A MAMMALIAN PREDATOR–PREY IMBALANCE: GRIZZLY BEAR … · A MAMMALIAN PREDATOR–PREY IMBALANCE:...

August 2001 947ECOLOGICAL ISSUES IN CONSERVATION

947

Ecological Applications, 11(4), 2001, pp. 947–960q 2001 by the Ecological Society of America

A MAMMALIAN PREDATOR–PREY IMBALANCE: GRIZZLY BEAR ANDWOLF EXTINCTION AFFECT AVIAN NEOTROPICAL MIGRANTS

JOEL BERGER,1,4 PETER B. STACEY,2 LORI BELLIS,3 AND MATTHEW P. JOHNSON3

1Program in Ecology, Evolution, and Conservation Biology, University of Nevada, Reno, Nevada 89512 USA, andWildlife Conservation Society, P.O. Box 340, Moose, Wyoming 83012 USA

2Department of Biology, University of New Mexico, Albuquerque, New Mexico 87131 USA3Department of Environmental and Resource Sciences, University of Nevada, Reno, Nevada 89512 USA

Abstract. Because most large, terrestrial mammalian predators have already been lostfrom more than 95–99% of the contiguous United States and Mexico, many ecologicalcommunities are either missing dominant selective forces or have new ones dependent uponhumans. Such large-scale manipulations of a key element of most ecosystems offer uniqueopportunities to investigate how the loss of large carnivores affects communities, includingthe extent, if any, of interactions at different trophic levels. Here, we demonstrate a cascadeof ecological events that were triggered by the local extinction of grizzly bears (Ursusarctos) and wolves (Canis lupus) from the southern Greater Yellowstone Ecosystem. Theseinclude (1) the demographic eruption of a large, semi-obligate, riparian-dependent herbi-vore, the moose (Alces alces), during the past 150 yr; (2) the subsequent alteration ofriparian vegetation structure and density by ungulate herbivory; and (3) the coincidentreduction of avian neotropical migrants in the impacted willow communities. We contrastedthree sites matched hydrologically and ecologically in Grand Teton National Park, Wyo-ming, USA, where grizzly bears and wolves had been eliminated 60–75 yr ago and moosedensities were about five times higher, with those on national forest lands outside the park,where predation by the two large carnivores has been replaced by human hunting and moosedensities were lower. Avian species richness and nesting density varied inversely withmoose abundance, and two riparian specialists, Gray Catbirds (Dumetella carolinensis) andMacGillivray’s Warblers (Oporornis tolmiei), were absent from Park riparian systems wheremoose densities were high. Our findings not only offer empirical support for the top-downeffect of large carnivores in terrestrial communities, but also provide a scientific rationalefor restoration options to conserve biological diversity. To predict future impacts, whetherovert or subtle, of past management, and to restore biodiversity, more must be known aboutecological interactions, including the role of large carnivores. Restoration options withrespect to the system that we studied in the southern Greater Yellowstone Ecosystem aresimple: (1) do nothing and accept the erosion of biological diversity, (2) replace naturalcarnivores with human predation, or (3) allow continued dispersal of grizzly bears andwolves into previously occupied, but now vacant, habitat. Although additional science isrequired to further our understanding of this and other terrestrial systems, a larger con-servation challenge remains: to develop public support for ecologically rational conser-vation options.

Key words: biodiversity; conservation; extinction; Grand Teton National Park; Greater Yellow-stone Ecosystem; grizzly bears; moose; neotropical migrant birds; restoration ecology; riparian veg-etation; trophic cascades; wolves.

INTRODUCTION

Among organisms conspicuously absent from manyof this planet’s ecological communities are two typesof large mammals. In the Neotropics, sites that oth-erwise appear virtually intact have been referred to as‘‘empty forests’’ because of the depletion of game

Manuscript received 2 August 1999; revised 21 March 2000;accepted 3 April 2000; final version received 30 August 2000.Forreprints of this Invited Feature, see footnote 1, p. 945.

4 Address correspondence to: P.O. Box 340, Moose, Wy-oming 83012 USA. E-mail: [email protected]

(Redford 1992). In western North America, many pub-lic lands designated as wilderness lack grizzly bears(Arctos ursus) and wolves (Canis lupus). Althoughthese carnivores still occur across vast tracts of Asianand North American boreal forests and the Arctic, theirlocalized extinction has been both thorough and swift(within the last 100 yr), to the extent that they arecurrently absent from 95–99% of their former rangeswithin the contiguous United States and Mexico (Clarket al. 1999). Despite a few well-publicized natural re-colonizations, such as wolves into Glacier NationalPark (Boyd et al. 1994), most ecosystems and nature

948 INVITED FEATURE Ecological ApplicationsVol. 11, No. 4

reserves are more likely to lose, rather than gain, largecarnivores, irrespective of recent highly publicized ef-forts such as the restoration of wolves in Yellowstoneand central Idaho (Phillips and Smith 1996).

The widespread disruption or total cessation of in-teractions between large mammalian carnivores andprey species has had numerous consequences, includ-ing the local irruption of native and domestic herbi-vores (McCullough 1997, Sinclair 1998), mesocarni-vore release (Crete and Desrosiers 1995, Crooks andSoule 1999, Terborgh et al. 1999), site-specific changesin prey behavior (Berger 1998, 1999), and, perhaps,the facilitation of Lyme disease at the local level (Wil-son and Childs 1997). The scale at which such pred-ator–prey disequilibriums have occurred provides ex-ceptional opportunities to learn more about ecologicalprocesses (Berger and Wehausen 1991), including thelong-standing debate over the role of top carnivores inthe regulation of prey populations (Terborgh 1987,Wright et al. 1994), and whether top-down or bottom-up effects play a larger role in biological organization(Paine 1966, Polis and Strong 1996). Although mostresearch on trophic cascades has focused on aquatic ormarine systems (Power 1992, Estes et al. 1998), andhas involved heterotherms and invertebrates (Spillerand Schoener 1994, Carter and Rypstra 1995), the re-cent localized, but widespread, extinction of large ver-tebrate carnivores throughout much of their ranges al-lows us to examine how predation and food can struc-ture terrestrial communities and how species losses canaffect the maintenance of biological diversity.

There is a problem, however. Because so many ofthis planet’s terrestrial systems have already been great-ly modified, it is often impossible to find appropriateregions where human and nonhuman effects can bedisentangled. For instance, although overabundant el-ephant populations in Botswana affect avian diversityin woodland habitats (Herremans 1995), the degree towhich humans have facilitated these elephant popula-tions is unclear (Gibson et al. 1998). In other areas,including both boreal and neotropical systems, humansand nonhuman predators compete for the same mam-malian food base (Gasaway et al. 1992, Jorgensen andRedford 1993), although human patterns of selectionmay have more important demographic effects on theprey themselves (Ginsberg and Milner-Gulland 1994,Berger and Gompper 1999). From an ecological per-spective, however, what may be more critical thanknowing about predation per se is understanding thecommunity-level consequences of the removal of pre-dation as a selective force.

In many systems, the speed of ecological change hasbeen accelerated with human intervention (Sutherland1998). A notable example concerns demographic re-sponses of ungulates to the loss of large carnivoreswhich, in turn, through population irruptions, often af-fects habitat structure through herbivory-related pro-

cesses (Wagner and Kay 1993, Naiman and Rogers1997). With increasingly dense prey populations, it isnow known that vegetation alterations through brows-ing also change patterns of habitat selection in neo-tropical migrant birds (McShea et al. 1995, McSheaand Rappole 1997). Because habitat structure is oftenone of the best predictors of avian diversity (MacArthurand MacArthur 1961, Cody 1974, Wiens 1989, Jackson1992), the possibility that predation by large mam-malian carnivores, or its absence, can affect other com-ponents of biological diversity, such as birds, presentsserious challenges for restoration and conservation bi-ologists alike.

Our intent here is twofold: first, to assess evidenceof top-down effects of large carnivores by considering(1) their influences on a semi-obligate riparian spe-cialist, moose (Alces alces), (2) herbivory-related ef-fects of moose on willow community structure, and (3)interrelationships between structural modification ofhabitat and avian species diversity and abundance; andsecond, to suggest how our results can bear on theoutcomes of different conservation options.

BIODIVERSITY AND THE MOOSE–GRIZZLY

BEAR–WOLF SYSTEM AS A MODEL

We used the well-understood moose–grizzly bear–wolf system to explore potential interrelationships be-tween human intervention and trophic levels on long-term patterns that connect carnivores and biologicalorganization. Moose have a circumpolar distribution,occurring from Mongolia and the northern Great BasinDesert throughout boreal, subarctic, and Arctic regionsof North America, Europe, and Asia (Franzmann andSchwartz 1998). Moose populations are controlled bypredation and, in its absence, by food (Messier 1991,Gasaway et al. 1992, Orians et al. 1997), and theyachieve states of relative equilibrium that depend oncarnivore abundance (Fig. 1A). Where intact carnivorecommunities exist, such as in Alaska and the Yukon,predation on moose is intense, with up to 90% of thejuveniles being killed annually by grizzly bears andwolves (Gasaway et al. 1992, Bowyer et al. 1998).However, after humans remove bears and wolves,moose population growth is near maximum (Orians etal. 1997). In the relatively unique situation on Isle Roy-ale, Michigan, USA, where only moose and wolvesinteract, the situation differs because of weather andother factors (Peterson 1999).

Moose may have important localized effects on eco-systems (Pastor et al. 1993, Post et al. 1999), partlybecause they consume large quantities of woody shrubsand young trees including aspen, willow, and cotton-wood, and because they achieve high densities (Hous-ton 1968) that, in riparian zones, may exceed 20 in-dividuals/km2 (Fig. 2) for up to 5–6 months per year.In Alaska and the Yukon, moose density is affected by


FIG. 1. (A) Conceptualization of moose population states under equilibrium and disequilibrium conditions. Each scenariohas been supported empirically (for case studies, see Orians et al. [1997]). (B) Patterns of colonization and general populationtrends of moose in the Jackson Hole region of the Greater Yellowstone Ecosystem from the 1840s to 1990s in relation tothe loss of grizzly bears and wolves. (Source: J. Berger, unpublished data from trapper records, historical journals, andWyoming Department of Game and Fish files. Solid circles reflect estimates from the literature.)

predation (see Fig. 1A), but not in many areas of theRocky Mountains where large carnivores are extinct.

The moose–grizzly–wolf system has critical advan-tages that make it useful as a model for exploring issuesrelated to biodiversity and conservation. First, thesetwo predators influence moose populations in the ab-sence of humans (Gasaway et al. 1992), and, thus, theirloss has resulted in substantial demographic change(Ballard and Van Ballenberghe 1998). Both speciesprey on adult and juvenile moose, with the bear beingmore of a specialist on calves (Gasaway et al. 1992).Although the cessation of Native Americans as majorpredators (Kay 1994a, b) of moose also corresponds,at a broad level, to the localized extinctions of thesetwo carnivores, from a demographic point of view,moose populations increase when grizzly bears andwolves are reduced or absent (Boertje et al. 1996).

Second, when released from predation pressure (Fig.1), native species, including cervids other than moose,may achieve extraordinarily high densities (Alversonand Waller 1997, Schmitz and Sinclair 1997). Althoughdomestic species may severely impact sensitive habi-tats (Knopf and Cannon 1982, Kauffman and Krueger1984, Belsky et al. 1999), little is known about theeffects of colonizing native browsers such as moose.This is a particularly critical issue in riparian systemsbecause, in arid zones like the American West, theseareas represent ,1% of the total area but may containup to 80% of the local avian diversity (Ohmart 1994,Stacey 1995, Dobkin et al. 1998). Additionally, at alocal scale, abundant herbivores like elk may affectavian diversity through structural effects on vegetation(Jackson 1992). It seems reasonable to expect not onlythat intense herbivory might affect riparian biodiver-


FIG. 2. Top: moose and riparian zones in Grand Teton National Park. Bottom: nine of a group of more than 30 moosein a riparian zone (April 1995).

sity, but also that the moose–grizzly bear–wolf systemholds promise to further knowledge about biologicalorganization as it may relate to large carnivores andterrestrial community dynamics.

HISTORY, BACKGROUND, AND STUDY AREA

We studied interactions among moose populations,the structure of their major winter food supply in ri-parian zones, and avian species diversity in Grand Te-ton National Park (GTNP) and adjacent public lands

managed by Bridger-Teton National Forest (BTNF) inthe Jackson Hole region of Wyoming. Wolves havebeen extinct in this system since the 1930s (J. Craig-head and F. Craighead, personal communication), andgrizzly bears have not been known at either of our studyareas since 1936 or earlier (Hoak et al. 1981).

During the 1800s, moose were rare in much of west-ern North America (Karns 1998), and 150 yr ago theywere absent from the Jackson Hole region at the centerof the Greater Yellowstone Ecosystem (GYE). Moose


appeared from the north, their existence first being not-ed in Yellowstone Park in the 1880s (Schullery andWhittlesey 1995). With the reduction of wolves andgrizzly bears and the absence of significant predationby Native Americans, moose populations irrupted (Kay1994b). Based on data from Houston (1968), currentlocal moose densities and population size in the Jack-son Hole area appear to have reached a plateau, andhave remained relatively stable for about the last 30yr, a pattern characteristic of populations that reachdemographic ceilings (Fig. 1B).

The two primary types of federal lands (GTNP andBTNF) that exist in Jackson Hole are managed verydifferently. On BTNF lands, moose are subject to hu-man harvest; between 1971 and 1991, .10 800 werekilled (Houston 1992). In contrast, GTNP moose arenot managed, and this fundamental difference in federalpolicy has resulted in variation in moose density.

Moose depend upon willow as a primary food sourceduring winter (Peek 1998). If populations in GTNPhave reached the point at which food is limiting, thenriparian vegetation may be altered more when mooseoccur at high densities than when they are controlledby predation (whether by humans or natural carni-vores). Alternatively, moose may have little, if any,impact on riparian vegetation.

We evaluated these possibilities by contrasting ef-fects of moose herbivory on riparian willow commu-nities, using three areas within GTNP (with no humanor natural predation for ;60 yr) and three similar sitesin BTNF (where human hunting has replaced that bynative carnivores). Moose wintering densities variedby about five times as a consequence of the differentmanagement regimes (mean density is 5.2 individuals/km2 at GTNP sites and 1.1 individual/km2 at BTNFsites). Thus, our treatment was predation (absent orpresent), with each area containing three replicated ri-parian communities, all matched with the followingcharacteristics: (1) willow riparian habitat lacking cot-tonwood, spruce and aspen gallery forests; (2) majorwillow species being Wolf’s (Salix wolfii), Geyers (S.geyerianna), and Booth’s (S. boothii); (3) wet meadowspecies composition dominated by willow spp., sedges,and grass; (4) some standing water with evidence ofwetland histosol soils and a lack of rodent burrows andtunnels; (5) situated in different drainages; and (6)watersheds within a 25 km radius of, and flowing into,the Snake River. The specific sites for the ‘‘Predation’’treatment (all in BTNF), were: Grey’s River(43803.3709 N, 110849.4209 W), Fall Creek (43820.9909N, 110849.6499 W), and Ditch Creek (43841.5029 N,110834.8149 W). For the ‘‘No Predation’’ treatment (allin GTNP), sites were: Triangle X (43847.1399 N, 110833.1799 W), Pacific Creek (43852.3949 N, 110829.2569 W),and Buffalo Fork (43850.3039 N, 110829.2769 W).

METHODS

Sampling

Vegetation.—Our sampling occurred during springand summer 1998. At each site we measured vegetationheight, volume, and structure using three 100-m tran-sects, $200 m apart. Twenty 1 3 1 m plots per transectwere selected using a random number table, 10 plotssituated on each side of the transect line. Vegetationtype, percent cover, height, and vertical structure wererecorded at each 1 3 1 m plot. For each vegetation/ground cover type (water, bare ground, grass and sedge,herbs, willows, and other shrubs), percent cover andfour height categories (0–0.5 m, 0.5–1 m, 1–2 m, and$2 m) were recorded. Canopy volume at each heightcategory was estimated by the ‘‘stacked-cube’’ method(Kus 1998) by placing PVC pipe on the ground to mapquadrat boundaries and then by connecting lengths ofthe PVC to enclose the vertical cover. Willow volumewas the product of density and cover. We gaugedbrowsing intensity by recording in each plot the pro-portion of willow stems .10 mm in diameter that wereeither browsed or unbrowsed, and either alive or dead.

Birds.—Nesting densities were estimated using spot-mapping techniques (Ralph et al. 1993). Briefly, twoobservers slowly walked along the vegetation transectsand recorded all birds perched or singing within 50 mof the transect line. The location of each bird was re-corded on maps, and behaviors were noted, as well asthe presence or absence of a second bird or young.Three censuses at each transect were conducted duringthe breeding season (15 May–1 July) between sunriseand 1100, with a minimum interim of 7 d betweensampling.

To derive spatial configurations, convex polygonswere drawn around all observations of the same speciesat the same general location on different dates. A sec-ond polygon was then drawn around the mapped lo-cations of all individual species, this area being thatwhich was actually surveyed during the census. Theresulting data provided evidence of both presence/ab-sence for each species, and relative density (e.g., threeterritories within a surveyed area of 1200 m2 5 0.25pairs/100 m2). This approach has been used to estimateavian densities in riparian zones (e.g., Ammon and Sta-cey 1997), and it enabled us to calculate species di-versity (both richness and evenness) within the sur-veyed regions.

Moose.—Our starting premise, that predation releaseof moose has occurred with the localized extinction ofwolves and grizzly bears, would be untrue if other car-nivores prey on moose. Although these two carnivoreswere absent, both black bears (Ursus americana) andpumas (Felis concolor) were not, and they are capableof killing moose (Ballard and Van Ballenberghe 1998).If black bears and pumas prey on moose, juvenile re-cruitment could be affected. Without investigation, this


FIG. 3. Contrasting effects of the presenceand absence of grizzly bears and wolves onmoose calf survival at sites in Alaska, Canada,Norway, and the southern Greater YellowstoneEcosystem in Wyoming. Dotted lines indicatemean survival. Sources (in order from left toright) are: Berger et al. (1999), Swenson et al.(1999), Testa (1998), Bowyer et al. (1998), Bal-lard et al. (1991), Larsen et al. (1989), and Gas-away et al. (1992).

would render impossible any conclusions about wheth-er predation had (or had not) been relaxed with the lossof grizzly bears and wolves.

We examined this issue by using two analyses toevaluate whether the local extirpation of wolves andgrizzly bears affected juvenile recruitment. First, wecontrasted patterns of juvenile survival at sites withand without grizzly bears and wolves. Second, becausefood and other factors also affect juvenile survival, weassessed pregnancy rates. Pregnancy and other corre-lates of fecundity are useful indices of food quality andquantity (Franzmann and Schwartz 1985, Gasaway etal. 1996). We relied on fecal metabolites (Montfort etal. 1993) to assess pregnancy in radio-collared moosefrom 1995 to 1997, and we subsequently followed thefates of neonates from known mothers. This procedureenabled us to assess not only whether a recruitmentfailure occurred, but also whether it might have arisenas a consequence of reduced pregnancy rates or otherfactors. Details of laboratory and field protocols areprovided in Berger et al. (1999).

Analyses.—Effects of predation treatment on bothvegetation and avian parameters were examined bygeneral linear models (Norusis 1997) with transectnested within site (i.e., a repeated measure), or a two-way repeated-measures ANOVA to evaluate two-di-mensional vegetation profiles. The number of obser-vations in our data set was 18 (two treatments, threesites for each, and three transects per site), but for allcomparisons in this nested design, the degrees of free-dom were one (treatment) and four (site). Pairwiseposthoc multiple comparisons were conducted withDuncan’s Multiple Range Test, with similar groupingsdenoted by the same letter and contrasting ones (P ,0.01) with different letters.

RESULTS

Assumptions and their tests

Prior to interpreting our empirical data from the field,we made two important assumptions about the effects

of large carnivores on moose. First, predation is relaxedafter the extinction of wolves and grizzly bears. Sec-ond, after release from predation, populations will ap-proach or exceed an ecological food ceiling. If theseassumptions are supported, then it seems reasonable toexpect that other levels of biological organization maybe affected by the loss of these two carnivores.

Predation is relaxed.—Although grizzly bears werenot entirely extirpated from Grand Teton National Park,they were highly restricted to northern portions, wheredensities were extremely low and verified observationsaveraged less than one every three years (Hoak et al.1981). Most importantly, they did not occur at our spe-cific study sites. Wolves were extirpated in the 1930s(Fig. 1), and only in 1998 did dispersal from Yellow-stone National Park result in some recolonization intothe Tetons and other areas of Jackson Hole (J. Berger,unpublished data).

If the local extinction of grizzly bears and wolveshad an impact on moose, then differences in juvenilesurvival should have existed between areas where thesecarnivores were present and absent. If, however, me-socarnivores such as black bears and pumas killedmoose at roughly similar rates, or other factors com-promised juvenile survival, then it would be difficultto argue that predation has been relaxed. However, theavailable data support the premise that grizzly bearsand wolves have major impacts on neonatal moose sur-vival, and that their loss results in a threefold increasein mean neonate survival (Fig. 3).

In absence of predation, herbivore populations ap-proach or exceed ecological capacity.—Abundant lit-erature supports the idea that ungulate populations ex-perience density dependence, particularly in the ab-sence of predation (Clutton-Brock and Albon 1989,Sinclair 1995). However, because species includingmoose may respond differently to varied ecologicalconditions or reduced predator densities (Peterson1999), it is important to assess the evidence with re-


TABLE 1. Summary of vegetation contrasts involving browsing intensity on willow com-munities at sites with and without predation on moose.

Willow status

Predation on moose

Human None F P

Alive, not browsedAlive, browsedDead, not browsedDead, browsed

0.53 (0.15)0.22 (0.15)0.12 (0.07)0.13 (0.08)

0.10 (0.11)0.59 (0.10)0.01 (0.01)0.29 (0.09)

18.567.73

13.745.04

,0.0001,0.0025,0.0002,0.0128

Note: Mean values (and 1 SD) reflect proportions (N 5 360).

TABLE 2. Statistical summary of effects of moose browsing(proportion of stems browsed relative to total available) inriparian willow communities at sites in different drainageswithin 25 km of the Snake River in northwest Wyoming.

Sources ofvariation df MS F P

PredationSiteVHCSite 3 VHCError

1433

0.1900.0742.4840.0600.008

23.449.11

307.097.41

0.00010.00010.00010.0003

Note: The table presents results of two-way repeated-mea-sures ANOVA of the effects of predation (present/absent) andstudy site on vegetation height categories (VHC).

TABLE 3. Duncan’s grouping examining the effects of pre-dation (present/absent) on moose browsing at designatedvegetation height categories.

Predation

Vegetation height category (m)

0–0.5 0.5–1 1–2 21

PresentAbsent

aa

bc

cd

ee

Note: Different letters indicate significant differences at P, 0.01.

spect to GTNP moose. Although we cannot directlytest whether moose in GTNP are at or above their foodsupply, pregnancy rates can be used as a surrogate mea-sure that reflects body condition, which, in turn, is di-rectly related to the food supply (Franzmann andSchwartz 1985).

First, we examined changes in pregnancy as a func-tion of time. If food were limiting, then pregnancy ratesshould decline over time. The evidence is consistentwith this interpretation; pregnancy rates dropped as thepopulation apparently stabilized (Fig. 1B), from ;90%in 1966 (N 5 41; Houston 1968) to 75.5% in 1997 (N5 49; Berger et al. 1999), a change that approachessignificance (Gadj 5 3.36; P , 0.10). Second, popu-lations below a ‘‘food ceiling,’’ that is, in a growthphase, tend to have relative high pregnancy rates. How-ever, the current pregnancy rate (;75.%) for GTNPmoose is comparatively low, and it ranks in the bottom85th percentile of North American populations (Bergeret al. 1999). Hence, the low pregnancy rates in GTNPmay be a consequence of populations released frompredation and currently occurring at high densities.

Riparian willow communities and moose density

As previously indicated, moose densities varied, be-ing almost five times higher in the absence of predation(GTNP) than on adjacent forest lands where predationby humans occurred (mean density was 5.2 individuals/km2 at GTNP sites and 1.1 individuals/km2 at BTNFsites). No details are available on the degree of simi-larity between willow communities in Grand Teton Na-tional Park and adjacent forest lands prior to moose

irruptions. We assumed, however, that because otherlarge-bodied browsers were absent from these ripariansystems, willow structure was comparable.

Did vegetation at local sites that varied in moosedensity also differ? Areas that varied in predation treat-ments differed significantly in both mean willow heightand density (F5,17 5 7.95, P , 0.01; F5,17 5 8.43, P ,0.01, respectively). Where moose densities were lim-ited by humans, willows were taller (2.17 6 0.32 m,x 6 1 SD) and their percentage of volume was greater(0.72 6 0.05%) than in areas lacking predation (1.776 0.45 m and 0.63 6 0.09%, respectively). Addition-ally, the proportion of willow stems .10 mm that werebrowsed, and either alive or dead, was also associatedwith moose density (Table 1). These results substantiatethat moose density affected vegetation.

Predation treatment also affected vegetation struc-ture through influences on vegetation height categories,VHC (Tables 2 and 3). At both the ground level, whereheavy winter snowfall reduces willow availability, andin the vegetation canopy above which moose regularlyfeed, differences in the proportion of browsed willowstems were not detectable (Table 3). However, at in-termediate vegetation layers, moose densities had dra-matic effects on the proportion of browsed willows(Table 3). These data suggest that (1) treatment effectswere robust; and that (2) our measures are consistentwith predictions about VHC as a bioassay of browsing-related herbivory. Our findings that variation in moosedensity failed to produce browsing-related differencesbetween treatments for both the ground and top veg-etation layers, but not at intermediate sites within thewillow canopy, reinforce the idea that moose per se,rather than other factors, were responsible for the ob-served variation.


TABLE 4. Summary of mean migrant breeding bird densities per 100 3 100 m transect at riparian sites differing in moosedensities, with predation on moose as the treatment.

Species

No. bird pairs/ha

Human predationon moose

No predationon moose F P

Calliope Hummingbird (Stellula calliope)Willow Flycatcher (Empidonax trailii)Gray Catbird (Dumetella carolinensis)Yellow Warbler (Dendroica petechia)

2.00 (0.87)0.78 (0.44)0.88 (0.92)3.78 (0.97)

0.22 (0.44)0.55 (0.53)0.00 (0)2.33 (1.00)

10.672.405.123.21

0.00040.09940.00960.0453

Wilson’s Warbler (Wilsonia pusilla)Yellowthroat (Geothlypis trichas)MacGillivray’s Warbler (Oporornis tolmiei)Black-headed Grosbeak (Pheucticus melanocephalus)

0.11 (0.33)0.00 (0)0.22 (0.44)0.77 (0.44)

0.67 (0.87)0.11 (0.33)0.00 (0)0.22 (0.44)

1.861.000.803.00

0.17600.45820.57050.0552

Song Sparrow (Melospiza melodia)Fox Sparrow (Passerella iliaca)Lincoln’s Sparrow (Melospiza lincolnii)White-crowned Sparrow (Zonotrichia leucophrys)

1.67 (0.70)0.44 (0.53)0.22 (0.44)0.00 (0)

1.11 (0.93)0.11 (0.33)0.66 (0.71)0.55 (1.13)

3.910.850.910.82

0.02450.54050.50380.5612

Avian species diversity and willows

To determine whether moose density was also as-sociated with the distribution or abundance of the ri-parian avian community, we used five assessments ofavian distribution and abundance (Table 4). All weresimilarly affected by moose density, with each beingstatistically greater where moose densities were regu-lated by predation: (1) species richness of breedingbirds (N 5 23 vs. 18; F 5 14.44, P , 0.003); (2) nestingdensity (F 5 7.78, P , 0.002); (3) Shannon’s diversityindex (F 5 13.10, P , 0.004); (4) Hill’s Diversitymeasure 1 (F 5 10.30; P , 0.008); and (5) Hill’s Di-versity Measure 2 (F 5 7.10; P , 0.021). Where moosedensities were high, the nesting density of many mi-grants including Willow Flycatchers, Calliope Hum-mingbirds, Yellow Warblers, Fox Sparrows, and Black-headed Grosbeaks was substantially reduced (Table 4).Two other species, Gray Catbirds and MacGillivray’sWarblers, were absent. Overall, ;50% of the riparianwillow bird species were reduced or absent from sitesinside GTNP where moose were protected from pre-dation and thereby attained high local densities. Thefact that typical riparian species were present at all sites(e.g., Yellow and Wilson’s Warblers; see Table 4 forothers) substantiates that the samples were derivedfrom the same ecological pool.

At a more refined scale, we asked whether levels inthe willow canopy were differentially affected bymoose density and whether they, in turn, influencedeither avian richness or density. Thus, these two mea-sures were contrasted between the species guilds thatnest primarily near the ground (GR: Gray Catbirds andthe four species of sparrows) and those that nest aboveground (AG: Calliope Hummingbird, Willow Flycatch-er, all warblers, and Black-headed Grosbeak). Becauseour sample sizes were further subdivided, our resultsare qualitative only. For AG, mean richness was greaterat low than at high moose density (4.0 vs. 2.3 species),as was also the case in all other contrasts (AG densitywas 7.1 vs. 4.1 pairs/10 000 m2, respectively; GR rich-

ness was 2.2 vs. 1.5 species; GR density was 3.2 vs.2.4 pairs/10 000 m2 for low vs. high moose density).Notable, however, is that the apparent magnitude ofdifferences due to moose density was less in groundnesters, as would be expected given the lack of sig-nificant herbivory-related effects at ground level rel-ative to higher levels in the canopy (Table 3). Overall,species richness and density were reduced to a greaterextent in AG than in GR, patterns that are consistentwith the idea that moose browsing at higher densitiesnegatively impacts neotropical migrants.

Although these data suggest that moose browsingshaped avian communities at a microgeographicalscale, the link between the structural modification ofriparian willows at our sites and avian diversity hadbeen uncertain. A direct relationship between willowvolume (mean using the proportion of live willowstems) and avian species diversity (Y ) existed, althoughmuch variance remains unexplained (Y 5 0.51X 1 0.33,n 5 18, r2 5 0.24; P , 0.03). Nonetheless, it appearsthat avian species diversity is partially affected throughstructural modifications of the willow canopy by moosewhose densities, in turn, are controlled at our studyareas by humans, either by total protection (park) orhunting (national forest lands).

DISCUSSION

Large herbivores and trophic cascades.—Our find-ing that large herbivores affect both riparian vegetationand associated avian communities is not entirely novel;somewhat similar relationships have been detected atsites grazed by domestic livestock (Dobkin et al. 1998,Sanders and Edge 1998, Skagen et al. 1998). However,what differs about our results is that they are based onthe irruption of a native, rather than an introduced,species, a browser, after natural predators were re-moved from the system. Still, in today’s highly mod-ified world, the detection of indirect effects stemmingfrom human actions, including the loss of carnivores,and those of other processes has become increasingly


FIG. 4. Overview of conservation optionsand the linkages among biological tiers of or-ganization in a terrestrial ecosystem with largecarnivores.

difficult. For instance, where large, terrestrial carni-vores have been extirpated, one consequence has beenthe release of herbivores, with concomitant habitat al-terations (Wagner and Kay 1993, Kay 1994b, McSheaand Rappole 1997). However, factors other than theloss of carnivores, such as climatic change and fire,confound straightforward analyses (Houston 1982,Singer et al. 1998, Boyce and Anderson 1999).

Our analyses, which focused exclusively on areaswithin the Jackson Hole area of the Greater Yellow-stone Ecosystem, support the idea that a dynamic chainof interactions involving multiple tiers of biologicalorganization were set in motion some 60–100 yr ago,principally by the removal of large predators (Fig. 4).Among the key events were: (1) human decisions toexterminate large carnivores, especially wolves (Murie1940, Phillips and Smith 1996), but also grizzly bears(Craighead 1979) in Yellowstone Park per se and inadjacent regions; (2) a resultant growth of an apparentlow-density moose population (Fig. 1) that, althoughit began expanding from 1880 to 1910, irrupted par-tially due to a dampened effect on juvenile mortality(Fig. 3); (3) increasing herbivory in riparian willowcommunities at sites lacking predation or hunting (Ta-ble 1); (4) structural modification of these communities(Tables 2 and 3); and (5) decreased avian richness anddiversity (Table 4).

Although the idea of top-down regulation of com-munities by carnivores, especially in complex terres-trial ecosystems, has been both attractive and contro-versial (Polis and Strong 1996), our data offer supportfor its importance at least in this system. In other ter-restrial regions, the plausibility of trophic cascades hasbeen noted, as have confounding variables that cloudthe interpretation for top-down regulation. For exam-ple, jaguars may or may not regulate mammalian her-bivores in the Neotropics (Terborgh 1987, Wright et al.1994), but the nature of comparative evidence from

geographically disparate sites without adequate repli-cation limits broad generalization (Terborgh et al.1999). In temperate boreal systems, where at least onelong-term data set is available, the evidence is stronger,in that wolves affect moose that subsequently limitproductivity of fir trees (Post et al. 1999). However,many of the studies that target community regulationby terrestrial carnivores have not considered bottom-up effects (Boyce and Anderson 1999) or questionedthe role of potentially confounding variables.

Alternative hypotheses and potential confoundingvariables.—Although we argue that the loss of large,terrestrial carnivores eventually caused a loss in bio-diversity (Fig. 4), other factors may be involved andmay account for the differences in the avian commu-nities in riparian areas where there is, or is not, pre-dation on moose.

1. Climate or local weather conditions could varybetween study regions.—Although explanations alongthese lines have been used to suggest possible causallinks for elk population changes through time in Yel-lowstone’s Lamar Valley or the loss of riparian vege-tation (see Krausman 1998), these do not apply to ourstudy system. Our sites were all situated in drainageswithin a 25 km radius of the Snake River, they wereindependent of each other at the scale at which weworked, and all have experienced similar edaphic, hy-drological, and weather conditions.

2. Microhabitat variation in willow community struc-ture could exist.—This is a clear possibility, althoughwe suspect that it is not a strong one, because studyareas were matched for additional characteristics in-cluding the lack of both conifer and deciduous treecanopies, the amount of edge and adjacent habitats,and a minimum of roads and human disturbances. Inaddition, all sites shared the same basic suite of riparianbird species. Nevertheless, our data on avian speciesdiversity were based on a single year; additional mul-


tiyear monitoring will be necessary to affirm our initialfindings.

We presume, but do not know, that differential useof willows by moose in Grand Teton National Park mayhave begun around 1950, which is when the Park wasexpanded to cover areas that included our riparian zonestudy areas. Although moose populations in JacksonHole were already in an exponential phase of growthby 1950 (Fig. 1A), human harvesting of moose on ad-jacent national forest was still occurring. There is noreason to expect that moose density was not in theprocess of changing between areas of total (park) pro-tection and areas with human hunting. In essence, itseems more likely that variation among willow com-munities was due to differential herbivory and not mi-crogeographical variation.

3. Predation by humans and predation by nonhu-mans have different demographic consequences.—There is no doubt truth in this premise, and no reasonto expect that moose cannot also be affected. For in-stance, males and females of sexually dimorphic un-gulates (including moose) not only respond differentlyto types of predation but also select different habitatsand may even consume foods that vary in quality (Clut-ton-Brock et al. 1982, Miquelle et al. 1992). Life his-tory or behavior responses may also vary with pre-dation intensity and degree of past selection (Cheneyand Seyfarth 1990, Berger 1991). For example, withthe relaxation of predation for ;10 generations, mooseno longer responded to Ravens, which typically asso-ciate with grizzly bears and wolves. At sites wherethese predators were not extirpated, however, moosewere highly responsive (Berger 1999). Nonetheless, de-spite notable differences in the age and sex of ungulateprey taken by human vs. nonhuman hunters (Ginsbergand Milner-Gulland 1994), the fact remains that formoose and other species, harvesting by humans resultsin striking reductions in population density (Crete1987, Cederlund and Sand 1991).

Three points bear on this issue with respect to moosein the southern Greater Yellowstone Ecosystem. First,wolves and grizzly bears are capable predators ofmoose (Ballard and Van Ballenberghe 1998), but in theabsence of predation (whether by humans or nonhu-mans), moose populations increase rapidly and attainrelatively higher densities (Gasaway et al., 1992, Or-ians et al. 1997). Second, densities in the Jackson Holearea have been mediated either through total protection(e.g., park) or through human predation (on adjacentnational forest lands). Third, as shown in this study,levels of herbivory in riparian zones are affected bymoose density.

4. Herbivory-related effects may have been due toungulates other than moose.—Although this possibilityalso exists, it is unlikely. First, despite controversy overthe effects of elk (Cervus elaphus) browsing on willowsin Yellowstone National Park (Singer et al. 1994), there

has been little evidence of elk at our GTNP study zonesduring winter over the last 40 yr (Houston 1968, 1992;J. Bohne and G. Frahlick, personal communication),although exceptions exist. For instance, during the fivewinters from 1995 to 1999, 30–50 elk spent about sixweeks in 1997 at one of the GTNP study areas. How-ever, even this number of elk represented ,5% of theavailable ungulate biomass during that winter, and,1% during the other four winters (J. Berger, unpub-lished data). The other major ungulates in the GreaterYellowstone Ecosystem (bison, Bison bison; mule deer,Odocoileus hemionus; pronghorn, Antilopcapra amer-icana; and bighorn sheep, Ovis canadensis) use dif-ferent habitats than do moose. Second, feeding groundsprovide alternate foods (alfalfa pellets, primarily) dur-ing winter, and these concentrate and sustain a large(,15 000–20 000) elk population (Boyce 1989). Thistype of management by humans has resulted in a defacto release of winter herbivory by elk in the ripariansites that we studied. Thus, we have not observed thecompetitive displacement of moose by elk that has ap-parently occurred in Yellowstone National Park whenelk densities are sufficiently high (Tyers 1996, Singeret al. 1998). Hence, it is likely that the effects on veg-etation and bird communities that we report here wereattributable to moose and not to other ungulates.

Conservation options within the context of large car-nivores, biodiversity, and public lands.—It is ironicthat, in a protected area such as GTNP, where the intenthas been to enhance biological diversity through hands-off management, the opposite has occurred. Protection,after the localized extinction of large carnivores, hasresulted in the decline of avian taxa within the protectedboundaries; in contrast, lands outside the park haveinvolved multiple uses (administered by the USDAForest Service), where the active management of wild-life has resulted in greater levels of diversity (Table 4).Although grizzly bears and wolves, until very recently,have also been extinct on much of the national forestlands adjacent to Jackson Hole, moose have been con-trolled in these regions, as is evident by the harvest of.10 000 animals during a 20-yr period (Houston 1992).Upon closer inspection, however, it is probably notuncommon that subtle, but ecologically important,changes in landscapes occur, even after complete pro-tection. National parks in western North America aregood examples because, despite protection that can betraced to a century or longer, numerous species ex-tinctions still occur (Newmark 1987) and ecologicalprocesses are altered in the absence of management.

Where restoration is an active goal, a hands-off pol-icy may be inadequate. As our data show, a cascade ofevents triggered by the loss of large carnivores can leadto changes at multiple trophic levels, including a dim-inution in avian diversity. Nevertheless, in the absenceof detailed biological information, options for ecolog-ical restoration both in the southern Greater Yellow-


stone Ecosystem and elsewhere (Boyce 1998, Sinclair1998) will continue to depend largely on guesswork orlocal politics. This emphasizes the need for adminis-trative support in many parks for the acquisition ofbaseline data and the performance of credible science(Wagner et al. 1995).

Our data, although specific to a system in the RockyMountains, offer a basis to guide future conservationplanning and action. Where conservation objectives areto maintain or promote biotic diversity, several optionsexist (Fig. 4). The first is to reintroduce or to allowcontinued dispersal of large carnivores into previouslyoccupied, but now vacant, habitat. This currently isoccurring in formal recovery areas, for both grizzlybears and wolves, although localized conflicts involv-ing livestock management on public lands complicatethe restoration process (Niemeyer et al. 1994, Mattsonet al. 1996). On lands outside of U.S. national parks,for instance, hunting for large herbivores by the publicentails access to areas where recovering carnivores mayinadvertently (or purposefully) be shot. A relationshipalso exists between carnivore mortality and road den-sity (Mattson et al. 1998). Yet, there is also some prom-ise. In areas of the Canadian Rockies, aspen regener-ation may be occurring because of the reduction of elkby wolves (White et al. 1998). Nevertheless, effects ofcarnivores on abundance of prey species are likely tovary from area to area; what occurs in Canada may notapply to the Greater Yellowstone Ecosystem.

Where wolves and moose co-exist, elk are the pri-mary prey of wolves, but whether wolves or grizzlybears or some synergistic effect of these two large car-nivores will ultimately reduce moose population den-sity in GTNP remains unknown. One step in garneringopportunities to examine a longer term role of top car-nivores in terrestrial ecosystems, irrespective of local-ity, will require continually educating the public, bothwithin local regions of carnivore reintroduction and atbroader geographical levels (Smith et al. 1999).

A second, and perhaps equally controversial, optionfor the restoration of naturally functioning ecosystemsmight be to replace patterns of carnivore predation byhuman control actions (Fig. 4), a practice that we be-lieve should be directed at a broader spectrum of theprey population and less at trophy males per se. Clearly,this option does nothing to facilitate the restoration oflarge carnivores, but it does raise a practical means bywhich to look beyond carnivores per se and to considerinstead how carnivores may be represented as majoragents of change at a landscape level.

However, the mere suggestion of human harvest iscontroversial, particularly when a desired vegetationcondition can be considered arbitrary (Boyce 1998,Kay 1998). Additionally, humans and carnivores oftenhunt very differently, particularly with respect to thesex and age of their prey (Ginsberg and Milner-Gulland1994, Berger and Gompper 1999). If humans are to

simulate the predation practices of nonhuman carni-vores, then serious consideration of how prey har-vesting by humans should proceed will be necessary.If the goal is merely density reduction of ungulateswithout regard to juvenile recruitment, population tra-jectories, or how or where the sexes segregate (Bowyeret al. 1996), then the status quo may be acceptable. Wesuspect that some intermediate practice between thesetwo strategies will be the most appropriate.

Finally, one can do nothing. If this latter path ispreferred, our results suggest that there may be a con-tinued de facto erosion of avian diversity due to mul-tiple cascades of the effects of past manipulations atdifferent trophic levels. The extent to which suchchange may apply to other taxa awaits further inves-tigation.

ACKNOWLEDGMENTS

We thank the following people and agencies for support:J. Bohne and D. Brimeyer (Wyoming Department of Gameand Fish), S. Cain and R. Schiller (National Park Service),W. Weber and J. Robinson (Wildlife Conservation Society,WCS), and the Universities of Nevada and New Mew Mexico.Funding was provided by the Denver Zoological Foundation,the National Science Foundation, and WCS. We greatly ap-preciate the insights and critical comments of J. H. Brown,S. Cain, C. Cunningham, D. Finch, B. Kus, D. Ligon, B.Miller, R. Noss, K. Redford, F. Wagner, and an anonymousreviewer.

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