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Forest Ecology and Management 260 (2010) 983–993

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Forest Ecology and Management

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Forecasting vole population outbreaks in forest plantations: The rise and fall of amajor mammalian pest

Thomas P. Sullivana,∗, Druscilla S. Sullivanb

a Department of Forest Sciences, Faculty of Forestry, University of British Columbia, 2424 Main Mall, Vancouver, BC, Canada V6T 1Z4b Applied Mammal Research Institute, 11010 Mitchell Avenue, Summerland, BC, Canada V0H 1Z8

a r t i c l e i n f o

Article history:Received 5 April 2010Received in revised form 11 June 2010Accepted 16 June 2010

Keywords:AbundanceFeeding damageForecasting outbreaksForest plantationsGrass seedingLong-tailed volesMicrotus longicaudusPopulation dynamicsTree seedlings

a b s t r a c t

Voles of the genera Microtus and Myodes feed on tree seedlings planted on cutover forest land in temperateand boreal forests of North America and Eurasia. This damage may have serious economic implicationsas well as limit regeneration of appropriate tree species in certain forest ecosystems. Prediction of volepopulation outbreaks and feeding damage to forest plantations, across even a limited geographic range,has yet to be achieved in North America. Thus, a major objective was a detailed analysis of changes inpopulation dynamics of long-tailed voles (Microtus longicaudus), and to test three hypotheses (H) thatvole populations would: (H1) rise and fall in accordance with the abundance of herbaceous plants (grassesand forbs) during early vegetative succession after forest harvesting, (H2) be positively associated withgrass-seeded sites; and (H3) incidence of feeding damage to seedlings would be positively associatedwith vole abundance. Voles were live-trapped for 6 years (2004–2009) from the time of harvesting onintensive sites, as well as surveyed over a range of extensive sites. Population numbers were relatedto habitat characteristics and tree damage in young forest plantations near Golden, British Columbia,Canada.

Populations of long-tailed voles were low in the first two years after harvest with mean numbers<5–15/ha. Annual peaks of 49–84 voles/ha were recorded in 2006. In the fourth year (2007) after har-vesting, numbers of voles declined on two of three sites, deepened in 2008 and reached extirpation in2009. On the extensive sites, vole numbers increased 4.6–5.3 times from 1–2 to 3–6 years post-harvestbefore declining thereafter. Crown volume index of grasses and herbs, volume and abundance of downedwood, total species richness of vascular plants, and structural diversity of herbs were important habitatvariables. Vole numbers were higher on those sites seeded with pasture grasses and forbs. There was asignificant positive relationship of tree mortality and abundance of voles (Microtus) across a relativelywide geographic area.

This study is the first relatively long-term analysis of changes in population dynamics of the long-tailed vole and the predictions of H1 and H2 seemed to be supported. The positive relationship (H3)of the incidence of overwinter damage to trees and vole abundance is the first such analysis for forestplantations, on harvested sites, in North America. At 3–4 years post-clearcut harvesting is a critical timefor population buildups of voles and subsequent damage to plantation trees. Seeded grass species clearlycreate optimum habitat conditions for voles, generating population densities up to 30–50 voles/ha, whichis in the range of a “high” damage risk to seedlings. Risk ratings (voles/ha) for feeding damage to treeswere low (<7), moderate (7–34), high (35–88), and very high (>88).

© 2010 Elsevier B.V. All rights reserved.

1. Introduction

The problem of feeding damage to forest and agricultural cropsby herbivorous small mammals has a long history in temperate andboreal ecosystems of North America and Eurasia (Myllymäki, 1977;Gill, 1992). In forestry, voles of the genera Microtus and Myodes

∗ Corresponding author. Tel.: +1 604 822 6873.E-mail address: [email protected] (T.P. Sullivan).

are considered the major mammalian species affecting coniferousand deciduous tree plantations in North America (Radvanyi, 1980;Sullivan et al., 1990), Europe (Hansson, 1991; Huitu et al., 2009), andAsia (Shu, 1985). Populations of some species of voles tend to havecyclic fluctuations in abundance in northern latitudes with a peakevery 3–5 years, although these periods may be interspersed withannual fluctuations in abundance (Krebs and Myers, 1974; Taitt andKrebs, 1985; Korpimäki and Krebs, 1996). Abundance of Microtuspopulations and degree of damage is usually highest in early succes-sional habitats that develop after forest harvesting by clearcutting

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984 T.P. Sullivan, D.S. Sullivan / Forest Ecology and Management 260 (2010) 983–993

(Hansson, 1989; Sullivan and Sullivan, 2001), wildfires (Fisher andWilkinson, 2005), and in old fields (perennial grasslands) undergo-ing afforestation (Bergeron and Jodoin, 1989; Ostfeld and Canham,1993). Grasses, forbs, and shrubs in these habitats provide food andcover for Microtus voles (Batzli, 1985; Getz, 1985; Ostfeld, 1985).

Voles will feed on tree seedlings and saplings, with highest dam-age during winter months of peak years in abundance (Huitu et al.,2003). These rodents feed on bark, vascular tissues, and sometimesroots of trees (Baxter and Hansson, 2001). This damage resultsin direct mortality from girdling and clipping of tree stems orreduced growth of surviving trees which have sub-lethal injuries.The fertilization regime of nursery-raised seedlings enhances theirpalatability and nutrition, thereby predisposing them to prefer-ential feeding over wildlings that arise from natural regeneration(Sullivan and Martin, 1991). Voles also preferentially feed on par-ticular tree species (Bucyanayandi et al., 1990). Thus, in terms ofconservation and sustainability of temperate and boreal forests,this feeding damage may limit regeneration of appropriate treespecies in certain forest ecosystems. In addition, this damageincreases the cost to reforest these stands to “free growing status”,decreases net productive forested area, and results in loss of meanannual tree growth increment.

Amount of vegetative cover is an important component of habi-tat quality and selection (Getz, 1985), population density (Birneyet al., 1976), and demography (Adler, 1987) of Microtus species.In addition, some populations of microtines need a thresholdlevel of cover and plant production to generate population cyclesor fluctuations (Birney et al., 1976; Laine and Henttonen, 1983).The herbaceous component, particularly grasses, seems crucialto maintaining suitable habitat conditions of food and cover forvoles (Reich, 1981; Adler, 1987). The long-tailed vole (Microtuslongicaudus) and the meadow vole (M. pennsylvanicus) are majorconsumers of tree seedlings in inland areas of the Pacific Northwest(PNW) of North America (Sullivan et al., 1990).

The long-tailed vole occurs throughout most of the westernUnited States (U.S.) and Canada to eastern Alaska, and may occupyvarious successional forest seres, shrubs, and riparian sites (Smolenand Keller, 1987). Clearcut sites in forests have created variable-quality habitats for long-tailed voles in Montana (Halvorson, 1982),Alaska (Van Horne, 1982), and British Columbia (B.C.) (Sullivan etal., 1999; Sullivan and Sullivan, 2001). Habitats after forest har-vesting or wildfire disturbance are typically dominated by forbs,grasses, and shrubs for up to 5–10 years, depending on the ecosys-tem, as well as coarse woody debris left over from harvesting andthe original forest. The meadow vole is distributed throughoutCanada, northern and eastern U.S., and into Mexico, and commonlyfound in grasslands (Reich, 1981).

An additional floral component of some harvested and burnedsites is the seeding of forage species to prevent soil erosion, sitedegradation, and invasion of noxious plant species. However, thereis much disagreement as to the validity and necessity of this prac-tice (Beyers, 2004; Beschta et al., 2004), and the role of seededpasture plants that may provide ideal habitat for buildups of volepopulations. Also, the subsequent dispersal of these grasses andforbs may alter the regenerating ecosystems in unfavorable ways.

Microtus species may be the most studied mammals, particu-larly with respect to investigations of population regulation (Taittand Krebs, 1985). Various hypotheses invoking predation, food sup-ply, interspecific competition, dispersal, quality of habitat patches,changes in aggressive behavior from stress, genetically controlledbehavioral mechanisms, and so on, have been postulated and tested(Krebs, 1996). Unfortunately, there has been a dearth of studiesattempting to forecast when and where vole population outbreakswill occur in relation to feeding damage to plantations on cutoverforest land. Incidence of damage is highly variable across temperateand boreal forests despite the fact that many multi-annual fluctu-

ations of vole populations tend to be spatially synchronous acrossthese landscapes (Huitu et al., 2003; Sundell et al., 2004). Someprogress has been achieved at local scales relating vole abundanceto degree of damage (Hansson, 1986; Ostfeld and Canham, 1993;Sullivan and Sullivan, 2001), as well as at a nation-wide scale inFinland (Huitu et al., 2009). However, for species with apparentannual versus multi-annual fluctuations in abundance, such as M.longicaudus, we need to know the status of populations in differentsuccessional stages and years, in order to forecast those sites andtimes that trees are particularly susceptible to feeding damage.

To date, prediction of vole population outbreaks and feedingdamage to forest plantations, across even a limited geographicrange, has yet to be achieved, at least in North America. However,this problem may be related to the lack of simultaneous informationon vole dynamics, vegetative succession and other site parameters,and damage to seedlings. Thus, a major objective of this study wasto provide a detailed analysis of changes in population dynamicsof long-tailed voles over a relatively long-term period (6 years)in young forest plantations in the Rocky Mountains near Goldenin south-central B.C. This region has a 15-year history of planta-tion failures owing to severe feeding damage from the long-tailedvole. In addition, we tested the hypotheses (H) that populationsof M. longicaudus would: (H1) rise and fall in accordance with theabundance of herbaceous plants (grasses and forbs) during earlyvegetative succession after forest harvesting; and (H2) be positivelyassociated with grass-seeded sites. A further hypothesis was that(H3) incidence of feeding damage to seedlings, across a wide geo-graphic range, would be positively associated with vole abundance.

2. Methods

2.1. Study areas

Several study areas were located at Glenogle Creek and RothCreek, 25 km east of Golden, B.C. (51◦18′N; 116◦45′W). Theseareas were within the Interior Douglas-fir (IDFdm), Montane Spruce(MSdk), Interior Cedar-Hemlock (ICHmk), and Engelmann Spruce-Subalpine Fir (ESSFdk) biogeoclimatic (BEC) zones (Meidinger andPojar, 1991). Topography ranged from hilly to very steep terrain at1125–1540 m elevation.

The upper IDF and MS have a cool, continental climate with coldwinters and moderately short, warm summers. The average tem-perature is below 0 ◦C for 2–5 months, and above 10 ◦C for 2–5months, with mean annual precipitation ranging from 30 to 90 cm.Open to closed mature forests of Douglas-fir (Pseudotsuga menziesii)cover much of the IDF zone, with even-aged post-fire lodgepolepine (Pinus contorta) stands at higher elevations. The MS landscapehas extensive young and maturing seral stages of lodgepole pine,which have regenerated after wildfire. Hybrid interior spruce (Piceaglauca × P. engelmannii) and subalpine fir (Abies lasiocarpa) are thedominant shade-tolerant climax trees. Douglas-fir is an importantseral species in zonal ecosystems and is a climax species on warmsouth-facing slopes in the driest ecosystems. Trembling aspen (Pop-ulus tremuloides) is a common seral species and black cottonwood(Populus trichocarpa) occurs on some moist sites (Meidinger andPojar, 1991).

The ICH has an interior, continental climate with cool wet win-ters and warm dry summers. Mean annual temperature rangesfrom 2 to 8.7 ◦C. The temperature averages below 0 ◦C for 2–5months and above 10 ◦C for 3–5 months of the year. Mean annualprecipitation is 500–1200 mm, 25–50% of which falls as snow.Upland coniferous forests dominate the ICH landscape and com-prise the highest diversity of tree species of any zone in B.C. Westernred cedar (Thuja plicata) and western hemlock (Tsuga heterophylla)dominate mature climax forests, with lodgepole pine, white spruce,


T.P. Sullivan, D.S. Sullivan / Forest Ecology and Management 260 (2010) 983–993 985

Engelmann spruce, their hybrids, and subalpine fir common inthese stands (Meidinger and Pojar, 1991). The ESSF zone is theuppermost forested zone in the southern interior of B.C., and ischaracterized by a relatively cold, moist climate with cool shortsummers, and long cold winters. Mean annual temperatures rangefrom −2 to +2 ◦C and mean monthly temperatures are below 0 ◦Cfor 5–7 months and above 10 ◦C for only 2 months (Meidinger andPojar, 1991). Most precipitation (50–70%) occurs as snow.

Prior to harvesting, all stands at the Golden study areas werecomposed of a mixture of lodgepole pine with variable amounts ofDouglas-fir, western red cedar, spruce, and subalpine fir. Averageages of lodgepole pine ranged from 80 to 120 years and for Douglas-fir and other conifers ranged from 120 to 220 years. Average treeheights ranged from 10.5 to 19.5 m for lodgepole pine and from 16.7to 27.5 m for Douglas-fir and other conifer species. There were nosite preparation treatments on any of these harvested sites, priorto planting.

Study areas relating the incidence of tree damage to number ofvoles included Golden (9 sites), Summerland (6 sites), and GavinLake (6 sites). The Summerland sites were located 25 km west ofSummerland, B.C. (49◦40′N; 119◦53′W) in the MSdm subzone, andthe Gavin Lake sites were located 75 km northeast of Williams Lake,B.C. (52◦29′N; 121◦45′W) in the Sub-boreal Spruce (SBSdw) subzone(Meidinger and Pojar, 1991).

2.2. Study and sampling designs

The long-term monitoring of vole populations, on selected sites,used grid systems of live-traps, via mark and recapture sampling,to provide detailed measurements of population dynamics. Threesites in the MSdk, that were clearcut harvested in the winter of2003–2004, were used for long-term grid sampling of changes inlong-tailed vole populations. These sites were selected on the basisof operational scale, reasonable proximity to one another, and weresampled since the time of harvesting. All sites were far enoughapart (0.75–3.0 km) to be statistically independent. An extensiveapproach was also needed, whereby many different plantationscould be surveyed to determine relationships between vole abun-dance and site characteristics. The 27 sites for index-line samplingof voles and site characteristics were all those units that wereavailable in the overall study area and represented the four BECsubzones, and were clearcut harvested from 1992 to 2004. Sim-ilarly, the 15 sites for index-line sampling of voles in grass vs.non-grass habitats were all the available units for this comparison,and were clearcut harvested from 1997 to 2004. These sites pro-vided a range of grass habitat conditions on landings, skid trails, androadsides. Much of the clearcutting practice was salvage harvestingof lodgepole pine from stands of mountain pine beetle (Dendroc-tonus ponderosae)-killed and susceptible trees.

2.3. Long-term sampling of vole populations

Vole populations were sampled at 4-week intervals from May toSeptember each year from 2004 to 2009. Live-trapping grids (1 ha)had 49 (7 × 7) trap stations at 14.3-m intervals with 1 or 2 Long-worth live-traps at each station. Number of traps at a station wasdependent on number of voles. Traps were supplied with wholeoats, a slice of carrot, and cotton as bedding. Each trap had a 30-cm × 30-cm plywood cover for protection from sunlight (heat) andprecipitation. Traps were set on the afternoon of day 1, checked onthe morning and afternoon of day 2 and morning of day 3, and thenlocked open between trapping periods. All animals captured wereear-tagged with serially numbered tags, breeding condition noted,weighed on Pesola spring balances, and point of capture recorded.The duration of the breeding season was noted by palpation ofmale testes and the condition of mammaries of the females (Krebs

et al., 1969). A pregnancy was considered successful if a femalewas lactating during the period following the estimated time ofbirth of a litter. Animals were released on the grids immediatelyafter processing. All handling of animals was in accordance withthe principles of the Animal Care Committee, University of BritishColumbia.

2.4. Vole population dynamics

Abundance estimates of long-tailed voles were derived from theJolly–Seber (J–S) stochastic model (Seber, 1982; Jolly and Dickson,1983), with small sample size corrections (Krebs, 1991). Num-ber of voles captured was used as the population estimate forthe first and last sampling weeks, and for the enumeration of therelatively uncommon heather voles and red-backed voles, whenthe J–S estimate was not calculated. Jolly trappability was calcu-lated according to the estimate discussed by Krebs and Boonstra(1984). There were six summer (May–September) and five winter(October–April) periods in this study.

Mass at sexual maturity was used to determine age classesof voles. The percentage of sexually mature animals was used todetermine the mass limitations for juveniles, subadults, and adultsassuming that juveniles were seldom, if ever, sexually mature; that<50% of the subadults in the upper mass class were mature; andthat at least 50% of the adults were sexually mature in the low-est mass class. Voles were classified as juvenile (includes juvenileand subadult classes pooled) or adult by body mass (long-tailedvoles: juvenile = 1–30 g, adult ≥ 31 g). Juveniles were considered tobe young animals recruited during the study. Recruits were definedas new animals that entered the population through reproductionand immigration.

Measurements of recruitment, number of successful pregnan-cies, and early juvenile survival were derived from the sample ofanimals captured in each trapping session and then summed foreach summer period. Early juvenile survival is an index relatingrecruitment of young into the trappable population to the numberof lactating females (Krebs, 1966). A modified version of this indexis number of juvenile animals at week t divided by the number oflactating females caught in week t–4. Mean survival rates (28-day)for the six summer and five winter periods were estimated fromthe Jolly–Seber model.

2.5. Index-line sampling of vole populations

An index-line system of live-traps was calibrated with a 1-hagrid system to provide a conversion factor for index-lines to giveaccurate measurements of vole abundance in many different sites(Petticrew and Sadleir, 1970). Superimposed on the center row ofeach of the three long-term sampling grids was an index-line oftraps consisting of 4 traps at each of the 7 stations. Trap set-up wasidentical to that described for the grids. The index-line traps wereset on the afternoon of day 1, checked on the morning and after-noon of day 2 and morning of day 3, and then locked open betweentrapping periods. The grid traps were then set on the subsequent3rd and 4th nights. Traps were then locked open between trappingperiods.

All small mammals (except shrews and weasels) captured wereear-tagged and immediately released at the point of capture asoutlined in the grid sampling procedure.

This approach allowed calculation of a regression relationshipbetween the estimated abundance of voles captured on an index-line to the estimated abundance of voles captured on a 1-ha gridsystem, throughout the summer and fall seasons of 2006.

For the extensive survey, one index-line was installed in eachsite and allowed to pre-bait for 4 weeks prior to the actual samplingof voles. Traps were supplied with whole oats and cotton and locked



open for the pre-bait period. For the survey, index-line traps wereset on the afternoon of day 1, checked on the morning and afternoonof day 2 and morning of day 3, and then picked up and moved to thenext site for a pre-bait period. Animals captured were processed inan identical manner to the grid sampling procedure.

The Petersen method of population estimation was used forvoles captured on index-lines (Krebs, 1999). This technique is amark and recapture method based on a single period (first night) ofmarking animals and second single period (second night) of recap-turing individuals. To control for the effect of seasonal change invole numbers due to reproduction and immigration over the sum-mer months of 2006, a monthly conversion factor was applied topopulation estimates so that the estimate for each index-line sam-ple was equivalent to the September period, which had the highestvole numbers. This conversion allowed us to sample sites withindex-lines during all summer–fall months (June–October) and stillbe able to accurately compare a site sampled in June with onesampled in October. The monthly changes in vole numbers fromthe three long-term sampling grids in 2006 were used to calcu-late the conversion factors. For example, mean abundance of volesincreased by a factor of 3.75 times from June to September, 2.77times from July to September, 1.75 times from August to Septem-ber, with no change (1.00 times) in abundance from September toOctober.

2.6. Grid and index-line sampling in grass and non-grass habitats

Three sites were selected that had grass-seeded habitats andthree sites that had little or no grass. A 1-ha live-trapping grid wasinstalled in each site and long-tailed voles were sampled over 8trapping periods from May to September 2005 and May to June2006. Additional grass and non-grass habitats were sampled bypermanent index-lines from May to September 2008 and May toJune 2009 (7 trapping periods). Methods of capture and process-ing of animals were identical to those described for the long-termsampling of voles.

One index-line was installed in each grass and non-grass habi-tat within a given site and allowed to pre-bait for 4 weeks priorto the actual sampling of voles. A total of 15 sites were sampledwith index-line surveys in 2007 and 2008. Traps were suppliedwith whole oats and cotton and locked open for the pre-bait period.For the survey, index-line traps were set on the afternoon of day1, checked on the morning and afternoon of day 2 and morning ofday 3, and then picked up and moved to the next site for a pre-baitperiod. Animals captured were processed in an identical manner tothe grid sampling procedure.

2.7. Vole abundance and tree mortality

Data relating vole abundance to mortality of planted trees werederived from three study areas in B.C. (Golden, Summerland, GavinLake) where the number of voles per ha was known in October,from standard 1-ha live-trapping grids, in a given year. At Golden,one-year-old nursery-raised trees were planted operationally onnine sites from 2002 to 2005. One hundred sample seedlings (lodge-pole pine, Douglas-fir, or interior spruce) were chosen randomly oneach site. At Summerland (1998) and Gavin Lake (1994), nursery-raised one-year-old lodgepole pine seedlings were planted in lateOctober-early November in groups of 10 × 10, one group of 100seedlings in each corner of a 1-ha trapping grid on each of sixsites at each study area. The total sample of trees exposed to voleson each grid was 400, which at a planting density every 3 m wasapproximately 1200 trees/ha. This density was within the rangeof typical new plantations of lodgepole pine in the interior of B.C.Clipping of terminal and lateral shoots, as well as gnawing onstems, was recorded for each sample seedling in the spring (May)

of each year. Removal of the terminal shoot was considered mor-tality unless another vigorous lateral shoot was available to replaceit. Girdling and removal of ≥50% of stem bark and vascular tissueswere also considered seedling mortality. Lodgepole pine, followedby Douglas-fir, are the most susceptible coniferous tree species tofeeding damage by Microtus spp. in the interior of B.C. (Sullivan etal., 1990). Overwinter feeding damage to trees (percentage mortal-ity) by voles was then related to the October population estimate.

2.8. Vegetation and site characteristics

At 5 of the 7 trap stations along each index-line or centerline of a sampling grid, a 3-m × 3-m plot for sampling shrubs andtrees and a 1-m × 1-m plot for sampling herbs was installed (afterStickney, 1980). Herb and shrub-tree layers were subdivided intoheight classes: 0–0.25, 0.25–0.50, 0.50–1.0, 1.0–2.0, 2.0–3.0, and3.0–5.0 m. A visual estimate of percentage ground cover was madefor each species/height class combination within the appropriatenested subplot. These data were then used to calculate crown vol-ume index (m3/0.01 ha) for each species. The product of percentcover and representative height gave the volume of a cylindroidwhich represented the space occupied by the plant in the commu-nity. Crown volume index values were then averaged by speciesfor each plot size, and converted to a 0.01-ha base to produce thevalues given for each species and layer (herbs, shrubs-trees). Totalpercentage cover for each layer was also estimated for each plot.Plant species were identified according to Parish et al. (1996) andHitchcock and Cronquist (1973). Species richness and species diver-sity (Shannon-Wiener index) were calculated for these data (Krebs,1999). Sampling was done in July–August 2006, 2007, and 2008.

Downed wood was recorded along two transect lines of 20 meach along each index-line on the 27 sites investigating habi-tat characteristics. As each piece was encountered, the followingattributes were recorded: (a) species, (b) diameter where linecrosses wood (cm), and (c) hardness (5 decay classes). Volumeof downed wood (m3/ha) was calculated by the method of VanWagner (1968). The physical characteristics of each index-line setof plots were recorded with respect to aspect, slope, site position,and other ecologically relevant features. Sampling was done in July-August 2006.

2.9. Seeding of forage species

Seeding of landings, road-sides, and skid-trails with forage grassand forb species for slope stabilization and erosion control was con-ducted, as an operational practice, on some harvested sites (Fig. 1).Typical pasture/forage seed mixtures include introduced speciesof: orchard grass (Dactylis glomerata), timothy (Phleum pratense),red fescue (Festuca rubra), crested wheatgrass (Agropyron crista-tum), red top (Agrostis alba), alfalfa (Medicago sylvatica), and clover(Trifolium pratense).

2.10. Statistical analysis

A one-way ANOVA was used to determine the effect of time(years) since clearcut harvesting and biogeoclimatic subzone onvole numbers in plantations (Zar, 1999). A linear regression analysiswas used to determine the relationship of vole numbers on index-lines to numbers on a 1-ha grid system, as well as the relationship oftree seedling mortality to number of voles. This regression analysiswas also used to relate vole numbers in plantations to abundance ofherbs, grasses, shrubs, and trees, species richness and diversity, andstructural diversity of total vascular plants, volume of down wood,and other site characteristics. A step-wise multiple regression anal-ysis was conducted on the six most meaningful relationships ofthe influence of site characteristics on vole abundance in planta-



Fig. 1. Photographs of linear habitats (skid-trails) (top) that have been seeded withpasture grasses (bottom) on a 5-year-old harvested site at Golden, southeasternBritish Columbia, Canada.

tions. Proportional data were arcsine-transformed prior to analysis.Means and 95% confidence intervals were calculated for the num-ber of voles in each year on the long-term sampling grids. A pairedsample t-test was used to compare the number of long-tailed volescaptured by index-lines in the 15 surveyed plantation units, and thegrid-based and index-line monthly samples of voles in grass andnon-grass sites. Duncan’s Multiple Range Test (DMRT) was used toevaluate mean values after statistically significant ANOVAs. In allanalyses, the level of significance was at least P = 0.05.

3. Results

3.1. Abundance

Vole populations were monitored on sampling grids for six years(2004–2009), since the time of harvesting, to follow how theserodents respond to successional change and reach densities capableof serious feeding damage to newly planted trees. Over 29 trappingperiods, the long-tailed vole was the most abundant microtine witha total of 625 individuals captured (96.7% of total Microtus), fol-lowed by 21 meadow voles, 113 red-backed voles, and 104 heathervoles. Susceptibility to capture was measured by Jolly trappabil-ity estimates with a mean value of 68.5% (range 66.9–70.2%) forlong-tailed voles. Populations of long-tailed voles were low in thefirst year after harvest with mean numbers <5/ha (Figs. 2 and 3a).Mean numbers in the second post-harvest year reached 15/ha andhad a strong annual cycle with up to 43 animals/ha in Septem-ber. Annual maximum densities of 49–84 voles/ha were recorded

Fig. 2. Abundance of long-tailed voles per ha on three replicate sampling grids fromthe time of harvest, 2004–2009.

in 2006, which seemed to be the peak populations on the threegrids (Fig. 2). However, in the fourth year (2007) since harvesting,numbers of long-tailed voles declined, particularly on grids D and F,while grid E remained high reaching an annual maximum of 82/ha.This decline deepened in 2008 and reached extirpation on two ofthree grids in 2009. For red-backed voles, in the first year after har-vesting, mean numbers ranged from 3.5 to 14.8/ha. However, theirnumbers declined dramatically at two years after harvesting. Theheather vole occurred at numbers ≤6/ha throughout 2004–2008and then declined to <1/ha in 2009.

3.2. Reproduction and survival

Reproductive performance of long-tailed voles, as measured bythe mean ± 1SE number of successful pregnancies, was highest in

Fig. 3. (a) Mean abundance of long-tailed voles per ha in (a) each year of the study(±95% C.I.), and (b) at 1–2, 3–4, 5–6, 7–8, and ≥9 years since clearcut harvesting, asper the 2006 survey (±1SE). Sample size is above upper bar.



Table 1Mean ± SE (n = 3 replicate sites) number of successful pregnancies, total recruits, juvenile recruits, and index of early juvenile survival for Microtus longicaudus during summerperiods, and 4-week Jolly–Seber (J–S) survival during summer and winter periods, 2004–2009 for the three sampling grids.

Parameter Year after harvesting

2004 2005 2006 2007 2008 2009

Pregnancies 2.67 ± 1.67 20.67 ± 6.69 36.00 ± 6.00 13.67 ± 4.67 9.67 ± 2.40 0.00 ± 0.00Total recruits 9.00 ± 5.00 38.00 ± 9.71 88.00 ± 16.09 47.67 ± 17.95 22.33 ± 6.74 3.33 ± 2.85Juvenile recruits 5.67 ± 3.67 27.67 ± 9.70 68.33 ± 14.24 36.67 ± 15.41 17.67 ± 4.41 3.33 ± 2.85Juvenile survival 2.06 ± 0.06 1.31 ± 0.04 1.86 ± 0.10 2.57 ± 0.27 1.88 ± 0.22 0.00 ± 0.00J–S Summer survival 0.98 ± 0.02 0.74 ± 0.07 0.72 ± 0.02 0.54 ± 0.12 0.62 ± 0.10 0.81 ± 0.11J–S Winter survival 0.85 ± 0.02 0.86 ± 0.02 0.86 ± 0.03 0.71 ± 0.06 0.70 ± 0.02 0.00 ± 0.00

2006 at 36.0 ± 6.0 (Table 1). The overall pattern was similar to thatof abundance, but with the next highest number at 20.7 ± 6.7 preg-nancies in the second rather than fourth year after harvesting asrecorded for abundance (Fig. 3a). Mean number of total recruitsand juvenile recruits followed the pattern of abundance with high-est numbers of new voles captured in 2006 (Table 1). The meanindex of early juvenile survival ranged from 1.32 to 2.57, peakingin 2007 (Table 1). Mean J–S summer survival was lowest in 2007(0.54 ± 0.12) corresponding with declining populations of voles, atleast on two of the three grids. Mean J–S winter survival also fol-lowed this pattern declining to 0.70 ± 0.02 in 2007–2008 (Table 1).

3.3. Index-lines and vole populations

The long-tailed vole was the most abundant microtine with atotal of 340 individuals captured (93.2% of total Microtus), followedby 25 meadow voles, 17 red-backed voles, and 15 heather voles onthe 57 (27 habitat characteristics plus 15 grass and 15 non-grass)sites surveyed by index-lines. The conversion (y = 0.18x + 3.68) ofindex-line numbers to per ha was based on the positive linear rela-tionship (r = 0.69; P = 0.02) (Fig. 4). These vole abundance estimateswere used in all subsequent analyses of habitat characteristics inthe 27 sites.

3.4. Voles and habitat characteristics

The relationship of mean vole abundance per ha to time sinceclearcut harvesting ranged from 17.3 at 1–2 years, and then upto 79.0 at 3–4 years, and 91.2 voles at 5–6 years, post-harvest(Fig. 3b). Vole abundance then declined to 23.9 animals/ha at 9–10years post-harvest. There was no statistical difference (F4,25 = 0.86;P = 0.50) in vole abundance among these time periods. However,numbers did increase 4.6–5.3 times from 1–2 to 3–6 years beforedeclining thereafter (Fig. 3b). The relationship of vole numbersto BEC subzone, over years 3–6 post-harvest, indicated that therewas a significant (F2,14 = 4.40; P = 0.03) difference among the threesubzones during this period. The mean (±SE) number of voles/hafor the IDFdm, MSdk, and ICHmk was 190.8 ± 67.8, 63.5 ± 7.7, and

Fig. 4. Linear regression analysis relating number of total voles on index lines tonumber on grids per ha.

107.0 ± 36.7, respectively. The IDFdm and ICHmk numbers were sim-ilar as were the MSdk and ICHmk, while the IDFdm and MSdk numberswere significantly (DMRT; P = 0.05) different.

There were few significant relationships between abundanceof voles and any one of the habitat characteristics. Four weakpositive relationships were recorded between voles and crown vol-ume index of grasses (r = 0.33; P = 0.09), volume of downed wood(r = 0.32; P = 0.11), and total species richness of all vascular plants(r = 0.38; P = 0.05). Number of large (≥20 cm diameter) pieces ofdowned wood (r = 0.53; P < 0.01) and the relationship of vole num-bers to area (ha) of site (r = 0.46; P = 0.08) also followed this pattern.There was a significant negative (−r = 0.41; P = 0.05) relationshipbetween number of voles and crown volume index of shrubs andtrees combined. There were no other meaningful relationshipsbetween habitat characteristics (amounts and diversity of vegeta-tion components) and numbers of voles: including crown volumeindex of individual species of herbs, shrubs, and trees. A multipleregression analysis of the four best-fit independent variables plustwo components of the herb layer: crown volume index and struc-tural diversity, did yield an overall significant (r = 0.67; P = 0.04)result.

3.5. Grass and non-grass habitats

There was a significant (r = 0.46; P = 0.01) positive relationshipbetween numbers of long-tailed voles and percentage cover ofgrasses in the index-line survey (n = 15) of plantation sites (Fig. 5a).Mean (±SE) cover of grasses was 61.7 ± 4.6% in the grass habi-tats and 1.8 ± 1.0% in the non-grass habitats. This pattern was alsoobserved for percentage cover of total herbs, but the trend onlyapproached significance (r = 0.33; P = 0.07) (Fig. 5b). Mean cover ofherbs was 70.0 ± 3.0% in the grass habitats and 26.4 ± 3.0% in thenon-grass habitats. The number of long-tailed voles captured byindex-lines in the 15 surveyed plantation sites was significantly(t14 = 4.05; P < 0.01) higher in the grass than non-grass habitats. Athreshold level of approximately 50% grass cover was required togenerate suitable habitat for vole numbers to reach tree damagelevels.

On grid systems, mean (±SE) numbers of long-tailed voles weresignificantly (t7 = 4.04; P < 0.01) higher (1.5–2.5 times) in the grass(mean = 23.5 ± 4.6) than non-grass (mean = 12.8 ± 2.2) habitats dur-ing 2005 and early 2006 (Table 2). Mean (±SE) cover of grasseswas 20.0 ± 1.8% in the grass habitats and 0.9 ± 0.9% in the non-grass habitats in this grid-based analysis. Mean cover of herbs was34.2 ± 11.4% in the grass habitats and 26.5 ± 9.3% in the non-grasshabitats.

Similarly, on index-lines, mean numbers of long-tailed volesalso followed this pattern of significance (t6 = 4.27; P < 0.01), being1.3–5.1 times higher in the grass (mean = 6.4 ± 1.7) than non-grass(mean = 3.3 ± 1.2) habitats during 2008 and early 2009 (Table 2).Mean cover of grasses was 80.7 ± 6.9% in the grass habitats and1.1 ± 0.6% in the non-grass habitats. Mean cover of herbs was83.3 ± 4.9% in the grass habitats and 21.2 ± 5.4% in the non-grass



Table 2Mean ± SE (n = 3 replicate sites) abundance of long-tailed voles in grass and non-grass habitats in (a) 2005–2006 (grids) and (b) in 2008–2009 (index-lines).

Habitat Monthly samples

3-May-05 3-Jun-05 27-Jun-05 27-Jul-05 21-Aug-05 16-Sep-05 8-May-06 5-Jun-06

Grass 3.67 ± 0.33 12.33 ± 1.67 19.00 ± 1.53 24.67 ± 0.88 33.00 ± 3.79 47.00 ± 1.73 21.33 ± 2.73 27.00 ± 1.15Non-grass 6.33 ± 0.88 8.00 ± 3.00 10.67 ± 2.91 12.67 ± 5.46 19.67 ± 6.39 24.67 ± 6.64 8.67 ± 1.86 11.33 ± 4.10

Habitat Monthly samples

4-Jun-08 2-Jul-08 29-Jul-08 26-Aug-08 27-Sep-08 6-May-09 3-Jun-09

Grass 1.33 ± 0.67 10.33 ± 2.03 11.67 ± 5.24 10.33 ± 4.67 7.33 ± 2.85 1.67 ± 0.88 2.00 ± 1.15Non-grass 1.00 ± 0.58 6.33 ± 3.33 8.33 ± 2.33 5.33 ± 3.18 2.00 ± 1.00 0.33 ± 0.33 0.00 ± 0.00

habitats. Thus, in both these cases (2005–2006 and 2008–2009),mean abundance of voles was maintained at a higher level in thegrass than non-grass habitats through the summer, fall, and subse-quent spring seasons.

Thus, three independent analyses showed clearly that vole num-bers were higher on those units with grass-seeded sites, whetherthey were along skid-trails, roadsides, or miscellaneous seedings.

3.6. Vole abundance and tree mortality

Most cutover forest sites in the interior of B.C. are planted withtree seedlings at a density of 1400–1600 per ha. The incidence ofmortality of trees from feeding damage by voles in our generalGolden study area has ranged from 15% to 100%. Sites have beenre-planted (in some situations several times) whenever tree loss isunacceptably high (e.g., <700 surviving trees/ha). Presumably theincidence of damage is related to the abundance of voles. Therewas a significant positive (r = 0.57; P = 0.01) relationship of percent-age tree mortality and abundance of voles (Microtus) (Fig. 6). Thus,the number of voles on a given site can be related to the poten-tial for feeding damage to trees in that particular plantation. It is

Fig. 5. Linear regressions of the relationship of percentage cover of (a) seeded pas-ture grasses and (b) herbs, to number of long-tailed voles in sampled plantationsites.

important to note that in some cases there can be relatively highnumbers of voles (in the moderate category), but little damage totree seedlings. Conversely, a relatively low number of voles may, incertain situations, damage a high percentage of trees. Based on thisrelationship, a risk rating for damage to trees would be, in termsof number of voles/ha: Low <7; moderate 7–34; high 35–88; veryhigh >88 (Fig. 6).

4. Discussion

4.1. Population changes of long-tailed voles

Our study is the first relatively long-term analysis of changesin population dynamics of the long-tailed vole. Mean numberson the grids were highest in 2006 and 2007, 3–4 years post-harvest, at 50–80 voles/ha. The index-line samples of maximummean numbers of 79 and 91 voles/ha in sites 3–4 years and 5–6years post-harvest, respectively, corroborated the grid measure-ments. This range was higher than the 30–50 voles/ha reportedfor three successional stages (2, 7, and 23 years post-harvest) incoastal forest of southeast Alaska (Van Horne, 1982), and 13–18/hain the third post-harvest year in montane spruce forest (Sullivanand Sullivan, 2001) in south-central B.C. Sampling of long-tailedvole populations continued on these same sites for up to 8 yearspost-harvest, but did not reach >2 voles/ha after the third post-harvest year (Sullivan et al., 2008). Klenner and Sullivan (2003)reported a maximum of 15–22 long-tailed voles/ha on clearcutsin the fourth post-harvest year in ESSF forest in south-central B.C.However, Sullivan et al. (1999) reported a strong annual cycle inclearcut sites 9–13 years post-harvest with maximum numbers

Fig. 6. Linear regression relationship of percentage tree mortality to abundance ofvoles. The two datapoints (empty circles) were outliers and not part of the analysis.



of 50–70 voles/ha in ESSF forest in northwestern B.C. In naturalhabitats, the range in maximum abundance of long-tailed voleswas 30–120/ha for two years in high elevation meadows in NewMexico (Conley, 1976), 16–29/ha for three years in high eleva-tion meadows in Utah (Cranford, 1984), and 8–12 voles/ha overtwo years in shrub-grassland of southeastern Washington (Farris,1971).

Harris (1968) also found high numbers of long-tailed voles 2–3years after logging in southeast Alaska, with low numbers in sub-sequent years. Similar results were reported for other Microtusspecies in the early years after clearcut harvesting (Gashwiler,1970; Martell and Radvanyi, 1977). Later successional stages hadfewer voles, owing to reduced understory vegetation and thick for-est canopies (Van Horne, 1982; and others). These studies all seemto suggest an annual cycle in population fluctuations of long-tailedvoles that is tied to changes in early successional vegetation, atleast in post-harvest sites. The persistence of long-tailed voles onolder logged sites in southeast Alaska and northwestern B.C., com-pared with our study, may be related to higher plant productivity,or alternatively, delayed vegetative succession, in these northernecosystems.

The over-arching concept of multi-annual cycling of Microtuspopulations sometimes dominates the interpretation of datasets.Our high populations in 2006 may have seemed like a “peak”,and 2009 a “low” in such a population cycle. However, this pat-tern appeared not to be the case, based on the high abundance(mean of 134 voles/ha) in the index-line sampling of 3-year-old har-vested sites in 2009. If long-tailed voles had declined in 2008–2009as part of a multi-annual cycle, this pattern would presumablyhave occurred throughout the general study area. However, as dis-cussed by Taitt and Krebs (1985), some Microtus species exhibitboth annual and multi-annual cycles. The amplitude for numericalchange is typically less than 5-fold for annual fluctuations and usu-ally over 10-fold for multi-annual cycles (Taitt and Krebs, 1985).Our grid populations had a mean maximum density of 27.9 andmean minimum density of 6.6 voles/ha, thereby yielding a 4.2-folddifference in amplitude across the six years. In addition, long-tailed vole populations declined in abundance each spring whichis another feature that is typical of annual fluctuations in abun-dance (Taitt and Krebs, 1985). Patterns in reproduction and survivalgenerally followed those of abundance throughout the six years ofpopulation changes.

4.2. Voles and habitat

Why did long-tailed voles reach such high densities in our studyareas? Voles reached their highest abundance in early successionalhabitats after clearcutting and this seemed to be associated withrelatively high crown volume index of grasses, as well as speciesrichness and diversity of vascular plants. Richness and diversitycould be considered a measure of vegetation complexity. In addi-tion, the influx of pasture grasses clearly led to higher numbersof long-tailed voles on seeded sites. While we had a weak posi-tive relationship with volume of downed wood, and a significantone with number of large pieces, Van Horne (1982) and Craig(2002) both reported significant relationships of this microtinewith amount of downed wood. Similarly, Sullivan et al. (1999)reported few long-tailed voles on clearcut sites that had beenbroadcast burned, with the consequent reduction in amount ofdowned wood and arrested vegetative succession. Therefore, long-tailed voles, as well as other Microtus, seem to require a giventhreshold of cover (both thermal and security) to occupy a habi-tat and increase in abundance (Van Horne, 1982; Getz, 1985). Forexample, a threshold level of 50% cover was required in linear habi-tats composed of pasture grasses to generate suitable habitat forvole numbers to reach tree damage levels. Adler and Wilson (1989)

reported abundance and survival of meadow voles increased lin-early up to a critical level of grass cover of 30–40%. Thus, biomassof grasses (including pasture grasses), vegetation complexity, andamount of downed wood may all have combined to provide opti-mum habitat for vole population buildups and enhanced survival,particularly during overwinter periods (Birney et al., 1976; Getz,1985; Bergeron and Jodoin, 1989; Adler, 1987).

Downed wood is an important moisture reservoir and may pro-vide a sheltered, cooler microclimate attractive to voles in summer(Getz, 1985). In winter, along with vegetation (particularly grasses),it may provide a mechanical support creating a snow-free space atthe ground surface (Spencer, 1984). Clearly, the substantial levels ofpost-harvest woody debris on our clearcut sites provided the struc-tural aspect for establishment of a subnivean space for long-tailedvoles. This space is crucial to the winter distribution, activity, andsurvival of small mammals, and potentially to feeding on conifer-ous trees (Spencer, 1984). It is this cover of vegetation and downedwood that predisposes seedlings, planted in these sites, to feedingby voles (Ostfeld and Canham, 1993).

Did diet play a role in long-tailed vole abundance patterns inthese young plantations (clearcuts)? Reports indicated a diet pri-marily of dicots, fruits, and seeds in Alaska (Van Horne, 1982),and grass and sedge components from Festuca and Carex, respec-tively, as well as tree bark and cambium during winter in Colorado(Spencer, 1984). Sullivan and Sullivan (2001) reported a positiverelationship between abundance of Microtus and number of lodge-pole pine cones and their potential source of seeds on clearcut sites.Long-tailed voles readily ate large numbers of pine seeds whenavailable (Sullivan and Sullivan, unpublished). We did not mea-sure abundance of cones on our sites, but the principal tree speciesharvested in these primarily MPB-attacked and MPB-susceptibleforests was lodgepole pine with its serotinous cones opening torelease seeds in the post-logging slash (Lotan, 1975). Vole popula-tions may have declined when the pine seed supply was exhausted.Again, although we have no direct evidence, the pasture grass mixcontained several species of grasses as well as alfalfa and clover.These latter two plant species have high nutritional value and mayhave contributed to enhancing reproduction, growth, and survivalof long-tailed voles (Batzli, 1985). Otherwise, our lack of any sig-nificant relationships of long-tailed vole abundance with crownvolume index of individual plant species was not helpful in thisregard.

Another possible explanation for the relatively high numbers oflong-tailed voles is the lack of other Microtus species as competitors.Long-tailed voles are reportedly timid and avoid encounters withother species of Microtus (Colvin, 1973; Randall, 1978). Sullivan andSullivan (2001) recorded a pattern of potential competitive dis-placement of M. longicaudus by M. pennsylvanicus. This behavioralinteraction has been suggested for the decline of red-backed volesafter harvesting, as Microtus, particularly the meadow vole, cometo dominate the microtine community (Iverson and Turner, 1972;Morris and Grant, 1972).

High numbers of long-tailed voles may also have been relatedto the absence of cattle grazing on our early successional sites.Cattle on summer range are ubiquitous in virtually all accessiblesites in the southern interior of B.C. and some other parts of thePacific Northwest. As reviewed by Johnston and Anthony (2008),Microtus have responded negatively to grazing as their preferredherbaceous and grass-dominated habitats have been altered. Thus,it is not surprising that long-tailed and meadow voles have not com-monly reached densities >20 animals/ha on recent clearcuts grazedby livestock. Our study areas near Golden, and those of Van Horne(1982) in southeast Alaska, did not have any grazing pressure fromlivestock. The major effect on understory vegetation was driven bysuccession as shrubs and trees came to dominate the habitat, withthe concurrent decline in grasses and forbs.



Thus, the prediction of H1 that populations of long-tailed voleswould rise and fall with the abundance of herbaceous plantsseemed to be supported. Similarly, these populations respondedpositively to the seeded pasture grasses as per the prediction ofH2. Abundance of native grasses (e.g., pinegrass (Calamogrostisrubescens)) appeared to decline with succession as shrubs andtrees came to dominate at ≥9 years after harvest. However, thepasture grasses in the various linear habitats did not decline, butrather expanded in area (see Fig. 1). Presumably the relatively smallarea (range of 0–23% coverage of sites) of these seeded grassesdid not provide enough essential habitats as shrubs increasedin abundance with time. The weak positive relationship of volenumbers with area of plantation sites tended to support this obser-vation.

4.3. Voles, plantations, and tree damage

The positive relationship of the incidence of overwinter damageto trees and vole abundance in the previous autumn, in three geo-graphic areas of B.C., is the first such analysis for forest plantationson harvested sites in North America. Ostfeld and Canham (1993)and Ostfeld et al. (1997) reported a similar relationship betweenmeadow vole density and seedling predation in old fields. Hansson(1986) related vole abundance to degree of vole de-barking oftrees at a small local scale in Sweden. Huitu et al. (2009) pro-vided a density-dependent vole damage analysis, based on surveyquestionnaires, at a nationwide scale in Finland. These resultsstrongly support population monitoring of voles as an effectivemeans to forecast future outbreaks in damage to new planta-tions.

There was considerable variation in our density-dependent voledamage relationship, as was also reported for the Finnish study(Huitu et al., 2009). In some cases, there was high (>80% mortality)incidence of damage, but few voles (<15/ha) recorded in that plant-ing. Alternatively, there were few trees eaten (<10%) at a very highabundance of voles (>70/ha) in another experimental plantation(see Fig. 6). However, the prediction of H3 that damage incidenceand vole abundance would be positively related seemed to be sup-ported. Our regression relationships were based on vole numbersand various habitat characteristics. The implicit understanding wasthat vole numbers and incidence of tree damage were highly cor-related, which was supported by a reasonably strong relationship.The Golden study area had a history of vole damage and all siteshad been replanted, some a multitude of times, and hence it wasnot possible to use incidence of tree damage as a dependent vari-able. Most plantations had several cohorts of trees from successiveplanting events. Newly planted seedlings are primarily damaged inthe first winter when the fertilization regime renders them partic-ularly palatable to voles (Sullivan and Martin, 1991; Sullivan andSullivan, 2008). This immediacy of vole predation on seedlings wasalso recorded in old fields by Ostfeld and Canham (1993). Dam-age may still occur in subsequent winters, but tends to be minorby 2–3 years post-planting. This pattern is likely related to volepopulations starting to decline by 4–5 years after harvest. Feedingdamage to trees in older plantations tends to be in “hotspots” wherea few long-tailed voles reside.

There was considerable variation in the relationship of voleabundance (and hence tree damage) to habitat characteristics. Themultiple regression of six factors explained 45% of this variation,but other site specific factors such as moisture (may be relatedto aspect), proximity to source populations of long-tailed voles,and incidence of predators could also be important. Most Microtusspecies respond favorably to moisture-bearing sites with enhancedgrowth of herbaceous plants (Getz, 1985). Long-tailed voles werecaptured mainly on seepage sites in north-aspect burned unitsin Montana (Halvorson, 1982). Similar results were recorded for

creeping voles (M. oregoni) in Oregon (Gashwiler, 1970; Hooven,1973).

Source populations of long-tailed voles pose an interesting sce-nario. This microtine was recorded at low abundance (<10/ha)in closed canopy forests (Van Horne, 1982; Klenner and Sullivan,2003). Thus, it seemed unlikely that older uncut forests were sourceareas; rather openings and natural meadows supplied sufficientearly successional forbs and grasses (Smolen and Keller, 1987).Contiguous units of clearcut harvesting over relatively short peri-ods provided several hundred ha of early successional habitat forlong-tailed voles at the Golden study area. This rapid sequence ofharvesting was typical of salvage operations for MPB-susceptiblelodgepole pine dominated stands and has occurred in many partsof the PNW over the last decade. Thus, long-tailed voles presum-ably move from one harvested site to another as new grass and forbcommunities develop. They may be assisted in this migration byroad corridors with banks and edges seeded with pasture grasses.This practice occurs in many new road and cutblock installationsin B.C. and perhaps other parts of the PNW as well. It has beendiscontinued in some nature reserves and National Parks becauseof the migration of alien flora (Tyser and Worley, 1992). Althoughwe do not have any data on vole movements, it seemed likely thatlong-tailed voles would disperse along these linear, and potentiallyoptimum, habitats since these microtines were so abundant in thegrass index-lines. Moving to optimum habitats that maximize theirfitness has been reported for other vole species (Lin and Batzli,2004).

4.4. Study limitations

Our study design relied on independent sampling sites asper the recommendations of Hurlbert (1984) for experimentalunits. Sites were deemed independent since no long-tailed voleswere captured on more than one grid or index-line. The sizesof experimental units (harvested sites) were those of typicalforestry operations. There was a wide range of sizes of cut-blocks and successional ages since harvesting. We sampled onlyclearcut sites and utilized all available units in the Glenogle andRoth Creek drainages. Inferences from index-line samples of voleabundance patterns in different-aged sites and vegetative succes-sion should be interpreted with caution as these surveys wereessentially “snapshots in time”. In addition, we did not sam-ple undisturbed habitats such as perennial grasslands or matureforests where long-tailed vole populations might have been presentand changing in abundance independently of vegetative succes-sion.

Although we did detect a significant difference in abundance ofvoles among BEC zones, additional sites in the IDFdm and ESSFdksubzones would have strengthened our inference. In general, thelower elevation (IDF and ICH) zones tend to be more productivethan their higher elevation (MS and ESSF) counterparts (Meidingerand Pojar, 1991). Sampling of vegetation along index-lines rarelyencountered the linear habitats of seeded pasture grasses, proba-bly because they covered a relatively low proportion of a given site.Additional vegetation sampling that included grass strips mighthave proven useful as part of the analysis of habitat character-istics. The separate grass versus non-grass analysis did serve toinclude this important aspect, but was limited to percentage groundcover, rather than crown volume index, for the vegetation samplingregime.

It would have been ideal for continued sampling of long-tailedvoles on these sites for several additional years to determineif a multi-annual population fluctuation might have occurred.Based on our sampling of habitat characteristics, long-tailed voleswere uncommon on older successional sites as they were insouth central B.C. (Sullivan et al., 2008), but they did occur in



Table 3Factors contributing to vole population outbreaks and feeding damage to plantations in south-central British Columbia, Canada.

Alaska (Van Horne, 1982) and northwestern B.C. (Sullivan et al.,1999).

5. Conclusions

A list of factors contributing to vole population outbreaks andfeeding damage to plantations is summarized in Table 3. Time sinceclearcut harvesting at 3–4 years is a critical time for populationbuildups of voles and subsequent damage to plantation trees. Com-parison of vole responses to clearcutting and variable retentionsystems may help clarify the role of harvesting method, where thisis a flexible operational scenario. Larger patch sizes (area of clearcutsite) tend to have a higher abundance of voles. Large contiguousopenings, typical of MPB salvage harvesting, provide substantialhabitat. Clearcut sites in the IDFdm and ICHmk subzones appearedto be most susceptible to vole abundance and consequent damage,although it must be noted that the IDFdm sites also had a high degreeof area seeded with pasture grasses. Seeded grass species clearlycreate optimum habitat conditions for voles, generating popula-tion densities up to 30–50 voles/ha, which is in the range of a “high”damage risk to seedlings. Risk ratings (voles/ha) for feeding dam-age to trees were low (<7), moderate (7–34), high (35–88), andvery high (>88). However, there was considerable variability amongplantation sites, incidence of damage, and abundance of voles.

Acknowledgements

We thank the Forest Science Program (British Columbia Ministryof Forests and Range) and Louisiana-Pacific Canada Ltd. for finan-cial and logistical support for this project. S. King was particularlyhelpful in assisting with this project. We thank J.H. Sullivan for helpwith the field work.

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