6.507 5 0.26

Calocedrus decurrens 10.298 3 0.01Quercus kelloggii 0.001

White fir forest Abies concolor 13.717 6 0.033Giant sequoia groves Abies concolor 7.838 4 0.098Red fir forest Abies magnifica 8.532 5 0.13Lodgepole pine forest Pinus contorta 2.7 5 0.74

Note: Species with at least 20 individuals at both sample dates were analyzed, and Fisher’s exact tests were used if >25% of the cells in thecontingency table had values <5.

Table 3. Chi-square or Fisher’s exact tests comparing the proportions of stems in different size classes from 1969 to 1996.



(Scolytidae), and root rot (Sherman and Warren 1988). In-creased tree densities due to fire suppression as well as nee-dle damage from air pollution may makePinus ponderosamore susceptible to stress (Savage 1994) and also less resis-tant to fire (Swezy and Agee 1991).Pinus ponderosais gen-erally fire resistant (Kilgore 1973a), so if it only kills smalltrees, then it is not a major factor in the decline of the spe-cies. The number of trees was too small to make statisticalcomparisons, but in unburned plots, 25% of the trees>40 cm in diameter died, while in burned plots 60% of thetrees >40 cm and 80% of the trees <40 cm (all <20 cm)died. Consequently, burning thins smaller trees and maycontribute to the mortality of larger individuals as well.

With the decline ofPinus ponderosa, species compositionis changing as the relatively open canopy is becoming even

more open than it was in 1969. Although tree and shrub di-versity decreased because of decreased equitability amongspecies (Roy 1997) the primary compositional change is thedecrease inPinus ponderosaand the increase ofAbiesconcolor. Similar patterns of decliningPinus ponderosaandincreasing densities of shade-tolerant trees such asAbiesconcolor have been found in the northern Sierra Nevada(Ansley and Battles 1998), and the San Jacinto Mountains ofsouthern California (Savage 1994). Savage (1994) alsofound that Q. kelloggii was resistant to pathogens andstresses that affected other trees, so its abundance in the SanJacinto Mountains was more stable.Quercus kelloggiiis

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also relatively stable in Sequoia National Park, where highmortality has been partially offset by ingrowth in burnedplots.

Our findings of decreased densities of most tree species inthe Ponderosa Pine Forest represent a reversal of the trendsreported by Vankat and Major (1978) that densities ofAbiesconcolor and Calocedrus decurrenswere increasing withfire suppression. Decreases in these two species have beenespecially large in burned plots because of mortality. In thecase of Calocedrus decurrens, numbers decreased in thesmallest size class as prescribed burning apparently in-creased mortality in smaller trees while maintaining largertrees in the canopy. This is similar to what is thought to haveoccurred with the low intensity fires that characterized thepresuppression fire regime. Decreased densities in the un-burned plots may be a result of disease or other density-dependent factors.

For the shrub layer, Vankat and Major (1978) reportedthat Arctostaphylos viscidadecreased with fire suppression,but we found that its cover was constant during this studyperiod. We also found thatCeanothus integerrimusgreatlyincreased since 1969, especially in burned areas, andKilgore (1973a) reported that the species responded favor-ably to fire as well.

Jeffrey pine forestWe relocated five plots in the Jeffrey pine forest, and they

indicate that the forest has become more dense, with a 58%(311 trees/ha) increase in total density and correspondingincrease in basal area of 15% (14 m2/ha; Table 1).Pinusjeffreyi decreased in density (11%; 22 trees/ha) and cover(33%; 9% absolute cover), but the importance percentage ofAbies concolordoubled, reflecting a 400% (267 trees/ha)increase in density.Chamaebatia foliolosaremained nearlyconstant in absolute cover, but a decrease in the coverof Arctostaphylos patulaled to an overall decrease inshrub cover (Table 2). Some plots changed little in speciescomposition (Fig. 1C), but in three plots the importance ofPinus jeffreyidecreased and importance percentages of ei-ther Abies concoloror Pinus contortaincreased. In the fourunburned plots, all species had increases in basal area(Fig. 4).

Vankat and Major (1978) stated that tree densities had in-creased more in the Jeffrey pine forest than in the other for-est types. We found thatAbies concolorand Calocedrusdecurrenshave continued to increase in density. Both spe-cies are shade tolerant and have seeds capable of germinat-ing on sites that are unburned (Adams et al. 1966; Kilgore1973a), so without fire these species might be expected toincrease until density-dependent mortality occurs. Despitethe increase inAbies concolor,its density remains lowerthan the other midelevation forests for either 1969 or 1996.Therefore, we hypothesize that densities have not reachedlevels necessary for self thinning. All but one of the plotsare located in the Kern River drainage where lower precipi-tation may slowAbies concolorregeneration.

White fir forestWe relocated 14 plots in the white fir forest, which de-

creased in total density (5%; 35 trees/ha), basal area (13%;22 m2/ha), and cover (15%; 16% absolute cover; Table 1).

Abies concolorremained the most important tree, althoughdensity decreased 16% (93 trees/ha). The density of theother three important canopy tree species increased, al-though their importance percentages remained nearly con-stant. Total absolute shrub cover increased dramatically(2–21%), as the result of increases ofCorylus cornutaMarsh andSymphoricarpossp., as well as colonization byeight taxa new to the plots (Table 2). In fact, new shrub spe-cies were found in 8 of the 14 plots. Despite stability in theoverall composition of the canopy trees in this vegetationtype, the ordination position of most plots generally movedaway from the central area, indicating increased heterogene-ity in this forest type (Fig. 1D). Seven of these nine plotshad newly colonized shrub species in 1996, suggesting thedecreased similarity among the plots was at least partly re-lated to the composition of the shrub layer.

Changes in the white fir forest also include a significantdifference in the distribution ofAbies concoloramong sizeclasses between 1969 and 1996 (Table 3). However, the dif-ferences in individual size classes were not significant in the2 × 2 table (adjustedα = 0.007, andχ2 42= . , df = 1, p >0.025 in the >50-cm size class). There is a trend, however,towards fewer trees in the three largest size classes(Fig. 2D). Ingrowth compensated for a high mortality rate(55%) in the 2.5- to 9-cm size class, but mortality rates of20, 50, and 20% in the three largest size classes, respectively,account for the apparent decreases in these size classes.

Changes differed between the eight unburned and sixburned plots (Fig. 5). The basal area ofAbies concolorde-creased in unburned plots and increased in burned plots, butdensity decreased much more in burned plots (33%; 183trees/ha) than in unburned plots (2%; 12 trees/ha). Amongthe differences between unburned and burned plots in otherspecies was an increase in the density ofCalocedrusdecurrens in burned plots resulting from ingrowth. Someburned plots were stable, while others shifted on the DCA(Fig. 1D), so burning did not produce consistent changes inthe composition of plots.

Our results differ from previous studies that documentedchanges due to fire suppression (Vankat 1970; Kilgore1973a; Kilgore and Briggs 1972; Parsons and DeBenedettii1979; Vankat and Major 1978) in that densities, especially ofAbies concolor, are no longer increasing. Ingrowth in un-burned plots nearly compensated for mortality, so densitiesof Abies concolorare nearly stable in the unburned plots. Inaddition, prescribed burning has reducedAbies concolordensity. Other studies have found that fire generally killsAbies concolorin smaller size classes (Kilgore 1973b; Lam-bert 1983), so burning may have contributed to the 55%mortality in the smallest size class. However, at least oneburned plot south of Crescent Meadow had scorched deadtrees >40 cm DBH, and the trend for decreased density inlarge size classes also suggests burning affects large trees.

Vankat and Major (1978) reported decreases inCeanothussp., andArctostaphylossp. in the shrub layer. In contrast, wefound an increase in shrub cover, mostly due to the increasein Corylus cornutacover. Burning may be related to the in-crease becauseCeanothus integerrimus, a species that in-creased in cover, germinates after fire (Kilgore 1973a), andfour of the five burned plots had new taxa, compared withonly four of the eight unburned plots.

© 1999 NRC Canada

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Giant sequoia grovesWe relocated 11 plots in giant sequoia groves that suggest

the groves were more open in 1996 than 1969 because of de-creases in total tree density (39%; 391 trees/ha) and cover(28%; 33% absolute cover; Table 1). An even larger de-crease occurred in basal area (62%; 588 m2/ha), but thebasal areas of largeSequoiadendron giganteawere overesti-

mated because circumferance measurements were not cor-rected where concave fire scars were present. Nevertheless,the combined basal area of the other species also decreased(26%; 35 m2/ha). Abies concolor decreased 40% (318trees/ha) in density as well. Although total shrub cover re-mained low (3%), five species colonized the plots, while onespecies was lost (Table 2).

Most individual plots changed in composition, but therewas only a small increase in the breadth of distribution ofplots on the DCA ordination, indicating little change in theoverall heterogeneity of the forest (Fig. 1E). The decreaseddensity ofAbies concolorwas not accompanied by a signifi-cant change in its size class distribution (Table 3); however,there was a trend towards a decrease in the two smallest sizeclasses (Fig. 2E), which had mortality of 43 and 39%,respectively.

Forest structure changed more in the seven burned plotsthan the four unburned plots (Fig. 5).Abies concolor, for ex-ample, decreased 18% (10 m2/ha) and 35% (40 m2/ha) inbasal area and decreased 21% (175 trees/ha) and 49% (370trees/ha) in density in unburned and burned plots, respec-tively. Mortality was higher in burned plots, and ingrowthwas greater in unburned plots. Density values forSequoia-dendron giganteumwere stable in both plot types with nomortality or ingrowth. DCA ordinations indicated that bothunburned and burned plots changed in species composition(Fig. 1E).

Previous studies reported that densities ofAbies concolorincreased as the result of fire suppression in the giant se-quoia groves (Vankat 1970; Kilgore 1973a; Vankat and Ma-jor 1978; Kilgore and Taylor 1979), but our data indicatethat this trend has reversed, even in unburned plots. Perhapsthe higher density in the giant sequoia groves than the whitefir forest has led to self thinning; Kilgore (1973a), for exam-ple, suggested that high densities facilitate the spread ofpathogens and insects. Of course the larger decrease in den-sities of Abies concolorin burned plots indicates that pre-scribed fire effectively reduced tree densities. Althoughmortality and burning cannot be directly correlated in thisstudy, prescribed fire has been shown to reduceAbiesconcolor densities in smaller size classes (Kilgore 1973b;Keifer and Stanzler 1995, Stephenson 1996).

Vankat and Major (1978) reported that shrub cover, espe-cially for Arctostaphylos patulaand Prunus emarginata(Hook.) Walp., decreased as the result of fire suppression.Our data do not show that prescribed fire reversed the de-cline of either species, as neither were recorded as present.Fire may have been important, however, in thatRosaspithamea, Ribes roezliiRegel, andSymphoricarpos albus(L.)S.F. Blake, all new to the plots, occurred only in two burnedplots (in the Muir Grove). This localized pattern of shrubcolonization is different from the more widespread coloniza-tion of shrubs in the white fir forest.

Previous researchers also viewed the absence ofSequoia-dendron giganteumreproduction as a change in the giantsequoia groves (Vankat and Major 1978). BecauseSequoia-dendron giganteumrequires exposed mineral soil for seed-ling establishment, fire suppression was believed to haveinhibited reproduction (Harvey and Shellhammer 1991). Al-though Sequoiadendron giganteumingrowth did not occurwithin the plots, we observed saplings in burned plots that

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were too small to be measured as ingrowth. Fire has been re-ported to increase the number of seedlings and saplings inother studies (Harvey and Shellhammer 1991; Stephenson1996).

High-elevation forests

Red fir forestIn the nine plots we relocated in the red fir forest, total

tree density decreased 24% (244 trees/ha) and tree cover de-creased 10% (8% absolute cover). Total basal area remainedalmost constant, the only vegetation type where this oc-curred (Table 1). Species composition remained nearly sta-ble, with the importance percentage ofAbies magnificaat90%, although its densities decreased 22 (189 trees/ha).Pinus contorta, is no longer present in the plots.Symphori-carpossp. was the only shrub species at both sampling dateswith 1% absolute cover in 1996 (Table 2).

Two of the three plots that experienced disturbance fromfire or windthrow shifted towards greater similarity with un-disturbed plots between 1969 and 1996 (Fig. 1F), resultingfrom shifts in composition toward increasedAbies magnificacover. Although the size class distributions ofAbiesmagnifica did not change significantly between 1969 and1996 (Table 3), there is a nonsignificant trend of fewer treesin the 2.5- to 9-cm size class (Fig. 2F). This possible de-crease would be accounted for by diameter growth, as mor-tality was only 4% in the size class. Ingrowth ofAbiesmagnificaalso occurred (Fig. 6).

The decrease inAbies magnificadensity represents a re-versal in the trend previously identified by Vankat and Ma-jor (1978), although they also reported a decrease inAbiesmagnifica in stands wherePinus contorta was present.Changes in individual plots occurred where there was eitherfire or windthrow, suggesting disturbance influences thecomposition of red fir forests. The scale of disturbance,however, affects the changes which occur (Vankat and Major1978; Pitcher 1987; Taylor 1993). Large fires or other stand-replacing disturbances cause a flush of ingrowth, producingeven-aged stands (Rundel et al. 1977; Pitcher 1987). Smallerscale disturbance events, such as the windthrow we ob-served, produce stands with a varied aged structure (Vankatand Major 1978; Taylor 1993).

Lodgepole pine forestsWe relocated five plots in the lodgepole pine forest, and

total tree density in the plots was unchanged (Table 1). Theimportance percentage ofPinus contortadecreased from 93to 88%, and density decreased 6% (60 trees/ha). The impor-tance percentage ofAbies magnificaincreased from 7 to12% with a doubling in density (60 trees/ha). In the shrublayer,Ribes montegeumMcClatchie is now the only shrub inthe plots (2% absolute cover) as two other species are nolonger present (Table 2). The size-class distribution ofPinuscontortadid not change, but there was a nonsignificant trendof more trees in the two largest size classes (Table 3,Fig. 2G), suggesting that mid-sized trees are gradually grow-ing into larger size classes. None of the plots burned, andmortality of Pinus contorta was higher than ingrowth(Fig. 6).

Vankat and Major (1978) had reported an increase inPinus contortadensity and an invasion byPinus contortainto herbaceous meadows adjacent to the lodgepole pine for-est. These changes were correlated with the end of sheepgrazing in the late 1800s, when sheep no longer ate andtrampled young conifer seedlings. Invasion of meadows alsomay have been enhanced by grazing-induced erosion thatlowered water tables. Our finding of a decrease in density ofPinus contortaindicates that the trend may have reversed.

Vankat (1970) placed six plots across meadow edges todocument any further shifts in the forest–meadow ecotone.Years with low precipitation have been correlated with theinvasion of meadows in Yosemite National Park (Helms1987), but we did not observePinus contortacolonizing

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410 Can. J. For. Res. Vol. 29, 1999

further into the meadows. In addition, changes among themeadow plots were too inconsistent to be from climatic fluc-tuations because only two of five plots had ingrowth (300trees/ha) along the forest edge, two exhibited no change, andthe fifth had mortality of 670 trees/ha. The reasons for thishigh mortality in one plot are unclear, although the Park Ser-vice cut young trees from some meadows in the early 1970s.

Subalpine forestTree importance percentages remained similar in the sub-

alpine forest (Table 1), but this has little meaning becausethe only two species sampled,Pinus balfourianaand Pinuscontorta,occur in different stands. A decrease in sheep graz-ing in 1890 may have led to increasedPinus balfourianaes-tablishment between 1890–1895 (Vankat and Major 1978),but densities have been stable since then (Vankat and Major1978; Table 1). Increases of 8% (12 m2/ha) in basal area and16% (4% absolute cover) in cover are due to the growth of ex-isting trees since mortality and ingrowth were absent (Fig. 6).

We found a decrease in tree density in all woodland andforest types except the Jeffrey pine, lodgepole pine, and sub-alpine forests. Where the density of trees decreased, thecomposition of species remained largely unchanged, exceptin the ponderosa pine forest. We also found that tree densi-ties decreased in both unburned and burned plots but morein burned plots, especially those in midelevation forests. Al-though many species specific changes are evident, their in-terpretation is limited in some woodland and forest typeswhere there were relatively few plots.

Our study plots represented samples of locally variableforest patches, and such patches are dynamic, developing in-dividually through time (Pickett and White 1985). Accord-ing to this patch dynamic concept, disturbances initiate newpatches of regenerating forest, and therefore, changes in thedisturbance regime result in shifts in the proportion ofpatches at different stages of development. Vankat and Ma-jor (1978) examined the impact of two broad-scale changesin Sequoia National Park’s disturbance regime. First, firesuppression removed fire as an important factor in patchvariability in woodlands and forests and resulted in increasesin tree densities in several forest types. Second, livestockgrazing, an addition to the disturbance regime, apparentlycaused a flush of tree regeneration in some woodlands andforests when it ended. Because these changes in the distur-bance regime occurred throughout the park, woodlands andforests may have become more homogenous.

More recently, small-scale prescribed burning increasedvegetation heterogeneity. Because prescribed burning hasbeen used only in scattered areas in the park, the woodlandsand forests now contain some patches influenced by pre-scribed burning while the composition of other patches morereflects past fire suppression and grazing. Furthermore, pre-scribed fires appear to have burned at various levels of in-tensity, resulting in even greater patch heterogeneity.

Increased heterogeneity is illustrated on several of theDCA ordinations in which plots sampled in 1996 are fartherapart than the same plots sampled in 1969. The fact thatburned and unburned plots usually shifted individually on

the ordinations indicates a diversity of changes likely relatedto site-specific conditions.

Vankat and Major (1978) previously documented in-creases in tree density in many woodlands and forests anddecreases in shrub cover in several forests. Our findings in-dicate that these trends have reversed, as tree densities havedecreased or remained stable in all woodlands and forestsexcept the Jeffrey pine forest, and shrub cover has increasedin several forests.

Although we have documented these reversals, we cannotdetermine when they occurred. Vankat and Major (1978)dated the beginning of increased tree densities to near 1900;however, densities could have stabilized or started decreas-ing anytime after that initial increase.

Despite the difficulty in determining the timing of thetrend reversals, some of the causes of the reduced tree densi-ties are clear. Minnich et al. (1995) proposed that competi-tion and stress from high tree densities in southernCalifornia would cause densities to decrease where fire sup-pression continued. Similar factors may limit tree densitiesin Sequoia National Park as well because tree densities de-creased in both unburned and burned plots. In addition,given that tree densities decreased more in burned than un-burned plots in the forest types, prescribed burning is an-other obvious factor in reduced densities.

Reducing tree densities in the forests was one of the goalsof the prescribed burning program in Sequoia National Park(Butts 1985), and this goal has been accomplished withoutmajor changes in species composition. Prescribed firescaused increased mortality in dominant species as well as inspecies of lower importance. Therefore, prescribed burningappears to create more open forest conditions while conserv-ing the characteristic species of each forest type.

Similar changes due to fire suppression, grazing, and pre-scribed burning reported for Sequoia National Park may alsooccur in other national parks in the western United Statesand Canada, at least where similar shifts in disturbance re-gimes have occurred. More generally, the trends illustratethe dramatic effects that changes in management policy anddisturbance regimes may have on vegetation. Of course, fu-ture changes in natural resource policies have the same po-tential to produce broad-scale change, as grazing and firesuppression once did. Unfortunately, factors that are beyondthe control of local land managers, such as air pollution andclimate change, may have similar far-reaching effects.

We thank the National Park Service for permission towork in Sequoia National Park, and Annie Esperanza wasespecially helpful with logistical support. Linda Mutch, NateStephenson, Patricia Muir, and an anonymous reviewer pro-vided constructive criticism of the manuscript. The Aca-demic Challenge Grant Program of the Department ofBotany, Miami University provided financial support.

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Reversal of human-induced vegetation changes in Sequoia National Park, California