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BARK BEETLE DISTURBANCE AND
NITROGEN CYCLING IN CONIFER FORESTS
OF GREATER YELLOWSTONE
by
Jacob M. Griffin
A dissertation submitted in partial fulfillment
of the requirements for the degree of
Doctor of Philosophy
(Zoology)
at the
UNIVERSITY OF WISCONSIN-MADISON
2011
i
TABLE OF CONTENTS
PREFACE……….………………………………………………………………………………iii
CHAPTER 1
Nitrogen cycling following mountain pine beetle disturbance in lodgepole pine
forests of Greater Yellowstone…………………………...………………………………...……1
Abstract…………………………………………………………………………………...1
Introduction………………………………………………………………………………2
Methods…………………………………………………………………………………..5
Results…………………………………………………………………………………...11
Discussion……………………………………………………………………………….16
Conclusions……………………………………………………………………………...23
Acknowledgements……………………………………………………………………..23
References……………………………………………………………………………….24
Tables……………………………………………………………………………………31
Figure legends…………………………………………………………………………..35
Figures…………………………………………………………………………………..37
CHAPTER 2
Forest type influences ecosystem response to bark beetle disturbance……………………..43
Abstract………………………………………………………………………………….43
Introduction……………………………………………………………………………..44
Methods………………………………………………………………………………….47
Results…………………………………………………………………………………...52
Discussion……………………………………………………………………………….58
ii
Conclusions.......................................................................................................................61
Acknowledgements..........................................................................................................61
References.........................................................................................................................61
Tables................................................................................................................................66
Figure legends...................................................................................................................69
Figures...............................................................................................................................71
CHAPTER 3
Does salvage logging alter ecosystem response to bark beetle disturbance
in lodgepole pine forests?............................................................................................................76
Abstract.............................................................................................................................76
Introduction......................................................................................................................77
Methods.............................................................................................................................80
Results...............................................................................................................................84
Discussion.........................................................................................................................88
Conclusions.......................................................................................................................92
Acknowledgements..........................................................................................................93
References.........................................................................................................................93
Tables................................................................................................................................99
Figure legends.................................................................................................................100
Figures.............................................................................................................................102
CONCLUSIONS........................................................................................................................110
iii
PREFACE
Disturbance has long been considered an important driver of ecosystem dynamics
(Bormann and Likens 1979, Pickett and White 1985, Chapin et al. 2002), influencing a wide
range of key ecosystem functions such as carbon storage (Li et al. 2003, Kashian et al. 2006),
nutrient retention (Pardo et al. 1995), and biodiversity (Connell 1978). The influence of
disturbance on ecosystem function may be manifested through several key mechanisms,
including the redistribution of biomass, nutrients, and energy among ecosystem pools, the
alteration of spatial pattern (Hadley 1994, Turner et al. 1999, Turner et al. 2004), and changes in
species composition (Chapin et al. 2002, Ellison et al. 2005). Effects of disturbance on
ecosystem function have been quantified for a wide range of spatial scales from small gaps
(Scharenbroch and Bockheim 2008) to whole continents (e.g.Potter et al. 2005), and often
exhibit threshold spatial extents or severities where qualitative shifts in ecosystem response
occur (Parsons et al. 1994). Temporal scale is also a key consideration, as many disturbance
types have either long-lived or delayed effects on ecosystem function (Foster et al. 1998,
Compton and Boone 2000, DeLuca and Aplet 2008). Disturbance interactions, often crossing
multiple scales, will be key to predicting future ecosystem consequences and have been
demonstrated in several systems (Paine et al. 1998, Throop and Lerdau 2004, Allen 2007, Raffa
et al. 2008).
Insect activity and insect-induced tree mortality can be significant components of the
disturbance regime in many temperate forests with substantial consequences for elemental
nutrient cycling. Mechanisms of insect disturbance involve both direct influence of herbivory
and insect-induced mortality on biomass stocks, growth rates, and ecosystem fluxes such as
litterfall and decomposition, as well as indirect effects through altered abiotic conditions or shifts
iv
in plant community composition (Mattson and Addy 1975, Schowalter 1985, Schowalter et al.
1986, Haack and Byler 1993, Hunter 2001). The great majority of studies regarding insect
disturbance and nutrient cycling have focused on folivorous insects as the study system (e.g.
Lovett and Ruesink 1995, Eshleman et al. 1998, Kosola et al. 2001, Christenson et al. 2002,
Townsend et al. 2004). Significantly less attention has been given to potential nutrient cycling
changes following outbreaks of non-folivorous insects, with notable exceptions (Huber et al.
2004, Huber 2005, Morehouse et al. 2008). By targeting specific tissues of host plants, various
insect guilds can have widely differing effects on host mortality and the transfer of materials
among ecosystem pools (Feller 2002, Lovett et al. 2006)
Even less studied is the role of forest type in determining ecosystem response to
outbreaks of the same insect guild. This knowledge gap is particularly important when
considering foundation species, whose characteristics largely determine ecosystem function
(Ellison et al. 2005) and whose dominance in forests often increases susceptibility to insect
attack (Safranyik and Carroll 2006). Variability in the baseline biogeochemistry of pure stands of
different tree species (Prescott et al. 1992, Thomas and Prescott 2000, Giardina et al. 2001,
Lovett et al. 2004) may lead to unique ecosystem responses following the same disturbance type.
Furthermore, additional compound disturbances following insect attack such as fire or salvage
logging may lead to qualitatively different effects on nutrient cycling compared to insects alone
(Paine et al. 1998, Peters et al. 2004). This is particularly important if insect outbreak affects the
probability or severity of other subsequent disturbances (Lindenmayer and Noss 2006, Page and
Jenkins 2007, Simard et al. 2011). Disturbance types vary in the specific ecosystem components
that are directly affected. For example, direct effects of bark beetles are limited to canopy
vegetation, and can be contrasted with other disturbance types such as forest fire, windstorm, or
v
logging which cause immediate change in a wider range of ecosystem components including
understory vegetation and soils (Figure 1).
Throughout western North America, several species of native bark beetles of the genus
Dendroctonus (Coleoptera: Curculionidae: Scolytinae) have been in outbreak phase over the past
decade (Raffa et al. 2008). The impacts of this disturbance type on some aspects of ecosystem
function such as carbon sequestration (Kurz et al. 2008) and hydrology (Boon 2009) are
beginning to be understood, but the consequences for nitrogen (N) cycling are less well-known.
In the Greater Yellowstone Ecosystem of WY, MT, and ID, USA, current bark beetle outbreaks
have occurred in multiple conifer forest types including lodgepole pine (Pinus contorta Dougl.),
Douglas-fir (Pseudotsuga menziesii [Mirb.] Franco), Englemann spruce (Picea engelmannii
Parry), and whitebark pine (Pinus albicaulis Engelmann). Outbreaks are allowed to progress
naturally within the National Park lands of the region, however salvage logging is being applied
as a management technique on National Forest lands to recover usable forest products and reduce
a perceived risk of increased forest fire.
Bark Beetles
Salvage logging
Bark Beetles
Salvage loggingSalvage logging
Figure 1. Impacts of bark beetle outbreak and other disturbance types on forest ecosystem components. Beetles remove large portions of the canopy, but direct impacts are limited to this component. Other disturbances including salvage logging cause additional canopy disturbance, but also may have direct impacts understory vegetation and soil. Adapted from Roberts (2004).
vi
For my dissertation I will address three general questions related to the issues described
above. (1) How do forest structure, soil microclimate, and N cycling through the litter, soil,
and vegetation change during and decades after beetle outbreak?, (2) How does host tree
forest type moderate the response of these ecosystem parameters to similar bark beetle
outbreaks?, and (3) How do forest structure, microclimate, and nitrogen cycling following
post-beetle salvage logging differ from bark beetle outbreak alone? I will approach these
questions by taking advantage of concurrent outbreaks of two Dendroctonus bark beetle species
in the Greater Yellowstone Ecosystem. First, I will use a replicated chronosequence approach to
measure the effects of mountain pine beetle (Dendroctonus ponderosae Hopkins) disturbance
over a thirty year time span in lodgepole pine forests. Second, I will explore the influence of
forest type on ecosystem response to bark beetles by comparing changes in the mountain pine
beetle/lodgepole pine system to those resulting from Douglas-fir beetle (Dendroctonus
pseudotsugae Hopkins) outbreak in Douglas-fir forests over a shorter 5 year time scale. Thirdly,
I will use a paired and replicated before-after-control-impact (BACI) experimental design in
lodgepole pine forests to compare the combined effects of mountain pine beetle outbreak and
salvage logging to the effects of bark beetles alone (Figure 2).
Figure 2. Three dissertation chapters on the influence of bark beetle disturbance on N cycling in conifer forests using multiple approaches. Chapter 1 describes a single beetle/host system over decadal scales. Chapter two compares two beetle/host systems over shorter time scales. Chapter three is an experiment testing short-term effects of the compound disturbances of beetles and salvage logging.
Chapter 1
Time scale of study
Num
ber
of sy
stem
s
Years Decades
Tw
oO
ne
Chapter 2
Influence of host forest type on ecosystem response
Chapter 1
Decadal-scale impact of bark beetles on the N cycle
Chapter 3
Compound disturbances
Bark beetles + salvage logging
Chapter 1Chapter 1Chapter 1
Time scale of study
Num
ber
of sy
stem
s
Years Decades
Tw
oO
ne
Time scale of study
Num
ber
of sy
stem
s
Years Decades
Tw
oO
ne
Chapter 2
Influence of host forest type on ecosystem response
Chapter 1
Decadal-scale impact of bark beetles on the N cycle
Chapter 3
Compound disturbances
Bark beetles + salvage logging
vii
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Lovett, G. M., C. D. Canham, M. A. Arthur, K. C. Weathers, and R. D. Fitzhugh. 2006. Forest ecosystem responses to exotic pests and pathogens in eastern North America. Bioscience 56:395-405. Lovett, G. M., and A. E. Ruesink. 1995. Carbon and Nitrogen Mineralization from Decomposing Gypsy-Moth Frass. Oecologia 104:133-138. Lovett, G. M., K. C. Weathers, M. A. Arthur, and J. C. Schultz. 2004. Nitrogen cycling in a northern hardwood forest: Do species matter? Biogeochemistry 67:289-308. Mattson, W. J., and N. D. Addy. 1975. Phytophagous Insects as Regulators of Forest Primary Production. Science 190:515-522. Morehouse, K., T. Johns, J. Kaye, and A. Kaye. 2008. Carbon and nitrogen cycling immediately following bark beetle outbreaks in southwestern ponderosa pine forests. Forest Ecology and Management 255:2698-2708. Page, W. G., and M. J. Jenkins. 2007. Mountain pine beetle-induced changes to selected lodgepole pine fuel complexes within the intermountain region. For Sci 53:507-518. Paine, R. T., M. J. Tegner, and E. A. Johnson. 1998. Compounded perturbations yield ecological surprises. Ecosystems 1:535-545. Pardo, L. H., C. T. Driscoll, and G. E. Likens. 1995. Patterns of nitrate loss from a chronosequence of clear-cut watersheds. Water Air and Soil Pollution 85:1659-1664. Parsons, W. F. J., D. H. Knight, and S. L. Miller. 1994. Root Gap Dynamics in Lodgepole Pine Forest : Nitrogen Transformations in Gaps of Different Size. Ecol App 4:354-362. Peters, D. P. C., R. A. Pielke, B. T. Bestelmeyer, C. D. Allen, S. Munson-McGee, and K. M. Havstad. 2004. Cross-scale interactions, nonlinearities, and forecasting catastrophic events. Proceedings of the National Academy of Sciences of the United States of America 101:15130-15135. Pickett, S. T. A., and P. S. White, editors. 1985. The ecology of natural disturbance and patch dynamics. Academic Press, Orlando, FL. Potter, C., P. N. Tan, V. Kumar, C. Kucharik, S. Klooster, V. Genovese, W. Cohen, and S. Healey. 2005. Recent history of large-scale ecosystem disturbances in North America derived from the AVHRR satellite record. Ecosystems 8:808-824. Prescott, C. E., J. P. Corbin, and D. Parkinson. 1992. Availability of Nitrogen and Phosphorus in the Forest Floors of Rocky-Mountain Coniferous Forests. Canadian Journal of Forest Research-Revue Canadienne De Recherche Forestiere 22:593-600.
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Raffa, K. F., B. H. Aukema, B. J. Bentz, A. L. Carroll, J. A. Hicke, M. G. Turner, and W. H. Romme. 2008. Cross-scale drivers of natural disturbances prone to anthropogenic amplification: The dynamics of bark beetle eruptions. Bioscience 58:501-517. Roberts, M. R. 2004. Response of the herbaceous layer to natural disturbance in North American forests. Canadian Journal of Botany-Revue Canadienne De Botanique 82:1273-1283. Safranyik, L., and A. L. Carroll. 2006. The biology and epidemiology of the mountain pine beetle in lodgepole pine forests. Pages 3-66 in L. Safranyik and B. Wilson, editors. The mountain pine beetle: A synthesis of biology, management, and impacts on lodgepole pine. Nat. Res. Can., Can. Forest Service, Pacific Forestry Centre, Victoria, BC. Scharenbroch, B. C., and J. G. Bockheim. 2008. The effects of gap disturbance on nitrogen cycling and retention in late-successional northern hardwood-hemlock forests. Biogeochemistry 87:231-245. Schowalter, T. D. 1985. Adaptations of Insects to Disturbance. Pages 235-252 in S. T. A. Pickett and P. S. White, editors. The Ecology of Natural Disturbance and Patch Dynamics. Academic Press, Orlando. Schowalter, T. D., W. W. Hargrove, and D. A. Crossley. 1986. Herbivory in Forested Ecosystems. Annual Review of Entomology 31:177-196. Simard, M., W. H. Romme, J. M. Griffin, and M. G. Turner. 2011. Do mountain pine beetle outbreaks change the probability of active crown fire in lodgepole pine forests? Ecological Monographs 81:3-24. Thomas, K. D., and C. E. Prescott. 2000. Nitrogen availability in forest floors of three tree species on the same site: the role of litter quality. Canadian Journal of Forest Research- Revue Canadienne De Recherche Forestiere 30:1698-1706. Throop, H. L., and M. T. Lerdau. 2004. Effects of nitrogen deposition on insect herbivory: Implications for community and ecosystem processes. Ecosystems 7:109-133. Townsend, P. A., K. N. Eshleman, and C. Welcker. 2004. Remote sensing of gypsy moth defoliation to assess variations in stream nitrogen concentrations. Ecological Applications 14:504-516. Turner, M. G., W. H. Romme, and R. H. Gardner. 1999. Prefire heterogeneity, fire severity, and early postfire plant reestablishment in subalpine forests of Yellowstone National Park, Wyoming. International Journal of Wildland Fire 9:21-36. Turner, M. G., D. B. Tinker, W. H. Romme, D. M. Kashian, and C. M. Litton. 2004. Landscape patterns of sapling density, leaf area, and aboveground net primary production in postfire lodgepole pine forests, Yellowstone National Park (USA). Ecosystems 7:751-775.
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CHAPTER 1
Nitrogen cycling following mountain pine beetle disturbance in
lodgepole pine forests of Greater Yellowstone
ABSTRACT
Widespread bark beetle outbreaks are currently affecting multiple conifer forest types
throughout western North America, yet many ecosystem-level consequences of this disturbance
are poorly understood. We quantified the effect of mountain pine beetle (Dendroctonus
ponderosae) outbreak on nitrogen (N) cycling through litter, soil, and vegetation in lodgepole
pine (Pinus contorta var. latifolia) forests of the Greater Yellowstone Ecosystem (Wyoming,
USA) across a 0-30 year chronosequence of time-since-beetle disturbance. Recent (1-4 years)
bark beetle disturbance increased total litter depth and N concentration in needle litter relative to
undisturbed stands, and soils in recently disturbed stands were cooler with greater rates of net N
mineralization and nitrification than undisturbed sites. Thirty years after beetle outbreak, needle
litter N concentration remained elevated; however total litter N concentration, total litter mass,
and soil N pools and fluxes were not different from undisturbed stands. Canopy N pool size
declined 58% in recent outbreaks, and remained 48% lower than undisturbed in 30-year old
outbreaks. Foliar N concentrations in unattacked lodgepole pine trees and an understory sedge
were positively correlated with net N mineralization in soils across the chronosequence. Bark
beetle disturbance altered N cycling through the litter, soil, and vegetation of lodgepole pine
forests, but changes in soil N cycling were less severe than those observed following stand
replacing fire. Several lines of evidence suggest the potential for N leaching is low following
bark beetle disturbance in lodgepole pine.
2
INTRODUCTION
Bark beetle disturbance and nitrogen cycling
Disturbance regulates a wide range of ecosystem properties (Bormann and Likens 1979,
Pickett and White 1985, Chapin et al. 2002), and in forests, disturbance-induced tree mortality
can change stand structure, productivity, and nutrient cycling (Lodge et al. 1994, Ostertag et al.
2003). Because canopy foliage stores nutrients, determines litter quantity and quality, and
moderates litter-soil microclimate (Prescott 2002), canopy disturbance may influence soil
nutrient availability (Chapin et al. 2002). Previous studies of disturbance and nutrient cycling in
forest ecosystems have provided insight into the magnitude, timing and mechanisms of response
(e.g.Vitousek et al. 1979). However, disturbances vary spatially and temporally in both pattern
and severity, and ecosystem response varies among disturbance types and forest communities.
Insect outbreaks are common disturbances in forest ecosystems, though most studies of
insect herbivores and nutrient cycling have addressed Lepidopteran (e.g. Lovett and Ruesink
1995, Kosola et al. 2001, Christenson et al. 2002, Houle et al. 2009) and Homopteran (e.g.
Kizlinski et al. 2002, Stadler et al. 2006) insects. Few studies have focused on stem phloem-
feeding insects, such as Dendroctonus and Ips bark beetle species (Coleoptera: Curculionidae-
Scolytinae). In contrast to folivores, bark beetles kill trees by consuming cambium and
disrupting phloem flow, thereby acting as agents of selective mortality within a forest. Because
bark beetle outbreaks have reached unprecedented levels throughout western North America
(Raffa et al. 2008), understanding the ecological effects of widespread beetle-induced tree
mortality has become increasingly important.
Bark beetles are native to temperate and boreal coniferous forests and intermittent
outbreaks typically recur at decadal-scale intervals (Raffa et al. 2008). Successful beetle attack in
3
summer causes tree death by the following spring, when needles turn red and begin to fall; the
stand progresses from this “red stage” to the “gray stage” as all needles are shed (Safranyik and
Carroll 2006). Large trees are most susceptible to bark beetle attack, and mortality of large trees
can reach 100% (Safranyik and Carroll 2006). When the outbreak subsides, release of subcanopy
surviving trees is often the major mechanism of regeneration (Romme et al. 1986, Nigh et al.
2008, Vyse et al. 2009). Herbaceous species and shrubs also increase presumably in response to
light and water availability (McCambridge et al. 1982, Stone and Wolfe 1996), but increased
nitrogen availability may play a role. The effects of bark beetles on stand structure (tree density,
basal area, and species composition) have been well documented (Shore et al. 2006, Dordel et al.
2008, Klutsch et al. 2009). However, the consequences of bark beetle outbreaks for nitrogen (N)
cycling have not been widely studied.
The effect of bark beetle outbreak on N pools and fluxes depends on the balance between
factors that enhance or limit nutrient supply. Stand-level transpiration declines as trees within the
stand begin to die, and soil moisture increases. Nutrient uptake by trees also declines, but soil
nutrient pools will increase only if the rate of nutrient supply via mineralization exceeds the rate
of nutrient removal via uptake and leaching. A post-outbreak pulse of litter could increase soil
inorganic N pools (Cullings et al. 2003), but conifer litter can also immobilize soil N (Fahey et
al. 1985). Canopy and litter changes also affect incident radiation (Hais and Kucera 2008), air
flow (Boon 2009) and soil insulation (Byers 1984), and experimental additions of lodgepole pine
litter have been shown to modify litter-soil microclimate and soil N availability (Cullings et al
2003). In an N-saturated Norway spruce (Picea abies (L.) Karst.) forest in Bavaria, Germany,
soil nitrate concentrations were elevated for 7 years after an outbreak of Ips typographus, but
returned to pre-outbreak levels after 17 years (Huber 2005). In soils of southwestern US Pinus
4
ponderosa forests, disturbance by Ips and Dendroctonus bark beetle species increased soil
ammonium and laboratory net nitrification rates, but did not affect soil nitrate or laboratory net
mineralization rates (Morehouse et al. 2008). Additional studies that compare a broader range of
N pools and fluxes, both during and after an outbreak, with unaffected stands are needed to
understand how bark beetle disturbance alters N flow through the litter, soil, and vegetation over
time.
Approach and hypotheses
We studied a 0-30 year chronosequence of mountain pine beetle (MPB; Dendroctonus
ponderosae Hopkins) disturbance in lodgepole pine (Pinus contorta var. latifolia Dougl.) forests
of the Greater Yellowstone Ecosystem. Specifically, we asked how litter quantity and quality,
soil nitrogen pool and fluxes, and foliar nitrogen varied with time-since-beetle outbreak and
evaluated several hypotheses. We expected litter depth, litter mass, and total litter N
concentration to increase during an outbreak and decrease to below undisturbed levels 30 years
after an outbreak. As observed by Morehouse et al. (2008) for Pinus ponderosa, we expected N
concentrations in fresh needle litter to be elevated during the outbreak due to lack of N resorption
prior to litterfall. For soil inorganic N, we expected nitrate and ammonium pools to increase
during an outbreak but not differ from undisturbed stands by 30 years after an outbreak..
Expectations for net N mineralization and net nitrification during the outbreak were less clear;
abiotic changes may stimulate mineralization rates, but microbial populations, surviving
vegetation and a pulse of needle litter could also act to immobilize soil N. By 30 years after an
outbreak, however, we expected N mineralization to be lower than undisturbed stands because
soils would likely be warmer and drier under an open canopy, and litter inputs would be
diminished relative to undisturbed stands since canopy biomass is not yet recovered. Falling
5
beetle-killed snags may also reduce soil N availability and turnover during this period because
coarse wood may be a net N sink (Fahey et al. 1985, Busse 1994, Laiho and Prescott 1999) and
can reduce decomposition and mineralization by altering microsite conditions (Remsburg and
Turner 2006, Metzger et al. 2008). For foliar nitrogen in lodgepole pine, we expected the foliar N
concentration of surviving trees to increase during an outbreak and be positively correlated with
net N mineralization, but total foliar N pool to decline with tree mortality. We further expected
that these differences would persist 30 years after an outbreak because canopy biomass would
still be greatly reduced relative to undisturbed stands, increasing resource availability per tree
even if soil N pools and fluxes return to or fall below undisturbed levels. In the understory sedge
Carex geyerii, we similarly predicted foliar N concentration to increase during an outbreak and
be positively correlated with N mineralization.
METHODS
Study area
This study was conducted in subalpine forests of the Greater Yellowstone Ecosystem
within Yellowstone National Park and Bridger-Teton National Forest in northwestern Wyoming,
USA. Lodgepole pine is a common forest type of the region, and most soils are nutrient-poor and
derived from volcanic rhyolitic and andesitic deposits. Climate is characterized by cool winters
and dry summers, with mean July temperatures of 12.8 ºC and mean January temperatures of -
11.7 ºC at Yellowstone Lake (WRCC 2010b). Mean annual precipitation is 563 mm, mostly
occurring as snow, and increases with elevation (Dirks and Martner 1982). Extensive outbreaks
of the MPB occurred in southern and western Yellowstone National Park in the 1970s and 1980s
(Lynch et al. 2006). Since 2003, large areas of forest in the eastern and southern portions of
6
Greater Yellowstone have been affected by outbreaks of MPB in lodgepole pine and whitebark
pine (Pinus albicaulis Engelmann), Douglas-fir beetle (Dendroctonus pseudotsugae Hopkins) in
Douglas-fir (Pseudotsuga menziesii [Mirb.] Franco), and spruce beetle (Dendroctonus rufipennis
Kirby) in Engelmann spruce (Picea engelmannii Parry). This study focused on lodgepole pine
forests affected by MPB in both the current outbreak and previous outbreaks of the 1970’s and
1980’s.
Sampling design
We sampled a time-since-beetle (TSB) chronosequence including four classes with five
replicates each (N = 20 plots). TSB classes included undisturbed stands; stands within the current
outbreak, both red stage (2 years post-outbreak) and gray stage (4 years post-outbreak); and
stands attacked by bark beetles approximately 30 years ago. Potential sites were identified using
maps of current (USFS 2006) and historic (Lynch et al. 2006) mountain pine beetle outbreaks,
soil type, and forest age. Field inspection assured selection of stands with comparable species
composition, basal area, soils, and bark beetle mortality. Undisturbed and 30-year TSB stands
were located in Yellowstone National Park, while red and gray stands were located within the
current outbreak on the Bridger Teton National Forest (Figure 1). Disturbed plots were located
within stands (0.25 ha) of homogenous structure and beetle disturbance intensity; undisturbed
sites were located in comparable beetle-susceptible stands. In lodgepole pine, both beetle
susceptibility (Safranyik and Carroll 2006) and ecosystem N status (Smithwick et al. 2009) are
largely determined by stand structure and age. Thus, site selection based upon stand structure,
age, and substrate largely controlled for pre-disturbance differences in these factors among
classes. To validate the chronosequence, dendrochronological analyses were used to determine
post-fire stand age for all plots and to reconstruct pre-outbreak stand structure and MPB outbreak
7
severity for the 30 year TSB plots (Simard et al. 2011). An 8-m radius plot (201 m2) was
established at each site in summer 2007, and slope and aspect were recorded at the plot center.
Microclimate
To evaluate variation in growing-season microclimate, soil and air temperature were
measured hourly in three of the five plots within each TSB class (N = 12 instrumented plots)
from June 18 through August 7, 2008, using three pairs of iButton datalogger probes (Maxim
Integrated Products Inc., Dallas Semiconductor, Sunnyvale, CA) per plot. One iButton of each
pair was installed at the litter-organic soil interface, and the second was installed 10 cm below
this interface. Air temperature was recorded using another temperature probe hung inside a
covered but well-ventilated and open-bottomed white PVC housing attached to a tree at breast
height near the plot center.
For each soil depth, temperature probe data were summarized as follows. First, plot-
level hourly temperatures were determined by averaging data from the three probes per plot. To
account for geographic variation and normalize comparisons among plots, hourly differences
between air and soil temperature were calculated (difference = soil – air). Plot-level daily mean
temperatures, temperature ranges, and soil-air temperature differences were then calculated from
the hourly data, and averaged by class to determine class-level daily means. Plot-level growing
season mean temperatures, ranges, and soil-air temperature differences were calculated by
averaging daily means across the sampling period, and class-level growing season means were
then calculated from these plot-level data.
Vegetation
All live and dead standing trees > 1.4 m tall were identified to species and measured for
diameter at breast height (DBH). MPB-killed trees were identified by the presence of pitch tubes
8
(tree resin accumulation at boring hole entrances), boring dust, and J-shaped galleries under the
bark (diagnostic of Dendroctonus (Safranyik and Carroll 2006)). Downed logs rooted in the plot
were also measured for DBH, identified to species, and scored as MPB-killed if J-shaped
galleries were visible. Trees < 1.4 m tall were measured by 10 cm height classes in the northeast
quadrant of the plot (50m2). Pre-outbreak basal area for red and gray stands was calculated by
summing the basal area of live and beetle-killed trees. In the 30-year TSB stands post-outbreak
growth of survivors was substantial, so pre-outbreak basal area was determined from
dendrochronology using stand reconstruction techniques (Simard et al 2011). Beetle-killed basal
area in the 30-year TSB stands was determined by summing the basal area of downed logs with
evidence of MPB galleries. Lodgepole pine canopy biomass was calculated for each plot using
allometrics developed for this region (Brown 1978). Ground cover was visually estimated to the
nearest 10% by plant functional groups (forbs, sedges, grasses, and seedlings) in ten 0.25-m2
circular microplots. The microplots were located within the inner 5-m plot radius using a
stratified random design of fixed distances (one at 0.5-m; two at 1.5-m, 2.5-m and 3.5-m; and
three at 4.5-m) and random bearings (in 10º increments) from the plot center, and averaged by
plot. Biomass of the sedge Carex geyerii was calculated from allometrics previously reported for
Yellowstone National Park (Turner et al. 2004).
Litter quantity and quality
In each 0.25-m2 microplot, litter depth was recorded at three locations and a 400-cm2
sample of the litter layer was collected and oven-dried at 60 ºC. Plot-level litter depth and mass
were obtained by averaging values from the 10 microplots. Litter from each microplot was sorted
into three categories: fresh current-year needle litter, identified by bright red color and lack of
mottling on surface (Morehouse et al. 2008); all-needle litter (all ages of needles combined); and
9
total litter (all foliar litter components and woody litter components < 1.0 cm wide). Sorted litter
was composited by plot for each litter category, and ground to powder for C:N analysis on a
Leco CNS-2000 at the University of Wisconsin Soil and Plant Analysis Laboratory (UWSPAL
2010)
Soil chemistry, N pools, and N fluxes
One soil core was collected from each 0.25-m2 microplot using a 5-cm diameter x 15-cm
long PVC corer. Soils were sieved (2 mm), weighed, and divided into three subsamples: 30 g
oven-dried at 60 ºC for gravimetric percent moisture; 20 g extracted in 75 ml of 2M KCl for 2
hours, with the extract then filtered and frozen for later analysis of NH4+ and NO3
- pools; and 20
g air-dried and bulked by plot for soil texture and chemical analyses. Air-dried soil was analyzed
for pH, total N, exchangeable Ca, Mg, and K, available P (Bray P1 extract), and organic matter
content at the University of Wisconsin Soil and Plant Analysis Laboratory (UWSPAL 2010).
Soil organic N was determined by difference using total N and inorganic N values, and soil
texture was determined using the Bouyoucos hydrometer technique (Bouyoucos 1962).
Net N mineralization and net nitrification were measured using ion-exchange resin cores
(Binkley et al. 1992), with one core incubated in situ for approximately one year (July 2007 -
June 2008) in each 0.25-m2 microplot. Incubated cores consisted of a 5-cm diameter x 15-cm
long PVC tube of soil with an ion-exchange resin bag placed at the bottom. Resin bags were
constructed using 20 g of mixed bed ion exchange resin (J.T Baker #JT4631-1) tied inside a
piece of un-dyed nylon stocking material. Upon retrieval in summer 2008, core soils were sieved
(2 mm) and core soils and resin bags were extracted separately in 2M KCl in the same manner as
the 2007 initial soil samples described above. All KCl extractions were analyzed for [NH4+] and
10
[NO3-] using colorimetric methods on an Astoria Pacific II continuous flow autoanalyzer. For
each microplot, mineralization and nitrification rates were calculated as:
rate = [(final soil N + resin bag N) – initial soil N ] / incubation time
and expressed in units of µg N g soil-1 year-1. Values were then averaged by plot. Atmospheric N
inputs in the region are low (~ 1 kg N ha-1 year-1; and were not factored into the soil N
calculations. (www.epa.gov/castnet/charts/YEL408_totn.png)
Foliar N
In each plot, pole pruners were used to collect canopy foliage from three live mature
lodgepole pines unattacked by MPB. Two fully sunlit branches (0.5-m long) were clipped from
each tree and separated into two subsamples: current-year foliage (from one branch per tree), and
all-years foliage (from the second branch per tree). Understory sedge (Carex geyerii Boott)
foliage was collected by clipping 15-20 pieces from each quadrant of the plot. For both
lodgepole and sedge foliar N, samples from two of the five undisturbed sites were excluded from
the analyses reported here because they were sampled much earlier in the growing-season than
the remaining eighteen sites. All foliar samples were oven-dried at 60ºC and ground to powder
for C and N analysis on a Leco CNS-2000 analyzer at the University of Wisconsin Soil and Plant
Analysis Lab. Tree-level canopy data (N=3 per sample type) and quadrant-level sedge data (N
=4) were averaged by plot. To compute canopy and Carex geyerii N pool sizes, the mean foliar
N concentration for each plot was multiplied by biomass values derived from the allometrics
cited above.
Statistical analyses
All statistical analyses were performed in SAS (SAS Institute Inc. 2003), and unless
otherwise noted all reported variance values are two standard errors. ANOVA was used to test
11
for differences among TSB classes in stand basal area, site characteristics (slope, elevation,
aspect, and soil chemistry), understory cover, soil temperatures, litter depth, litter quantity and
quality, soil N pools, foliar N concentrations, and foliar N pools. When ANOVAs were
significant, Tukey’s HSD test (α = 0.05) was used to identify differences among means.
Because variability in soil characteristics of P. contorta forests in this region is known to
influence nitrogen transformations irrespective of bark beetle outbreaks (Smithwick et al. 2005),
we used general linear models to account for potential covariate effects (soil cations, C, OM, pH,
bulk density, and C:N ratio) when testing for differences in net mineralization and net
nitrification rates among TSB classes. To explore which beetle-induced changes were related to
differences in soil N cycling rates among TSB classes, we used linear regression to identify
relationships among litter, soil temperature, and N mineralization variables. Because only three
out of five plots per TSB class were instrumented with soil temperature probes, regressions
including soil temperature variables have N = 12 rather than N = 20. We also used linear
regression to test whether foliar N concentrations in unattacked P. contorta and C. geyerii were
positively related to N net mineralization rates.
RESULTS
Stand structure and site characteristics
Pre-disturbance stand structure and disturbance severity were similar across the
chronosequence. There were no significant differences among TSB classes in pre-outbreak post-
fire stand age (168 ± 18 years), pre-outbreak P. contorta basal area (43.4 ± 2.4 m2 ha-1 ),
background P. contorta mortality (7.2 ± 1.1 m2 ha-1), or beetle-killed P. contorta basal area
12
(disturbed classes only; 33.7 ± 3.2 m2 ha-1) (Simard et al. 2011). In the 30-year TSB sites, peak
mortality occurred 29 ± 2 years prior to sampling (Simard et al. 2011).
Topographic position (elevation, aspect, and slope) did not vary among undisturbed, red,
and gray stands (Table 1). However, the 30-year TSB sites were on average 260 m lower in
elevation than other classes and had shallower slopes than gray sites (Table 1). Soil texture was
similar among TSB classes, with only percent sand and bulk density being slightly lower in the
30-year TSB compared to gray stands (Table 1). Soil K (average 160 ± 9 ppm), P (average 18 ±
2 ppm), and percent organic matter (average 4.5 ± 0.5) did not differ among TSB classes (Table
1). Soil pH, calcium and magnesium differed with the regional distribution of sites, with slightly
higher pH (0.5 pH units) and 2-3 fold higher [Ca] and [Mg] in the Bridger Teton National Forest
sites (red and gray) than in the Yellowstone National Park sites (undisturbed and 30-year TSB)
(Table 1).
Total percent cover of understory vegetation (F = 4.49, R2 = 0.46, P = 0.02) and percent
cover of forbs (F = 7.06, R2 = 0.57, P = 0.003) varied with TSB class (Figure 2). Total
understory cover in the red class averaged 60%, approximately double that of undisturbed forest;
understory cover in the gray and 30-year TSB classes was intermediate (Figure 2). Forb cover
averaged 25% in the red class, approximately 2.5x greater than in undisturbed forest (Figure 2).
Grass, sedge and shrub cover did not vary with TSB class (P > 0.05).
Microclimate
Air and soil temperatures during the 2008 growing season differed among TSB classes,
and the 30-year TSB class was both the warmest and the most variable (Table 2). Mean daily air
temperature in the 30-year TSB class was ~1 ˚C warmer than undisturbed, red, and gray classes,
and the mean daily range of temperature extremes was greater (Table 2). Temperatures at the
13
litter-soil interface varied more among classes than did air temperatures, and were 4-6 ˚C warmer
in the 30-year TSB class compared to undisturbed, red, and gray. Temperatures at the litter-soil
interface did not differ from air temperature in the undisturbed class, but were ~2˚C cooler than
air temperature in the red and gray classes, and 3˚C warmer than air temperature in the 30-year
TSB class (Table 2). At 10 cm soil depth, temperatures were 3-4˚C warmer in the 30-year TSB
class compared to the undisturbed, red, and gray. Soil temperatures at 10 cm depth were always
cooler than air temperatures but were 5.3 ± 0.2 ˚C less than air in undisturbed, red, and gray
sites, and only 2.5 ± 0.9˚C less than air in the 30-year TSB class (Table 2). There were no
differences among classes in the slope or aspect of instrumented sites (Table 2). Mean elevation
of the instrumented undisturbed sites was not significantly different from the instrumented sites
in any other class (Table 2). Thus, differences in soil temperature between undisturbed and any
disturbance class suggest an effect of TSB rather than topographic differences among classes.
Litter quantity and quality
Total litter depth was approximately 3 cm in red and gray stands, about twice that of
undisturbed and 30-year TSB stands (Figure 3a). However, total litter mass averaged 1577 ± 149
g m-2 and did not vary with TSB class (Figure 3b). The N concentration in fresh current- year-
needle litter (Figure 3c) and all-needle litter (Figure 3d) varied with TSB class, with the highest
values (~0.75%) observed in the red and gray classes and lower concentrations (~0.4-0.5%) in
the undisturbed and 30-year TSB classes. However, N concentration of total litter averaged 0.78
± 0.02% and did not vary with TSB class (Figure 3e). Litter N pool size averaged 13.8 ± 1.2 g N
m-2 in undisturbed, red and gray stands and was significantly lower (7.4 ± 2.5 g N m-2) in the 30-
year TSB class compared to the gray (Figure 3f).
14
Soil N pools and fluxes
Among soil N pools, only soil ammonium pool size varied with TSB class (Figure 4a).
Extractable NH4+ was six times greater in the red stands (4.75 µg N g soil-1) compared to the
undisturbed stands (0.74 µg N g soil-1) and intermediate in the gray and 30-year TSB stands.
Extractable NO3- was low (averaging 0.63 ± 0.20 µg N g soil-1) and did not vary with TSB class
(Figure 4b). Total extractable inorganic N averaged 3.11 ± 0.63 µg N g soil-1 (data not shown)
and also did not vary with TSB class, nor did soil C:N ratio, total percent soil N, or soil organic
N pool (Table 1).
Annual net N mineralization rate was approximately double (21 ± 7 µg N g soil-1 year-1)
in the red and gray stages relative to the undisturbed and 30-year TSB classes (Figure 4c). After
accounting for variation due to soil bulk density, pH and C:N ratio, the effect of TSB class on net
N mineralization was significant; collectively, these variables explained 70% of the variation in
net N mineralization (Table 3). Similar to annual net N mineralization, annual net nitrification
rate was highest in the red stage (8.2 ± 4.2 µg N g soil-1 year-1) and about three times greater than
in the undisturbed and 30-year TSB stages (Figure 4d). After accounting for soil bulk density,
there was a significant effect of TSB class on annual net nitrification (Table 3). Net N
mineralization and net nitrification were positively correlated (R2 = 0.66, P < 0.0001), and
nitrification fraction (ratio of nitrification to mineralization) averaged 0.32 ± 0.05 and did not
vary among TSB classes (data not shown). In post-incubation soils, neither nitrate concentration
(average 2.23 ± 0.67 µg N g soil-1) nor the ratio of nitrate to total inorganic N (average 0.29 ±
0.07) differed among classes (P = 0.1314 and P = 0.1710, respectively; data not shown).
Net nitrogen mineralization rates were more strongly correlated with soil temperature
(Figure 5b; R2=0.39) than with litter quality variables (Figure 5c-f; R2 <0.19). Increased litter
15
depth in the red and gray classes corresponded to decreased soil temperatures, and shallow litter
layers in the 30-year TSB class were associated higher soil temperatures (Figure 5a). Net N
mineralization rates declined with increasing soil temperature across the TSB chronosequence
(Figure 5b). Net N mineralization was weakly related to the nitrogen concentration in fresh
needlefall (Figure 5c), but showed no relationship to the nitrogen concentration in all-needle
litter or total litter, or to the total litter N pool size (Figures 5d-f).
Foliar biomass and N
Foliar biomass of live lodgepole pines was 69% lower in the current outbreak (red and
gray sites) relative to undisturbed sites, and 48% lower in the 30-year TSB class (Table 4). Pre-
outbreak foliar biomass of beetle-killed trees in the red and gray classes averaged 687 ± 49 g m-2
(data not shown), which also represents the total mass of needle litterfall induced by the
outbreak. The N concentration in current year foliage of unattacked P. contorta averaged 0.99 ±
0.02% in the red, gray and 30-year TSB classes compared to 0.82% in the undisturbed, and foliar
C:N was concomitantly lower in the disturbed classes compared to undisturbed forest (Table 4).
The N concentration in all-age foliage of unattacked P. contorta was highest in the red and gray
classes and lower in the undisturbed and 30-year TSB classes (Table 4). P. contorta foliar N pool
averaged 8.3 g N m-2 in undisturbed forest, was about 64% lower in the red and gray classes, and
48% lower in the 30-year TSB class (Table 4). Across all TSB classes, the N concentration of
current year P. contorta foliage was positively related to soil N mineralization (Figure 6a).
Biomass of the understory sedge Carex geyerii did not differ across the chronosequence (Table
4), but foliar N concentration of C. geyerii was 50% greater (1.5-1.6% N) in red and gray stands
compared to undisturbed and 30-year TSB stands. Foliar N concentration of Carex geyerii was
also positively related to annual net N mineralization across all TSB classes (Figure 6b).
16
DISCUSSION
This study demonstrates the substantial effects of MPB outbreak on N cycling through
the litter, soils, and vegetation of lodgepole pine forests across a replicated chronosequence of
time-since-beetle disturbance. Beetle-induced tree mortality triggered a cascade of effects
through increased needlefall and reduced N uptake, which in turn altered soil microclimate,
increased soil N availability, and increased N concentration of surviving vegetation in recent
outbreaks. Thirty years after the disturbance soil N parameters returned to undisturbed levels,
though changes in soil temperature, %N of new foliage, canopy biomass, canopy N pool size;
and the % N of needle litter were still evident. Our results also suggest several mechanisms that
contribute to N retention in beetle-affected lodgepole pine forests and make N loss from the
system unlikely.
Microclimate
A strong and persistent effect of bark beetle outbreak was observed on soil microclimate
during the growing season. Within the current outbreak, soils were notably cooler than in
undisturbed sites, most likely in response to increased litter depth. Though soil moisture was not
measured in this study, soil moisture probably increased due to deeper litter and reduced
transpiration caused by tree mortality (Stottlemyer and Troendle 1999, Tan et al. 2008, Griffiths
et al. 2010). Experimental additions of needle litter have been shown to increase soil moisture in
lodgepole pine forests even in the absence of tree mortality (Cullings et al. 2003). In the 30-year
TSB sites soils were warmer and likely drier than in undisturbed forest, which can decrease
decomposition (Remsburg and Turner 2006) and net N mineralization (Metzger et al. 2008) rates
in post-fire lodgepole pine forests. Though we did not measure soil moisture or decomposition
17
directly, our results suggest similar mechanisms may be important in beetle-disturbed lodgepole
pine forests as well.
Litter quantity and quality
The elevated N concentration of fresh needle litter in the current outbreak was consistent
with our hypothesis and prior studies, and is likely due to a lack of N resorption prior to needle
drop (Morehouse et al. 2008). However, we were surprised that the mass, N concentration, and
N pool size of the total litter layer did not differ as expected. Gradual inputs of litterfall over
several years combined with ongoing litter decomposition may explain why litter mass did not
increase significantly. Foliar mass of beetle-killed trees in red and gray stands was estimated to
be 687 ± 49 g m-2 (data not shown), which is approximately half the mass of litter observed in
undisturbed stands. Thus, if all beetle-killed foliage fell at once one may expect a 50% increase
in the litter mass of attacked stands. However, beetle-induced litterfall extends over 3-4 years,
and P. contorta litter can lose 30% of its mass after two years (Remsburg and Turner 2006).
These processes may explain why the mass gained in the litter layer is smaller than the mass lost
from the canopy by ~200 g m-2, and why increases in litter mass following beetle outbreak (498
g m-2) are small relative to the amount of litter already present on the forest floor prior to
outbreak (1512 ± 457 g m-2).
Thirty years following outbreak, we also did not see the hypothesized declines in total
litter biomass, N concentration or N pool size relative to undisturbed forest, even though stand
basal area and therefore fresh litter inputs had not yet recovered. We suggest that persistent
warming and drying of the forest floor and drier soil conditions under elevated coarse wood
(Remsburg and Turner 2006) may have reduced decomposition rates during the post-outbreak
period enabling more litter accumulation despite reduced input rates. The elevated N
18
concentration in all-aged needle litter also may represent a legacy effect of outbreak-induced
litterfall undergoing an extended period of decomposition and N immobilization.
Soil N pools and fluxes
Although we expected both soil NH4+ and NO3
- to increase during the outbreak, only soil
NH4+ was elevated despite observed increases both net N mineralization and net nitrification.
Increased soil NO3- has been observed following bark beetle outbreaks in other forest types
(Huber 2005), however several studies of lodgepole pine have shown that mortality must be
extensive before soil NO3- increases. Parsons et al. (1994) found elevated NO3
- only in large
experimental root gaps in which 30 trees were killed, and Knight et al. (1991) found no elevated
NO3- in soil solution following a 60% thinning but a 10-40 fold increase following clearcut.
Plot-level tree mortality in this study averaged 76%, but surviving trees were intermixed with
beetle-killed trees. Thus, undisturbed vegetation and microbial biomass were likely able to
assimilate the products of increased nitrification yielding no change in NO3- pool size. In contrast
to many N-saturated forests of Europe and Eastern North America where atmospheric N
deposition is high and nitrate levels can rise dramatically following disturbance (Huber 2005;
Bormann and Likens 1979), conifer forests of western North America receive little N deposition
and often show little response of nitrate following disturbance (Prescott et al. 2003; Titus et al.
2006).
Observed rates of net N mineralization and net nitrification in the undisturbed stands are
within ranges reported by other studies from Greater Yellowstone for both mature (Turner et al.
2007) and 15 year-old post-fire (Metzger et al. 2008) lodgepole pine forests. Nitrification is
typically low (<2 µg N g soil-1 year-1) in undisturbed lodgepole pine (Turner et al. 2007), and
despite a significant increase in the red stage, net nitrification was never above 8.2 µg N g soil-1
19
year-1 in any class. The magnitude of increased mineralization in the current outbreak is
comparable to that reported by Parsons et al. (1994) for artificial root gaps of 15 lodgepole pine
trees two years after disturbance, but less than increases reported for silvicultural gaps in
Douglas-fir forests 6-8 years after disturbance (Thiel and Perakis 2009). Though we did not see
any change in net N mineralization in the 30-year TSB stands relative to the undisturbed stands,
decadal-scale effects of lodgepole pine mortality on net N mineralization have been documented
in other studies. Twenty-four years following treatment, Tan et al. (2008) observed twice the
rate of N mineralization in lodgepole stands thinned to 1600 stems ha-1, compared to unthinned
stands. However, stands in that study were much younger (22 years) at the time of disturbance
than ours (average 168 years), soils were generally more fertile (e.g., soil C:N ratios in control
and thinned plots ranged from 20-33), and thinning residues were removed.
Foliar N and biomass
Nitrogen concentration of new lodgepole pine foliage demonstrated a rapid response of
unattacked trees to increased nutrient availability. The foliar N concentration of current-year
needles was positively correlated with net N mineralization, and lodgepole pine is known to take
up both NH4+ and NO3
- (Hawkins et al. 2008). Fertilization studies in lodgepole pine have also
demonstrated rapid response of foliar N to increased soil N availability, though this may
attenuate after several years as foliage biomass increases (Blevins et al. 2005). Increased foliar N
concentration in unattacked trees provided a mechanism for N retention in the canopy following
bark beetle disturbance. Between the undisturbed and gray classes, beetle outbreak caused a 68%
reduction in foliar biomass but only a 58% reduction in canopy N pool size.
The canopy N pool is an important regulator of ecosystem N cycling (Prescott 2002), and
declines in canopy N pool size paralleled declines in live foliar biomass during the current
20
outbreak. Foliar N concentrations for undisturbed stands were similar to values previously
reported for the Yellowstone region (Litton et al. 2004) and elsewhere (Binkley et al. 1995,
Blevins et al. 2005). However, we were surprised to observe that foliar N concentration in
current-year needles remained elevated in the 30-year TSB stands, although N concentration of
composite foliage returned to undisturbed levels. Net N mineralization rates were comparable in
undisturbed and 30-year TSB stands but foliage biomass remained substantially lower; thus, the
availability of N per unit of foliage biomass was likely greater in the 30-year stands.
A flush of understory growth has been observed in many post-beetle conifer forest types,
including lodgepole pine (Stone and Wolfe 1996), Douglas-fir (McMillin and Allen 2003),
Engelmann spruce (Picea engelmanni) (Schmid and Frye 1977), subalpine fir (Abies lasiocarpa)
(McMillin et al. 2003), and ponderosa pine (McCambridge et al. 1982). Reduced competition
and increased light, moisture and nutrients are likely stimulants of this growth. As forbs prefer
NO3- uptake over NH4
+ in deciduous woodland (Falkengren-Grerup et al. 2004) and alpine
meadow (Miller and Bowman 2002) systems, increased forb cover observed in this study may be
related to increased net nitrification. The positive relationship between N mineralization and
Carex geyerii %N also suggests that herbaceous species benefit from increased N availability
during bark beetle outbreak. Foliar %N of Carex geyeri can increase after fertilization in
ponderosa pine forests (VanderSchaaf et al. 2004), and in lodgepole pine C. geyerii %N can
increase following fire-induced changes in soil N availability. Interestingly, the increased N
concentration we observed in C. geyerii is comparable to the increase observed two years after
stand-replacing fire in lodgepole pine forests of the same region (Metzger et al. 2006).
21
Interpretations and comparison to post-fire lodgepole pine forests
Multiple lines of reasoning suggest that increased rates of net N mineralization and net
nitrification during the current outbreak are likely not due to a pulse of litter N associated with
beetle-killed trees. Beetle outbreak did not significantly change the mass or N pool size of total
litter, and neither was correlated with N mineralization rates—a pattern that has been noted
across a wide range of North American forests (Scott and Binkley 1997). Furthermore, fresh
lodgepole pine litter immobilizes N for several years, acting as a net sink of N during this period
rather than a source (Fahey 1983, Remsburg and Turner 2006). Thirdly, increases in net N
mineralization following tree mortality are often limited to mineral soil horizons and not found in
the forest floor (Morehouse et al. 2008, Tan et al. 2008, Thiel and Perakis 2009), which suggests
belowground processes and abiotic changes may be more important than litter N inputs. Labile C
leaching from fresh litter or decaying roots could play a role in stimulating gross N production
and N consumption if microbial communities were C limited, although Giardina et al. (2001)
found lodgepole pine soils of the intermountain West to have relatively large amounts of high
quality soil C, with no relationship between C and N mineralization rates. Alternatively, declines
in belowground C transfer beyond the rhizosphere (Högberg et al. 2010) following beetle-
induced mortality could lead to decreased N demand by microbes and thus greater net N
mineralization rates and extractable N pools. Further study is needed to understand the role of C
availability on N transformations following bark beetle disturbance.
Despite the increases in available soil N associated with the current beetle disturbance,
surviving lodgepole pine did not appear to have an excess of available N. Reduced competition
and increased net N mineralization increased nutrient availability for surviving unattacked trees,
as evidenced by a 20-30% increase in foliar N concentration. However, N concentrations in
22
current year foliage remained < 1.2%, suggesting N may still be limiting (Moore et al. 2004).
When foliar N concentrations are considered along with the lack of elevated soil nitrate pools,
our results also suggest that the risk of nitrogen loss via NO3- leaching following bark beetle
outbreak is low.
The effects of bark beetle disturbance on soil and canopy N were substantially less than
observed following fire in Greater Yellowstone. In two-year post-fire lodgepole pine forests,
NH4+ and NO3
- pools were four times greater than in the red stage of a bark beetle outbreak (20
vs 5 µg NH4+-N g soil-1, 2 vs. 0.5 µg NO3
--N g soil-1), and net N mineralization and net
nitrification rates were twice as great (18 vs. 9 µg N g soil-1 year-1 for mineralization, 16 vs. 8 µg
N g soil-1 year-1 for nitrification) (Turner et al. 2007). Several studies also show that foliar N
concentration of lodgepole pine is higher after fire than after bark beetle disturbance. Romme et
al. (2009) found very high (1.87%) foliar N in 3 to 5-year old post-fire lodgepole pine seedlings,
and Turner et al. (2009) found current-year foliage averaged 1.38% N and composite foliage
1.08% N in 17-year old post-fire lodgepole pine. Contrary to fire, bark beetles typically affect
large mature trees with no direct disturbance to the understory or soils. Observed changes in soil
N following bark beetle outbreak are similar to those following other selective agents of
mortality in forests. In eastern hemlock (Tsuga canadensis) forests affected by the hemlock
wooly adlegid (Adelges tsugae Annand), Orwig et al.(2008) documented increased
mineralization, nitrification, and N pool sizes relative to undisturbed stands. Generally, the
consequences of bark beetle outbreak for ecosystem processes are likely similar to any agent of
mortality that selectively kills canopy trees and does not disturb soils or understory directly.
Though our chronosequence approach is strengthened by replication and by validation
from companion dendrochronological analyses (Simard 2011), there were some noteworthy
23
differences among classes. Current outbreak sites had greater soil pH, Ca, and Mg which may
have contributed directly to greater understory cover of forbs. The influence of soil cations on
soil N transformations, however, is considered to be indirect through control on species
composition and subsequent litter quality (Page and Mitchel 2008). Since all sites were
dominated by lodgepole pine, differences in soil cations were unlikely to influence soil N
parameters. Though not significant, thirty year old outbreak sites tended toward greater soil
organic matter and total soil N, which may have resulted from organic N inputs during the
outbreak rather than inherent pre-disturbance site differences.
CONCLUSIONS
Bark beetle disturbance significantly altered N cycling through the litter, soil, and
vegetation of lodgepole pine forests. Beetle-induced litterfall increased litter depth and altered
soil microclimate. Increased soil N in recently disturbed sites was more closely related to cooler
soil temperatures than to litter quality, and was positively correlated with increased foliar N in
unattacked vegetation. Effects of bark beetle disturbance on needle litter %N, soil temperature,
and canopy N pool size persisted 30 years following outbreak. However, changes in soil N
cycling during the outbreak were of lesser magnitude than those observed following stand
replacing fire. Although significant, the net effects of bark beetle disturbance on N cycling in
lodgepole pine were surprisingly minor given the extent of beetle-caused tree mortality.
ACKNOWLDEGEMNTS
We thank Bill Romme, Gary Lovett, Emily Stanley, and Erica Smithwick, and two
anonymous reviewers for constructive comments that improved this manuscript. For help in the
24
field and the laboratory, we appreciate the assistance of many University of Wisconsin students:
Jaclyn Entringer, Heather Lumpkin, Lucille Marescot, Erin Mellenthin, Corey Olsen, Ryan
Peasley, Alex Rahmlow, Amanda Rudie, Ben Ruh, and Greg Skupien.
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31
TABLES
Table 1. Site characteristics across a chronosequence of mountain pine beetle disturbance
in P. contorta forests. Superscript letters next to values denote significant differences among
classes (Tukey’s test, α = 0.05). N = 5 per class. Error ranges = 2 SE.
w ANOVA P value; n.s. = non significant x Post-fire stand age at time of mountain pine beetle disturbance. y Data for 30 year TSB sites are from Simard (2011). z Values are an index of southwest-ness ranging from -1 to 1, calculated as: cosine(aspect-225)
TSB Class Site variable Undisturbed Red Gray 30 years P w
Stand agex (year) 179 ± 46 a 164 ± 32 a 176 ± 43 a 152 ± 28 a n.s.
P. contorta basal area (m2ha-1)
Pre-outbreak live y 42.7 ± 10.0 a 48.3 ± 11.9 a 40.2 ± 7.7 a 42.4 ± 9.4 a n.s.
Beetle-killed ---- 41.7 ± 10.5 a 26.6 ± 6.5 a 33.0 ± 12.7 a n.s
Dead, non-beetle killed 5.6 ± 2.4 a 6.0 ± 2.9 a 10.9 ± 3.0 a 6.5 ± 7.2 a n.s.
Topography
Elevation (m) 2400 ± 17 a 2487 ± 34 a 2476 ± 48 a 2218 ± 160 b 0.001
Slope (º) 8 ± 7 ab 10 ± 9 ab 21 ± 11a 2 ± 2 b 0.025
Aspect z 0.10 ± 0.94 a 0.30 ± 0.66 a 0.56 ± 0.26a 0.04 ± 0.81 a n.s.
General Soil Characteristics
Sand (%) 62 ± 6 ab 58 ± 5 ab 69 ± 7 a 55 ± 6 b 0.034
Silt (%) 26 ± 6 a 27 ± 5 a 20 ± 6 a 31 ± 5 a n.s.
Clay (%) 13 ± 1 a 15 ± 2 a 11 ± 2 a 13 ± 2 a n.s.
Bulk density (g cm-3) 0.7 ± 0.1ab 0.7 ± 0.2 ab 0.9 ± 0.2 a 0.6 ± 0.1 b 0.042
pH 4.9 ± 0.1 b 5.4 ± 0.1 a 5.2 ± 0.1 a 4.9 ± 0.2 b <0.001
Ca (ppm; exchangeable) 452 ± 63 b 1192 ± 651 a 857 ± 244 ab 322 ± 182 b 0.014
Mg (ppm;exchangeable) 67 ± 12 b 143 ± 54 a 117 ± 35 ab 71 ± 21 b 0.018
K (ppm; exchangeable) 162 ± 22 a 186 ± 60 a 149 ± 29 a 145 ± 23 a n.s.
P (ppm; Bray P1) 17 ± 4 a 25 ± 12 a 20 ± 9 a 11 ± 3 a n.s.
OM (%) 3.9 ± 0.8 a 5.3 ± 3.9 a 3.3 ± 0.4 a 5.3 ± 1.4 a n.s.
C:N 72 ± 23 a 42 ± 26 a 41 ± 7 a 42 ± 11 a n.s.
Total N (%) 0.04 ± 0.03 a 0.05 ± 0.03 a 0.04 ± 0.01a 0.08 ± 0.04 a n.s.
Organic N (µg N g soil-1) 368 ± 271 a 509 ± 306 a 446 ± 109 a 759 ± 416 a n.s.
32
Table 2. Air and soil temperatures across a chronosequence of mountain pine beetle
disturbance. Superscript letters next to values denote significant differences among classes
(Tukey’s test, α = 0.05). N = 3 per class. Error ranges = 2SE.
w ANOVA P value; n.s. = non significant x Values are an index of southwest-ness ranging from -1 to 1, calculated as: cosine (aspectº-225)
TSB Class
Variable Undisturbed Red Gray 30 years P w
Air temperature (ºC)
Mean daily temperature 15.3 ± 0.4 b 15.1 ± 0.3 b 15.0 ± 0.5 b 16.3 ± 0.5 a 0.006
Mean daily range 19.4 ± 1.4 b 21.8 ± 0.8 ab 20.5 ± 1.6 b 25.1 ± 2.8a 0.012
Litter-soil temperature (ºC)
Mean daily temperature 15.4 ± 1.0 b 13.3 ± 0.8 bc 13.0 ± 0.6 c 19.2 ± 1.4 a < 0.001
Mean daily range 27.4 ± 3.5 b 21.8 ± 3.6 b 19.3 ± 1.0 b 38.0 ± 7.0 a 0.001
Mean daily difference from air 0.1 ± 0.8 b -1.8 ± 1.0 b -2.0 ± 0.2 b 2.8 ± 1.4 a <0.001
10 cm soil depth temperature (ºC)
Mean daily temperature 10.5 ± 0.3 b 9.4 ± 0.1 b 9.7± 0.4 b 13.8 ± 1.3 a < 0.001
Mean daily range 2.4 ± 0.2 a 2.9 ± 0.8 a 2.8 ± 0.6 a 4.6 ± 1.8 a n.s.
Mean daily difference from air -4.8 ± 0.1 b -5.6 ± 0.3 b -5.3 ± 0.8 b -2.5 ± 0.9 a < 0.001
Togography of instrumented sites
Elevation (m) 2401 ± 4 ab 2471 ± 40 a 2494 ± 68 a 2198 ± 227 b 0.0301
Slope (º) 10.8 ± 10.6 a 6.2 ± 3.4 a 13.7 ± 5.5 a 2.5 ± 3.2 a n.s.
Aspect x 0.26 ± 1.25 a 0.42 ± 0.70 a 0.55 ± 1.34a -0.09 ± 0.67 a n.s.
33
Table 3. General linear models of net N mineralization and net nitrification across a
chronosequence of mountain pine beetle disturbance in P. contorta forests. N = 5 per class.
Response variable Model Fit (R2, F, P) Parameter Estimate F P
Net N mineralization 0.70, 5.04, 0.007 TSB class -- 5.30 0.01
Undisturbed -- -- --
Red 18.2 -- 0.02
Gray 16.2 -- 0.03
30 years -8.6 -- 0.09
Bulk density -32.02 7.68 0.02
pH -22.17 4.37 0.06
C:N -0.19 5.66 0.03
Intercept 153.9 -- 0.02
Net nitrification 0.52, 4.06, 0.02 TSB class -- 4.58 0.02
Undisturbed -- -- --
Red 1.3 -- 0.01
Gray 1.2 -- 0.04
30 years -0.31 -- 0.53
Bulk density -3.56 8.52 0.01
Intercept 4.01 -- 0.0005
34
Table 4. Foliar N in unattacked lodgepole pine (P. contorta) and an understory sedge (C.
geyerii) across a chronosequence of mountain pine beetle disturbance in P. contorta forests.
Superscript letters denote significant differences among classes (Tukey’s HSD test, α = 0.05).
Error ranges = 2SE.
TSB Class
Foliage variable Undisturbed Red Gray 30 years P w
Pinus contorta
New foliage
N (%) 0.82 ± 0.03 b 1.05 ± 0.12 a 1.00 ± 0.06 a 0.94 ± 0.05 a 0.016
C:N 59 ± 4 a 49 ± 6 b 50 ± 2 b 53 ± 2 ab 0.022
All foliage
N (%) 0.75 ± 0.04 b 0.91 ± 0.08 a 0.91 ± 0.09 a 0.75 ± 0.12 b 0.014
C:N 63 ± 3 a 52 ± 3 b 53 ± 5 b 64 ± 10 ab 0.017
Biomass
(g m-2) 1107 ± 214 a 331 ± 88 b 358 ± 169 b 571 ± 78 b <0.0001
N pool size
(g N m-2) 8.3 ± 1.7 a 3.0 ± 0.9 b 3.5 ± 1.5 b 4.3 ± 0.8 b <0.0001
Carex geyerii
N (%) 1.1 ± 0.2 b 1.6 ± 0.4 a 1.5 ± 0.1 ab 1.1 ± 0.1 b 0.016
C:N 43 ± 9 a 28 ± 7 b 29 ± 2 ab 41 ± 4 ab 0.024
Biomass
(g m-2) 46 ± 46 a 32 ± 42 a 52 ± 42 a 179 ± 129 a n.s.
N pool size
(g N m-2) 0.5 ± 0.5 a 0.5 ± 0.8 a 0.6 ± 0.8 a 2.1 ± 1.7 a n.s.
35
FIGURE LEGENDS
Figure 1. Location of chronosequence sites in the Greater Yellowstone region. A) All sites; B)
Undisturbed and 30 yr TSB (time-since-beetle) sites in Yellowstone National Park; C) Red and
Gray sites in the Green River Lakes region of Bridger-Teton National Forest.
Figure 2. Understory cover by plant functional group across a chronosequence of mountain pine
beetle disturbance in lodgepole pine forest.
Figure 3. Litter variables across a chronosequence of mountain pine beetle disturbance in
lodgepole pine. A) litter depth, B) litter mass, C) fresh needle litter %N, D) all-needle litter %N,
E) total litter %N, F) litter N pool. P values are from ANOVA, and class differences were
determined using Tukey’s HSD test (α = 0.05). N = 5/class, error bars = 2 standard errors.
Values within the same panel with the same letter are not significantly different.
Figure 4. Soil N pools and fluxes across a chronosequence of mountain pine beetle
disturbance in lodgepole pine. For extractable NH4+ pool size (panel A), and extractable NO3
-
pool size (panel B), significance levels are from ANOVA and Tukey’s HSD test (α = 0.05). For
net mineralization rate (panel C), and net nitrification rate (panel D), significance levels were
determined from general linear models reported in Table 3. Summary statistics in each panel are
for the class terms in each respective mode, and values within the same panel with the same
letter are not significantly different. Error bars = 2 standard errors.
36
Figure 5. Relationships between litter depth and N content, soil temperatures, and net N
mineralization rates across a chronosequence of mountain pine beetle disturbance in
lodgepole pine. ● = undisturbed, ■ = red stage, ▲ = gray stage, ♦ = 30 year after outbreak. A)
Soil temperature vs. litter depth. Solid symbols = litter-soil interface temperatures, open symbols
= 10-cm soil depth temperatures; B) Net N mineralization vs. litter-soil interface temperature; C-
E) Net N mineralization vs. fresh needle litter %N (C), all needle litter %N (D), and total litter
%N (E); F) Net N mineralization vs. litter N pool. In panels A and B, n = 12 as only three plots
per class were instrumented with temperature probes; All plots are included in panels C-F (n =
20). Summary statistics and regression lines in each panel are for significant linear regressions.
Figure 6. Foliar %N of unattacked lodgepole pine (A) and %N of the sedge Carex geyerii (B)
vs. soil N mineralization rate in a chronosequence of mountain pine beetle disturbance in
lodgepole pine. ● = undisturbed, ■ = red stage, ▲ = gray stage, ♦ = 30 year after outbreak.
Summary statistics are from linear regression.
37
FIGURES
Figure 1
38
Figure 2
0
10
20
30
40
50
60
70
Undisturbed Red Gray 30 yr
% g
roun
d co
ver
SeedlingsForbShrubSedgeGrass
39
Figure 3
0.00.51.01.52.02.53.03.54.0
Undisturbed Red Gray 30 yrs
Litte
r dep
th (c
m)
BC
A
C
ABP =0.002
0
500
1000
1500
2000
2500
3000
Undisturbed Red Gray 30 yrs
Litte
r mas
s (g
m-2
)
P =n.s
0
5
10
15
20
25
Undisturbed Red Gray 30 yrs
Litte
r N p
ool (
g N
m-2
)
AB AB
B
AP =0.04
A)
B)
C)
0.0
0.2
0.4
0.6
0.8
1.0
Undisturbed Red Gray 30 yrs
Fres
h ne
edle
litte
r %N
B
A
B
AP < 0.0001
0.0
0.2
0.4
0.6
0.8
1.0
Undisturbed Red Gray 30 yrs
All
need
le li
tter %
N
P < 0.0001
CB
ABA
0.0
0.2
0.4
0.6
0.8
1.0
Undisturbed Red Gray 30 yrs
Tota
l litt
er %
N
P =n.s.
D)
E)
F)
40
Figure 4
0
5
10
15
20
25
30
Undisturbed Red Gray 30 yr
Net
nitr
ifica
tion
(µg
N g
soi
l-1 y
r-1)
P = 0.02D)
0
5
10
15
20
25
30
Undisturbed Red Gray 30 yr
Net
N m
iner
aliz
atio
n (µ
g N
g s
oil-1 yr
-1)
P = 0.01C)
0
1
2
3
4
5
6
7
8
Undisturbed Red Gray 30 yr
Ext
ract
able
NH4
+ (µg
N g
soi
l-1)
P = 0.05A)
B
ABAB
A
C
AB
BC
A
B
A
A
B
0
1
2
3
4
5
6
7
8
Undisturbed Red Gray 30 yrE
xtra
ctab
le N
O3- (µ
g N
g s
oil-1
)
P = n.s.B)
41
Figure 5
6
8
10
12
14
16
18
20
22
0.5 1.0 1.5 2.0 2.5 3.0Litter depth (cm)
Tem
pera
ture
(C)
Adj. R 2 = 0.31; P = 0.04
Adj. R 2 = 0.30; P = -226
10141822263034
12 14 16 18 20 22Temperature (C)
Net
N m
iner
aliz
atio
n(µ
g N
g s
oil-1
yr-1
)
Adj. R 2 = 0.39;P = 0.02
-226
10141822263034
0.2 0.4 0.6 0.8 1.0All needle litter %N
Net
N m
iner
aliz
atio
n(µ
g N
g s
oil-1
yr-1
)
P = n.s.
-226
10141822263034
0.2 0.4 0.6 0.8 1.0Fresh needle litter %N
Net
N m
iner
aliz
atio
n(µ
g N
g s
oil-1
yr-
1)
Adj. R 2 = 0.19; P = 0.03
-226
10141822263034
0.2 0.4 0.6 0.8 1.0Total Litter %N
Net
N m
iner
aliz
atio
n(µ
g N
g s
oil-1
yr-1
)
P = n.s.
-226
10141822263034
0 5 10 15 20 25 30Total litter N pool (g litter N m-2)
Net
N m
iner
aliz
atio
n(µ
g N
g s
oil-1
yr-1
)
P = n.s.
A)
D)
B)
C)
F)E)
42
Figure 6
0.75
0.85
0.95
1.05
1.15
1.25
P. c
onto
rta f
resh
folia
r % N
A) Adj. R 2 = 0.41; P = 0.003
0.75
1.00
1.25
1.50
1.75
2.00
-5 0 5 10 15 20 25 30 35
Net N mineralization (µg N g soil-1 yr-1)
C. g
eyer
i %
N
B) Adj. R 2 = 0.48 ;P = 0.007
43
CHAPTER 2
Forest type influences ecosystem response to bark beetle disturbance
ABSTRACT
Outbreaks of native Dendroctonus bark beetles are causing extensive tree mortality in
conifer forests throughout western North America, yet consequences for nitrogen (N) cycling in
different forest types are not well known. We quantified beetle-induced changes in forest
structure, soil microclimate, and N cycling through litter, soils, and vegetation in Douglas-fir
(Pseudotsuga menziesii) and lodgepole pine (Pinus contorta var. latifolia) forests of Greater
Yellowstone (WY, USA). We hypothesized N cycling responses would be greater in Douglas-fir,
which generally has higher overall N capital. Undisturbed Douglas-fir stands had larger litter and
soil N pools, and greater net N mineralization rates than lodgepole pine. However, responses to
disturbance were similar and included a pulse of N-enriched litter, a doubling of soil N
availability, a 30-50% increase in understory cover, and a 20% increase in N concentration of
composite foliage in unattacked trees. Soil temperature did not vary in Douglas-fir, but was
cooler in beetle-disturbed lodgepole pine. N concentration in fresh foliage was uncorrelated with
N mineralization in Douglas-fir, but positively correlated in lodgepole pine. Though pools and
fluxes of soil inorganic N doubled, they remained low in both forest types (< 6 µg N g soil-1
NH4+-N or NO3
--N; < 25 µg N g soil-1 yr-1 net N mineralization; < 8 µg N g soil-1 yr-1 net
nitrification). Our results suggest that bark beetle disturbance affects N cycling similarly in
Douglas-fir and lodgepole pine, and both systems appear to have mechanisms that could enhance
N retention following beetle outbreak.
44
INTRODUCTION
Insect outbreaks contribute significantly to many disturbance regimes, and can be strong
drivers of change in forest ecosystems (Haack and Byler 1993). Insect-induced mortality alters
ecosystem structure by redistributing biomass among ecosystem pools, often via the selective
mortality of mature canopy trees. In turn, vegetation-driven aspects of ecosystem function such
as primary production and elemental cycling may be affected from local to regional scales
(Schowalter 1981). Widespread outbreaks of native Dendroctonus bark beetles are occurring
within multiple conifer forest types of western North America, yet the consequences of this
disturbance type for many ecosystem functions are not well known. Furthermore, the pre-
outbreak condition and biogeochemical characteristics of infested stands can vary substantially
among forests dominated by different host tree species (Thomas and Prescott 2000, Giardina et
al. 2001, Lovett et al. 2004). In turn, ecosystems organized around such foundational species
(Ellison et al. 2005) could have unique responses to similar bark beetle disturbance (Lovett et al.
2006). This variation may arise from specific biogeochemical properties of the host (e.g. nutrient
uptake rates; nutrient allocation among tissues), host landscape position (e.g. elevation, slope,
soil type), or post-disturbance successional trajectory (Schowalter 1981).
Extensive outbreaks of the Douglas-fir beetle (DFB; Dendroctonus pseudotsugae
Hopkins) and mountain pine beetle (MPB; Dendroctonus ponderosae Hopkins) have affected
Douglas-fir (Pseudotsuga menziesii [Mirb.] Franco), and lodgepole pine (Pinus contorta Dougl.)
forests of Greater Yellowstone (Wyoming, USA) since 2003. At epidemic population levels,
these closely related Dendroctonus species utilize pheromone-mediated mass attack strategies to
overcome host trees (Raffa and Berryman 1987) and produce similar patterns of tree mortality,
litterfall, and canopy decay across host forest types. Both beetle species tend to attack large
45
canopy-dominant trees. Stand-level mortality rates frequently exceed 50% (Negron 1998,
Safranyik and Wilson 2006), leaving post-outbreak stands with decreased live stem density and
mean stem diameter (McMillin and Allen 2003, Shore et al. 2006). Foliage of beetle-killed trees
of both species turns red in the year following successful attack (red stage), and is shed to the
forest floor within two or three years (gray stage) (Negron 1998, McMillin and Allen 2003).
Canopy opening and subsequent inputs of needle litter to the forest floor induced by bark beetles
may influence soil microclimate (Griffin et al. 2011) by increasing soil insulation (Byers 1984),
incident radiation (Hais and Kucera 2008), and air flow through the stand (Boon 2009).
Despite the similar mortality, litterfall, and structural changes induced by both the MPB
and DFB in their respective host forest types, subsequent effects on nitrogen (N) cycling may
differ. Although lodgepole pine litter has a slightly higher N content than Douglas-fir, the
concentration of lignin in lodgepole pine litter is twice that of Douglas-fir, resulting in a greater
lignin:N ratio (Thomas and Prescott 2000). Lignin:N ratio is a strong control on litter
decomposition (Scott and Binkley 1997) and subsequent return of N to the soil profile (Prescott
2005). Higher decay constant (k) values of Douglas-fir litter relative to lodgepole pine support
this relationship (Keane 2008). Thus, in undisturbed systems N may be retained in the litter layer
longer in lodgepole pine forests than in Douglas-fir. Insect-induced changes in needle senescence
and nutrient content are known to influence litter decomposition rates in conifers (Koukol et al.
2008, Przybyl et al. 2008), and bark beetle mortality has been shown to increase fresh needle
litter N concentration in both ponderosa pine (Pinus ponderosae Laws.) and lodgepole pine
(Morehouse et al. 2008, Griffin et al. 2011). Furthermore, in undisturbed forests Douglas-fir
generally has greater pool sizes of soil inorganic N and greater rates of net N mineralization
relative to lodgepole pine (Thomas and Prescott 2000). Greater baseline N capital of soils is
46
correlated with accelerated N cycling between soil, foliage, and litter N pools (Prescott et al.
2000), and may increase the potential for nutrient loss following disturbance. Laboratory studies
show uptake rates of inorganic N in Douglas-fir roots are 73% greater for NO3- and 35% greater
for NH4+ than in lodgepole pine (Hawkins et al. 2008). These differences suggest that declines in
stand-level nutrient uptake following bark beetle outbreak may be greater in Douglas-fir forests
relative to lodgepole pine, leading to greater accumulation of inorganic N in soils and thus a
greater potential for N leaching. Differences in N uptake rates may also contribute to variable
responses of foliar N content in unattacked or surviving trees following disturbance.
In this study we address two overarching questions: (1) How does DFB disturbance alter
forest structure, soil microclimate, and N cycling through litter, soils, and vegetation in
Douglas-fir forests? (2) Are the patterns of change observed in beetle-disturbed Douglas-fir
forests similar to those in mountain pine beetle-disturbed lodgepole pine forests? Forest
structure, soil microclimate, and nitrogen pools and fluxes were measured in both undisturbed
and gray stage beetle-killed forests. Hypothesized bark beetle effects and suggested mechanisms
are summarized in Figure 1. We expected beetle disturbance to decrease live canopy biomass
and initiate a pulse of N-enriched litter to the forest floor. Shading by litter and growth of
understory vegetation was expected to cool soil temperatures, while soil N cycling rates and
pools of inorganic N were predicted to increase resulting in increased foliar %N of surviving
unattacked trees. When comparing forest types, we expected the absolute magnitude of increases
in soil N pools and fluxes to be larger in Douglas-fir. However, proportional increases were
predicted to be larger in lodgepole pine and therefore potentially more ecologically significant.
Foliar N response to changing soil N availability could also vary among forest type, with more
N-limited stands showing larger increases in foliar N. Bark beetle disturbance may also influence
47
the cycling of micronutrients in conifer forests, which can be important controls on
macronutrient cycles.
METHODS
Lodgepole pine study area, site characteristics, and data used for our comparisons of
forest type were reported in Griffin et al. (2011), with the exception of foliar micronutrients.
Vegetation, litter, and soil sampling methods in Douglas-fir were identical to those used by
Griffin et al. (2011) in mountain pine beetle-disturbed lodgepole pine stands.
Douglas-fir study area and sampling design
Douglas-fir study sites were located in the Greater Yellowstone Ecosystem of
northwestern Wyoming, USA, within Grand Teton National Park and Bridger-Teton National
Forest in the Moran Junction/Buffalo River Valley Region. July temperatures average 14.9 ºC,
January temperatures average -11.2 ºC, with 595 mm of annual precipitation occurring mostly as
snow (WRCC 2010c). The most recent outbreak of DFB in this area began in the early 2000’s,
though the region has experienced widespread outbreak of multiple bark beetle species in recent
years including the mountain pine beetle in lodgepole pine and whitebark pine (Pinus albicaulis
Engelmann), and the spruce beetle (Dendroctonus rufipennis Kirby) in Engelmann spruce (Picea
engelmannii Parry) (USFS 2006). In 2008 we sampled ten spatially independent (>2 km apart)
Douglas-fir stands (five undisturbed, five attacked by bark beetles in 2003 and 2004). DFB-
disturbed stands were identified using aerial surveys of insect damage (USFS 2006), and field
inspection assured selection of stands with comparable structure, basal area, soils, and
disturbance intensity. Within each stand, an 8m radius (201 m2) plot was installed with GPS
location, elevation, and slope recorded at the plot center.
48
Vegetation
All live and dead trees greater than breast height (1.4m) were identified to species and
measured for diameter at breast height (DBH). Beetle-killed trees were identified by the presence
of exit holes on the bark exterior and distinctive Dendroctonus galleries underneath the bark. For
the beetle-disturbed sites, pre-outbreak basal area was calculated by summing the basal area of
live and beetle-killed trees. Mortality rates in disturbed plots were as high as 100% of canopy
trees, and survivors were mostly smaller sub-canopy trees contributing little basal area. Thus, the
post-disturbance growth (~4yrs worth) of survivors was not considered in calculating the pre-
disturbance basal area of beetle-killed stands. Canopy biomass was determined for each tree
using allometrics developed for Rocky Mountain conifers (Brown 1978) and summed by plot.
Ground cover was visually estimated to the nearest 10% by plant functional groups (forbs,
sedges, grasses, shrubs, and tree seedlings) in ten 0.25-m2 circular microplots; total cover was
allowed to exceed 100% to account for multiple strata of ground vegetation. The microplots were
located within the inner 5-m plot radius using a stratified random design of fixed distances (one
at 0.5-m; two at 1.5-m, 2.5-m and 3.5-m; and three at 4.5-m) and random bearings (in 10º
increments) from the plot center, and averaged by plot.
Microclimate
To evaluate variation in growing-season soil microclimate, soil temperature was
measured hourly in three of the five plots within each disturbance class (N = 6 instrumented
plots) from Ju1y 4 through August 31, 2008, using three pairs of iButton datalogger probes
(Maxim Integrated Products Inc., Dallas Semiconductor, Sunnyvale, CA) per plot. One iButton
of each pair was installed at the litter-organic soil interface, and the second was installed 10 cm
below this interface. For each soil depth, temperature probe data were summarized as follows.
49
First, plot-level hourly temperatures were determined by averaging data from the three probes
per plot. Plot-level daily mean temperatures, ranges, maximums, and minimums were then
calculated from the hourly data, and averaged to determine disturbance class-level daily means.
Plot-level growing season values were obtained by averaging daily values across the sampling
period, and disturbance class-level growing season means were then calculated from these plot-
level data.
Litter quantity and quality
In each 0.25-m2 microplot, litter depth was recorded at three locations and a 400-cm2
sample of the litter layer was collected and oven-dried at 60 ºC. Plot-level litter depth and mass
were obtained by averaging values from the 10 microplots. Litter from each microplot was sorted
into three categories: fresh current-year needle litter, identified by bright red color and lack of
mottling on surface (Morehouse et al. 2008); all-needle litter (all ages of needles combined); and
total litter (all foliar litter components and woody litter components < 1.0 cm wide). Sorted litter
was composited by plot for each category, and ground to powder for C:N analysis on a Leco
CNS-2000 at the University of Wisconsin Soil and Plant Analysis Laboratory (UWSPAL 2010).
Soil chemistry, N pools, and N fluxes
One soil core was collected from each 0.25-m2 microplot (n = 10 per plot) using a 5-cm
diameter x 15-cm long PVC corer. Soils were sieved (2 mm), weighed, and divided into three
subsamples: 30 g oven-dried at 60 ºC for gravimetric percent moisture; 20 g extracted in 75 ml of
2M KCl for 2 hours, with the extract then filtered and frozen for later analysis of NH4+ and NO3
-
pools; and 20 g air-dried and bulked by plot for soil texture and chemical analyses. Air-dried soil
was analyzed for pH, total N, exchangeable Ca, Mg, and K, available P (Bray P1 extract), and
organic matter at the University of Wisconsin Soil and Plant Analysis Laboratory (UWSPAL
50
2010). Soil organic N was determined by difference using total N and inorganic N values, and
soil texture was determined using the Bouyoucos hydrometer technique (Bouyoucos 1962).
Net N mineralization and net nitrification were measured using ion-exchange resin cores
(Binkley et al. 1992), with one core incubated in situ for approximately one year (July 2008 -
June 2009) in each 0.25-m2 microplot. Incubated cores consisted of a 5-cm diameter x 15-cm
long PVC tube of soil with an ion-exchange resin bag placed at the bottom. Resin bags were
constructed using 20 g of mixed bed ion exchange resin (J.T Baker #JT4631-1) tied inside a
piece of un-dyed nylon stocking material. Upon retrieval in summer 2009, core soils were sieved
(2 mm) and core soils and resin bags were extracted separately in 2M KCl in the same manner as
the 2008 initial soil samples described above. All KCl extractions were analyzed for [NH4+] and
[NO3-] using colorimetric methods on an Astoria Pacific II continuous flow autoanalyzer. For
each microplot, mineralization and net nitrification rates were calculated as:
rate = [(final soil N + resin bag N) – initial soil N ] / incubation time
and expressed in units of µg N g soil-1 year-1. Values were then averaged by plot. Atmospheric N
inputs in the region are low (~ 1 kg N ha-1 year-1; www.epa.gov/castnet/charts/YEL408totn.png)
and were not factored into the soil N calculations.
Foliar Chemistry
In each plot, pole pruners were used to collect canopy foliage from three live trees
unattacked by bark beetles. Two fully sunlit branches (0.5-m long) were clipped from each tree
and separated into two subsamples: current-year foliage (from one branch per tree), and all-years
foliage (from the second branch per tree). Foliar samples were oven-dried at 60ºC and ground to
powder for chemical analyses at the University of Wisconsin Soil and Plant Analysis Lab. C and
N analysis for both current-year fresh foliage and all-year composite foliage was performed on a
51
Leco CNS-2000 analyzer. In current year fresh foliage, P, K, Ca, Mg, S, Zn, B, Mn, Fe, Cu, Al,
and Na were determined by ICP-OES (Thermo Jarrell Ash IRIS Advantage Inductively Coupled
Plasma Optical Emission Spectrometry) and ICPMS (VG Plasma Quad PQ2 Turbo Plus
Inductively Coupled Plasma Mass Spectrometry) (UWSPAL 2010). Tree-level canopy data (N=3
per sample type per plot) were averaged to obtain plot-level data. To compute canopy N pool
sizes, the mean foliar N concentration for each plot was multiplied by biomass values derived
from the allometrics cited above.
Statistical analyses
All statistical analyses were performed in SAS (SAS Institute Inc. 2003), and unless
otherwise noted all reported variance values are two standard errors. All variables were checked
for normality and transformed if necessary to satisfy the assumptions of statistical methods. N =
10 (5 per disturbance class) for all analyses of Douglas-fir forests except those using
microclimate data, where N = 6 as only three of five plots per disturbance class were
instrumented. For all analyses comparing forest types, N = 20.
General linear models were used to test for differences between both undisturbed and
beetle-killed Douglas-fir forests, as well as undisturbed Douglas-fir and undisturbed lodgepole
pine, in basal areas, site characteristics (slope, elevation, aspect, and soil chemistry), understory
cover, soil temperatures, litter quantity and quality, soil N pools, foliar N concentrations, and
foliar N pools. We also used general linear models to account for potential covariate effects of
soil pH, bulk density, and C:N ratio in addition to disturbance class when testing for differences
in net mineralization and net nitrification rates in Douglas-fir sites. Relationships between soil N
availability and fresh foliar N concentrations were explored using Pearson correlation and linear
regression. Linear regression was also used to test whether abiotic changes in soil temperature or
52
biotic changes in litter N were more closely related to changes in soil N cycling. To test whether
litter N, soil N, and foliar chemistry of lodgepole pine and Douglas-fir forests responded
differently to bark beetle disturbance, we used general linear models of the form: y = species +
disturbance class + species*disturbance class. A significant (P ≤ 0.05) interaction term in these
analyses indicates differing response to beetle disturbance by type.
RESULTS
Douglas-fir forests
Site characteristics, stand structure, and soil microclimate. Undisturbed and beetle-
killed Douglas-fir stands occupied similar landscape positions, with no difference in either slope
or aspect between disturbance classes (Table 1). However, due to the spatial distribution of the
Douglas-fir beetle outbreak, beetle-killed sites were on average 140 m higher in elevation than
undisturbed (2186 vs. 2325 m; P = 0.0400; Table 1). Soil texture (%sand, silt, clay), bulk
density, and organic soil horizon depth were similar between disturbance classes, as were soil
Ca, Mg, K, P, and pH (Table 1). Douglas-fir basal area comprised 97% (67 of 69 m2 ha-1) of total
stand basal area among all sites and did not differ between undisturbed and beetle-killed sites
(Table 1). Live Douglas-fir basal area averaged 98% of the total Douglas-fir basal area in the
undisturbed class and 7% in beetle-killed stands (Table 1). In beetle-killed stands, Douglas-fir
mortality averaged 93% of stem density, with an estimated concurrent decline in live canopy
biomass averaging 15,796 ± 1704 kg ha-1 (Table 1). Total biotic understory cover increased
significantly from 79 to 103%, driven by significant increases in forb (19 vs. 36 %) and grass (18
vs 49 %) cover (P ≤ 0.05; Figure 2). Growing season soil temperatures were not significantly
different between Douglas-fir disturbance classes at either soil depth, averaging 13.1 ± 1.0 ºC at
53
the litter-soil interface and 10.9 ± 0.9º C at 10 cm depth below the litter layer (Table 1).
Variability in soil temperature was also consistent between disturbance classes at both depths,
with no differences in daily temperature means, ranges, minimums, or maximums (Table 1).
Litter quantity and quality. Litter quantity did not differ between undisturbed and beetle-
killed Douglas-fir stands; litter depth averaged 3.2 ± 0.4 cm, and total litter mass averaged 1842
± 247 g m-2 (Table 1). However, litter quality did differ between disturbance classes. The N
concentration in needle litter of all ages increased by 55% in beetle killed stands (from 0.89 to
1.38 %N), and total litter layer N concentration increased by 25% (from 1.10 to 1.37 %N; Table
2). Current year needle litter N concentration showed a marginally significant 13% increase
(from 0.70 to 0.79, P = 0.0641; Table 2). Increases in litter %N resulted in significant concurrent
declines in litter C:N ratio in all three litter categories. However, total litter N pool size (product
of total litter mass and total litter %N) showed only a marginally significant increase (P =
0.0666) from 19.3 to 26.4 g N m-2 (Table 2), as litter mass varied considerably among sites.
Soil N pools and fluxes. Soil inorganic N pools and soil N transformation rates differed
between undisturbed and beetle-killed Douglas-fir stands (Table 2). Extractable ammonium in
beetle-disturbed stands was twice that of undisturbed stands (5.0 ± 0.8 vs. 2.5 ± 0.6 µg N g soil-
1), extractable nitrate was almost four times greater (5.5 ± 5.2 vs. 1.5 ± 1.0 µg N g soil-1), and
total extractable inorganic N (NH4+ + NO3
-) in beetle-disturbed stands was more than double that
in undisturbed (10.5 ± 4.7 vs. 4.0 ± 1.7µg N g soil-1). Annual net N mineralization rates were
highly variable and were not significantly different between disturbance classes, though a
marginally significant increase (P = 0.0611) was found when using soil bulk density as a
covariate, with higher rates observed in beetle-killed stands. Annual net nitrification rates
increased 50% in beetle-disturbed stands relative to undisturbed (7.2 ± 1.0 and 4.8 ± 0.6 µg N g
54
soil-1 yr-1 respectively). Neither net N mineralization nor net nitrification rates were significantly
correlated with any metric of soil temperature (Pearson P < 0.05; N =6), nor was net N
mineralization correlated with any metric of litter N (Pearson P < 0.05; N = 10). Nitrification
fraction (the ratio of net nitrification to net mineralization) averaged 0.38 ± 0.10 and did not
differ among disturbance classes, nor did soil C:N, soil %N, or soil N pool size (Table 2).
Foliar N. Beetle disturbance impacted both foliar N concentration and foliar N pool size.
N concentration in the fresh foliage (current year) of unattacked Douglas-fir trees averaged 1.28
± 0.11% and did not differ between disturbance classes. However, N concentration in all foliage
of unattacked Douglas-fir was 18% greater in beetle-disturbed stands than in undisturbed (0.90 ±
0.07 vs. 0.76 ± 0.04 %N), with a concurrent decrease in C:N ratio (Table 2). Foliar N pool size
mirrored changes in live foliar biomass and declined sharply by 79% from 11.5 ± 1.7 to 2.4 ± 1.5
g N m-2 (Table 2). Neither fresh foliar %N or C:N were significantly correlated with any metric
of inorganic soil N, including NH4+, NO3
-, total inorganic soil N, net mineralization, net
nitrification, or soil C:N (Pearson correlation; data not shown). However, foliar %N of all
needles was positively related to both soil NO3- (linear regression P = 0.0553, F = 5.03, Adj. R2
= 0.31) and total inorganic soil N (linear regression P = 0.0132, F = 10.06, Adj. R2 = 0.50), and
the C:N ratio of all needles was negatively related to total inorganic soil N (linear regression P =
0.0253, F = 7.53, Adj. R2 = 0.42).
Forest type comparison
Landscape position (slope and aspect) did not differ among Douglas-fir and lodgepole
pine sites (P > 0.08) though Douglas-fir sites were approximately 200m lower in elevation (2255
± 71m vs. 2438 ± 35m; P < 0.0001; Douglas-fir data in Table 1, lodgepole pine data reported in
Griffin et al. (2011)), consistent with the species’ distribution in the region (Despain 1990).
55
Douglas-fir soils were higher in pH and contained less sand and more silt than lodgepole pine
soils, as well as greater concentrations of organic matter, Ca, Mg, K, and P (P < 0.01; Douglas-
fir data in Table 1, lodgepole pine data reported in Griffin et al. (2011)). Compared to lodgepole
pine, Douglas-fir sites contained higher total basal area (all sites: 69.0 ± 8.5 vs. 42.7 ± 7.6; P =
0.0003; Figure 2), beetle-killed basal area (beetle-killed sites: 69.6 ± 11.2 vs. 13.7 ± 5.8; P <
0.0001; Figure 2), and percent mortality of pre-outbreak living stems <1.4m DBH (93 ± 3% vs.
52 ± 17%).
Undisturbed forests. In the absence of bark beetles, forest types differed in understory
cover composition and soil temperature. Total biotic cover was greater in Douglas-fir (79 ± 6%
vs. 29 ± 10 %; P < 0.0001), driven by greater cover of shrubs, total graminoids (grasses +
sedges), and tree seedlings compared to lodgepole pine (P < 0.02; Figure 2). Soil temperatures at
the litter-soil interface in undisturbed forests were cooler and had a lower range in Douglas-fir
(temperature: 13.5 ± 0.7 ºC vs. 15.4 ± 1.0 ºC; temperature range: 8.2 ± 1.2 ºC vs. 27.4 ± 2.9 ºC; P
≤ 0.03; Douglas-fir data in Table 1, lodgepole pine data reported in Griffin et al. (2011)).
Temperatures at 10-cm soil depth in undisturbed forests were not significantly different among
forest types, and averaged 10.9 ± 0.6 ºC. However, the range of 10-cm depth temperature was
lower in Douglas-fir (1.4 ± 0.6 ºC vs. 2.4 ± 0.2 ºC; P = 0.0236; Douglas-fir data in Table 1,
lodgepole pine data reported in Griffin et al. (2011)).
Undisturbed Douglas-fir and lodgepole pine forests also differed in several metrics of
litter, soil, and foliar N. Litter depth and all measurements of litter N content and pool size were
greater in Douglas-fir (Figure 3; P ≤ 0.03), though there was no difference in total litter mass. In
soils, extractable NH4+, NO3, total inorganic N, and total organic N were also greater in Douglas-
fir (P ≤ 0.005; Figure 4). Net N mineralization rates were not statistically different, however net
56
nitrification rates in undisturbed Douglas-fir were twice that of undisturbed lodgepole pine (4.8
vs 2.4 µg N g-1 day-1; Figure 4; P = 0.0242). Composite foliage in undisturbed forests did not
differ in foliar biomass, N content, or N pool size (Figure 5), however fresh foliar %N was
greater in Douglas-fir (P = 0.0185; Figure 5). P, K, Ca, and B in fresh foliage were also greater
in Douglas-fir, while Mn and Al were greater in lodgepole pine (P ≤ 0.03; Table 3).
Ecosystem response to bark beetle disturbance. Despite differing pre-disturbance
communities, the response of understory cover to bark beetle disturbance was similar between
forest types. Significant disturbance class effects were found for increased grass and forb cover
(P ≤ 0.04; Figure 2), but there were no significant interactions between species and disturbance
class for any biotic cover variable (Figure 2). However, soil temperature response did vary
among forest type. Significant disturbance class effects were found for decreases in the mean,
range, minimum, and maximum temperature at the litter-soil interface, though significant
interactions between species and disturbance class for the range and maximum show that
decreases in these variables were limited to lodgepole pine (P < 0.04; Douglas-fir data in Table
1; lodgepole pine data reported in Griffin et al. (2011)). No significant disturbance class or
species*disturbance class effects were found for 10-cm depth soil temperatures.
Outbreak-induced changes in N cycling also varied among forest types. Significant
effects of disturbance class were found for all metrics of litter quality and quantity except total
litter mass, which did not differ by either species or disturbance class (Figure 3). Nitrogen
concentration in both fresh and total litter also showed a significant interaction between species
and disturbance class, indicating that fresh litter %N increased only in lodgepole pine and total
litter %N increased only in Douglas-fir (Figure 3). In soils, significant disturbance class effects
were found for extractable NH4+ and NO3
-, net N mineralization, and net nitrification. No
57
significant effects were found for nitrification fraction, and there were no significant interactions
between species and disturbance class for any soil N metric (Figure 4). In trees unattacked by
bark beetles, all foliar N metrics showed a significant disturbance class effect except for fresh
foliar %N, which was marginally significant (Figure 5). N concentration of unattacked current
year foliage was higher in beetle-killed lodgepole pine, and composite foliar N was higher in
beetle-killed stands of both species. Live canopy biomass and canopy N pool size were lower in
beetle-killed stands of both types, but also showed a significant interaction between species and
disturbance class indicating that declines in these variables were greater in Douglas-fir compared
to lodgepole pine (Figure 5). Fresh foliar %N was positively related with net N mineralization in
soils of lodgepole pine (Adj. R2 = 0.56; P = 0.0202) but not in Douglas-fir.
Foliar micronutrients
Bark beetle disturbance significantly lowered foliar concentrations of Ca, Mg, and Mn in
fresh foliage (needles < 1yr) of unattacked Douglas-fir trees, while foliar concentrations of P, K,
Mg, S, Zn, B, Fe, Cu, Al and Na did not differ between Douglas-fir disturbance classes (Table
3). Foliar Mn also declined substantially in beetle-killed lodgepole pine forests, where decreases
in Al were also significant (Table 3). Forest types differed overall in foliar P, K, Ca, and S with
higher concentrations in Douglas-fir. Mn, Cu, and Al also differed by type, with greater
concentrations in lodgepole pine (Table 3). Across both forest types there was a significant
disturbance class effect for P, Mn and Al, with P increasing and Mn and Al decreasing in beetle-
killed stands relative to undisturbed (Table 3).
58
DISCUSSION
Bark beetle disturbance initiated similar changes in N cycling through litter and soil in both
Douglas-fir and lodgepole pine forests, despite substantial differences in pre-disturbance forest
structure and ecosystem N dynamics. Patterns of increased litter depth, needle litter N
concentration, and litter N pool size were consistent across forest types, though concurrent
increases in total litter %N were limited to Douglas-fir. In soils, patterns of increased extractable
inorganic N and N transformation rates following bark beetle disturbance were proportionally
similar among forest types, though Douglas-fir had more N-rich soils overall. Live foliar N pools
declined sharply with beetle-killed basal area in both forest types. Differences between these
foundation species (Ellison et al. 2005) in pre-disturbance ecosystem characteristics may explain
variable responses for some metrics following beetle outbreak. Total inorganic N in soils was
four-fold higher in Douglas-fir, and foliar N in undisturbed Douglas-fir averaged 1.23%
compared to 0.82% in lodgepole pine. These data suggest a higher overall N capital in Douglas-
fir forests and less potential N limitation. Thus, the response of foliar %N in unattacked trees to
increased soil N in beetle-killed stands was limited to lodgepole pine forests.
Forest types also differed in the response of abiotic soil conditions. Soil temperature was
cooler in beetle-killed lodgepole pine forests, presumably due to insulating effects of litter input
and increases in soil moisture. This effect was not seen in Douglas-fir, where the shading effect
of abundant understory cover may be the dominant control on soil temperature. Soil temperature
has been negatively correlated with net N mineralization across longer time scales of TSB in
lodgepole pine (Griffin et al. 2011), however neither forest type showed this relationship using
data from undisturbed and gray-stage TSB classes only. Furthermore, no litter N metrics were
correlated with net N mineralization in either forest type at this time-scale. Though unmeasured
59
in this study, soil moisture and soil C dynamics are also potential controls on soil N processes
and would be expected to change following bark beetle disturbance (Morehouse et al. 2008,
Spielvogel et al. 2009).
Observed changes in litter and soil N cycling of both forest types were consistent with
mechanisms of disturbance specific to the bark beetle insect guild. Rapid tree death appears to
inhibit the resorbtion of N from foliage prior to needlefall (Morehouse et al. 2008, Griffin et al.
2011), and deposits a large pulse of needle litter in a relatively short time. Though dead conifer
needles may leach some N compounds before dropping (Stadler et al. 2005), bark beetles do not
alter fluxes of inorganic N to soils in the same way as the frass of conifer defoliators (le Mellec
et al. 2009, Pitman et al. 2010). Beetle-induced inputs of N to the soil surface are largely in the
form of organic N in litter, which can be a N sink during a period of net N immobilization in the
early stages of decomposition (Fahey et al. 1985, Remsburg and Turner 2006). Thus,
disturbance-induced fluxes of N from canopy to soil pools may be spread over longer time
periods following bark beetle outbreak compared to defoliators. Minimal inputs of inorganic N to
the soil profile could explain why the magnitude of observed increases in net N mineralization
following bark beetle disturbance in lodgepole pine was smaller than those following defoliator
disturbance in another Pinus forest (Tokuchi et al. 2004).
Effects of bark beetle disturbance on micronutrient cycling could also be important for
post-beetle N cycling. Beetle-induced tree death reduces the delivery of photosynthate C to
ectomycorrhizal fungi. In lodgepole pine stands of the Yellowstone region, a similar loss of C
delivery to ectomycorrhizal communities following 50% defoliation has been shown to double
the level of manganese peroxidase activity in soils (Cullings et al. 2008), an extracellular fungal
enzyme used to metabolize lignin-based C sources. Greater demand for Mn to support increases
60
in this enzymatic pathway could reduce soil Mn availability to plants and may explain the
considerable declines in foliar Mn levels of unattacked, surviving trees in beetle-disturbed stands
of both forest types. In turn, litter produced from this foliage will also be lower in Mn content,
which is known to slow decomposition rates (Berg et al. 2007) and thus subsequent return of N
to soils from the litter layer. Furthermore, increased Mn peroxidase activity is likely to continue
for considerable time periods. Canopy foliage (an index of photosynthate C delivery to soil
fungi) in 30 yr post MBP lodgepole pine forest remains approximately half the biomass of
undisturbed forests (Griffin et al. 2011), and surviving trees may reduce C allocation to fungi
under conditions of elevated soil N (Hogberg et al. 2010). Concurrently, input of lignin and
woody C to soils increases with root turnover, more decomposing needle litter, and eventually
fine and coarse wood which has been shown to sequester large amounts of Mn during
decomposition (Daniel et al. 1997). Sustained declines in photosynthate C supply and increases
in lignin and woody C inputs could thus result in long-term declines of plant-available Mn.
Disturbance can lead to nutrient loss from forest ecosystems (Bormann and Likens 1979,
Chapin et al. 2004), however our data suggest three mechanisms of ecosystem N retention
following bark beetle outbreak. (1) Net N mineralization, net nitrification, and pool sizes of the
highly leachable NO3- ion remained low relative to other disturbance types in conifer ecosystems
(Turner et al. 2007, Tan et al. 2008), and nitrification fraction was unaffected. (2) Although the
foliar N pool declined in both forest types, increased foliar N of surviving trees partially
compensated for N lost in litterfall. (3) Declines in foliar Mn could affect future litter quality,
decreasing decomposition rate and slowing the return of inorganic N to the soil. Together, these
suggest that the potential for N loss as NO3- following bark beetle disturbance is low.
61
CONCLUSIONS
Forest types dominated by different foundation species (Ellison et al. 2005) can have
unique responses to a common disturbance type. N cycling was significantly impacted by bark
beetle outbreak in both Douglas-fir and lodgepole pine forests, which showed qualitatively
similar responses of increased litter N inputs and soil N processes. However, differences
between forest types in pre-disturbance ecosystem structure and N status resulted in variable
responses of soil temperature and foliar %N of unattacked trees. Though the magnitude of beetle
effects was relatively minor compared to other disturbance types, ecosystem impacts were
specific to disturbance by the bark beetle insect guild. Several mechanisms of ecosystem N
retention suggest that N losses following this native disturbance may be low.
ACKNOWLEDGEMENTS
We thank Gary Lovett, Bill Romme, and Dan Donato for constructive comments that
improved this manuscript. We also thank Emily Stanley, James Thoyre, and the Center for
Limnology at UW for access to laboratory equipment. For help in the field and the laboratory,
we appreciate the assistance of many University of Wisconsin students: Heather Lumpkin, Erin
Mellenthin, Alex Rahmlow, Amanda Rudie, and Ben Ruh.
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66
TABLES
Table 1. Site characteristics in undisturbed and bark beetle-killed Douglas-fir forests. N = 5 /
disturbance class. Significant (α < 0.05 ) ANOVA P values are shown in bold.
Douglas-fir disturbance class Site characteristic Undisturbed Beetle-killed ANOVA P Topography Elevation (m) 2186 ± 38 2325 ± 107 0.0400 Aspect (SW index) 0.367 ± 0.691 -0.085 ± 0.566 0.9201 Slope 26 ± 12 22 ± 11 0.6282 Canopy vegetation Basal area (m2 ha-1) Total 62.7 ± 9.9 75.3 ± 12.2 0.1468 Douglas-fir 58.7 ± 12.6 75.1 ± 12.1 0.0973 Live Douglas-fir 57.6 ± 13.2 5.2 ± 2.9 0.0001 Dead Douglas-fir 1.1 ± 1.2 69.9 ± 11.2 <0.0001 Live foliar biomass (kg ha-1) 15,179 ± 4021 2551 ± 1538 0.0004 Soils Growing season temperature Litter-soil interface (ºC) mean 13.5 ± 0.7 12.7 ± 2.0 0.5073 range 8.2 ± 1.2 9.4 ± 2.8 0.4607 maximum 18.1 ± 1.0 18.3 ± 3.7 0.9094 minimum 9.9 ± 0.8 8.9 ± 1.0 0.1953 10 cm soil depth (ºC) mean 11.2 ± 1.1 10.5 ± 1.6 0.4899 range 1.4 ± 0.6 1.5 ± 0.3 0.7494 maximum 11.9 ± 1.3 11.3 ± 1.7 0.5892 minimum 10.5 ± 0.8 9.8 ± 1.5 0.4284 Texture & Structure Sand (%) 56 ± 3 49 ± 3 0.1590 Silt (%) 33 ± 6 36 ± 5 0.3570 Clay (%) 12 ± 2 15 ± 2 0.0847 Bulk density (g cm-3) 0.66 ± 0.22 0.63 ± 0.10 0.8195 Organic soil depth (cm) 4.2 ± 1.5 5.5 ± 2.0 0.3078 Organic matter (%) 8.4 ± 4.8 8.6 ± 2.2 0.9299 Cations pH 6.2 ± 0.3 6.4 ± 0.2 0.2073 Ca (µg g-1; exch.) 2401 ± 808 2702 ± 253 0.4974 Mg (µg g-1; exch.) 187 ± 52 220 ± 14 0.2527 K (µg g-1; exch.) 225 ± 79 218 ± 42 0.8696 P (µg g-1; exch.) 49 ± 8.4 51 ± 15 0.7950 Litter Depth (cm) 3.0 ± 0.5 3.5 ± 0.6 0.2968 Mass (g m-2) 1760 ± 365 1924 ± 357 0.5385
67
Table 2. N concentrations, pools, and fluxes in undisturbed and bark beetle-killed Douglas-fir
forests. N = 5 / disturbance class; values in table are untransformed mean ± 2SE. Significant (α <
0.05 ) ANOVA P values are shown in bold; asterisk (*) denotes marginally significant (0.05 < α
< 0.07) ANOVA P.
1 µg N g soil-1 2 µg N g soil-1 yr-1 3 ANOVA performed on log(x) transformed data
4 ANOVA performed on 1/x transformed data
Douglas-fir disturbance class N concentration, pool, or flux Undisturbed Beetle-killed ANOVA P Litter Total litter N pool (g N m-2) 19.3 ± 3.3 26.4 ± 5.8 0.0666* Total litter %N 1.10 ± 0.06 1.37 ± 0.13 0.0054 Total litter C:N 44 ± 2 35 ± 3 0.0008 All needles %N 0.89 ± 0.07 1.38 ± 0.10 <0.0001 All needles C:N 57 ± 5 36 ± 2 <0.0001 < 1 yr old needles %N 0.70 ± 0.06 0.79 ± 0.07 0.0641* < 1 yr old needles C:N 72 ± 6 62 ± 5 0.0266
Soil Total soil N pool (g N m-2) 192 ± 78 237 ± 42 0.3353 %N 0.23 ± 0.15 0.26 ± 0.07 0.7782 C:N 19.2 ± 2.9 16.7 ± 0.4 0.11094 Organic N (mg N g-1) 2.34 ± 1.52 2.58 ± 0.70 0.7836 NH4
+ 1 2.5 ± 0.6 5.0 ± 0.8 0.0010 NO3
- 1 1.5 ± 1.0 5.5 ± 5.2 0.0520*3 Total inorganic N 1 4.0 ± 1.7 10.5 ± 4.7 0.00603 Net N mineralization 2 14.3 ± 9.3 24.8 ± 11.2 0.09503 Net nitrification 2 4.8 ± 0.6 7.2 ± 1.0 0.0295 Nitrification fraction 0.42 ± 0.18 0.34 ± 0.11 0.4724 Canopy foliage Foliar N pool (g N m-2) 11.5 ± 1.7 2.4 ± 1.5 0.0004 All needle %N 0.76 ± 0.04 0.90 ± 0.07 0.0101 All needle C:N 63 ± 3 54 ± 4 0.0046 <1yr needle %N 1.23 ± 0.19 1.34 ± 0.12 0.3853 <1yr needle C:N 40 ± 7 36 ± 3 0.3499
68
Table 3. Foliar chemistry of new needles from unattacked trees in undisturbed and beetle-killed stands of Douglas-fir and lodgepole
pine in the Greater Yellowstone Ecosystem. N = 5 stands per disturbance class; values in table are untransformed mean ± 2SE.
Significant (α < 0.05 ) ANOVA P values are shown in bold; asterisk (*) denotes marginally significant (0.05 < α < 0.06) ANOVA P.
1 ANOVA performed on log10(x) transformed data 2 ANOVA performed on x-2 transformed data 3 ANOVA performed on √x transformed data 4 ANOVA performed on 1/x transformed data 5 ANOVA performed on 1/x2 transformed data
Douglas-fir Lodgepole pine Species Comparison ANOVA P Foliar Nutrient Undisturbed Beetle-killed ANOVA P Undisturbed Beetle-killed ANOVA P Species Class Species*Class N (%) 1.23 ± 0.19 1.34 ± 0.12 0.3853 0.82 ± 0.03 1.00 ± 0.06 0.0048 <0.0001 0.0578* 0.6043 C:N 39.7 ± 7.1 35.8 ± 3.3 0.3499 59.1 ± 3.9 49.5 ± 2.4 0.0044 < 0.0001 0.0154 0.2613 N:P 5.5 ± 0.4 5.3 ± 0.4 0.5771 6.1 ± 0.2 6.4 ± 0.5 0.4200 0.0010 0.7615 0.3123 P (%) 0.23 ± 0.03 0.25 ± 0.03 0.2099 0.14 ± 0.00 0.16 ± 0.01 0.0555* < 0.0001 0.0426 0.9067 1 K (%) 1.2 ± 0.1 1.1 ± 0.1 0.4180 0.50 ± 0.04 0.58 ± 0.04 0.5153 < 0.0001 0.3032 0.9785 1 Ca (%) 0.40 ± 0.05 0.31 ± 0.04 0.0236 0.14 ± 0.04 0.15 ± 0.04 0.8026 < 0.0001 0.1104 0.0565* Mg (%) 0.11 ± 0.00 0.10 ± 0.01 0.0344 0.09 ± 0.01 0.10 ± 0.01 0.4742 0.2624 0.9954 0.1255 S (%) 0.09 ± 0.01 0.10 ± 0.01 0.7073 0.07 ± 0.00 0.08 ± 0.01 0.8023 0.0011 0.6660 0.8831 Zn (ppm) 28.5 ± 4.3 27.5 ± 4.0 0.7352 31.9 ± 1.5 31.5 ± 9.4 0.4993 5 0.2523 0.6608 0.9437 1 B (ppm) 13.6 ± 0.7 13.7 ± 2.8 0.9178 12.0 ± 0.9 11.9 ± 3.1 0.9644 0.1836 0.9822 0.9203 Mn (ppm) 132 ± 45 67 ± 22 0.0331 340 ± 18 200 ± 34 0.0010 < 0.0001 < 0.0001 0.0537* Fe (ppm) 44.4 ± 28.4 33.7 ± 7.9 0.5810 4 32.0 ± 2.4 29.1 ± 3.5 0.2780 0.3743 0.4461 0.6734 2 Cu (ppm) 7.3 ± 3.1 7.7 ± 2.5 0.8277 2.3 ± 0.6 3.1 ± 0.8 0.1714 0.0001 0.2954 0.5322 1 Al (ppm) 52.0 ± 41.9 18.8 ± 8.4 0.0780 1 492 ± 114 329 ± 48 0.0217 < 0.0001 0.0048 0.4533 3 Na (ppm) 12.4 ± 4.5 11.2 ± 1.4 0.6255 8.7 ± 3.2 11.3 ± 3.4 0.3455 0.3085 0.6951 0.2966
68
69
FIGURE LEGENDS
Figure 1. Hypothesized impacts of bark beetle disturbance in Douglas-fir and lodgepole
pine ecosystems. Bark beetle-induced tree mortality is predicted to increase litter N inputs and
moderate soil temperature. Both of these changes may accelerate fluxes of inorganic soil N and
increase pools of plant-available N. Unattacked trees within beetle-killed stands may respond to
these increases by allocating more N to foliage. Forest types are expected to differ in pre-
disturbance N pool sizes and flux rates, which may result in variable ecosystem responses.
Figure 2. Tree basal area (A) and understory cover by plant functional group (B) in
undisturbed and bark beetle-killed lodgepole pine and Douglas-fir forests. Plotted values are
untransformed mean ± 2SE. LP and DF denote a significant effect of disturbance class within
lodgepole pine (LP) or Douglas-fir (DF) forests only (P < 0.05; N = 10; GLM: y = disturbance
class). SPP, CL, and S*C denote significant (P < 0.05; N = 20) effects of species (SPP),
disturbance class (CL), and their interaction (S*C) in the global GLM: y = species + disturbance
class + species*disturbance class. Total basal area: SPP, S*C. Total cover: LP, DF, SPP, CL.
Figure 3. Litter quantity and litter N metrics in undisturbed and bark beetle-killed
lodgepole pine and Douglas-fir forests. Plotted values are untransformed mean ± 2SE. (A)
litter depth, (B) litter mass, (C) fresh (<1 yr) needle litter %N, (D) all needle litter %N, (E) total
litter %N, (F) total litter N pool. Values inset within each panel are P values from the GLM: y =
species + disturbance class + species*disturbance class.
70
Figure 4. Soil N metrics in undisturbed and bark beetle-killed lodgepole pine and Douglas-
fir forests. Plotted values are untransformed mean ± 2SE. (A) extractable NH4+, (B) extractable
NO3-, (C) net N mineralization, (D) net nitrification, (E) nitrification fraction, (F) organic N.
Values inset within each panel are P values from the GLM: y = species + disturbance class +
species*disturbance class.
Figure 5. Canopy foliar N metrics and foliar biomass in undisturbed and bark beetle-killed
lodgepole pine and Douglas-fir forests. Plotted values are untransformed mean ± 2SE. (A)
fresh foliar %N, (B) total foliar %N, (C) live canopy biomass, (D) canopy N pool. Values inset
within each panel are P values from the GLM: y = species + disturbance class +
species*disturbance class.
71
FIGURES
Figure 1
Similar effects; increased %N in canopy foliage and decreased N pool size in beetle-killed stands of both species.
Increased %N in canopy foliage.Sharp decline in N pool size.
Canopy of unattacked trees
Larger increases in pool size and flux rate in Douglas-fir; greater soil N in Douglas-fir overall
Increased NH4+, NO3
-, net N mineralization, & net nitrification
Soils
Greater increases in litter %N and N pool size in Douglas-fir; greater litter N in Douglas-fir overall
Increased litter mass, depth, and %N of fresh needle litter
Litter
Increased understory cover in beetle-killed forests of both types, with more cover overall in Douglas-fir.
Increased cover-canopy opening-increased light, H2O, nutrients
Understory cover
Similar effects; cooler soil temperatures in beetle-killed forests of both species
Decreased soil temperature-insulating litter input -shading by understory growth
Soil temperature
Hypothesized differences in beetle effects in Douglas-fir compared to lodgepole pine forests
Hypothesized bark beetle effects in Douglas-fir forests
Ecosystem component
Similar effects; increased %N in canopy foliage and decreased N pool size in beetle-killed stands of both species.
Increased %N in canopy foliage.Sharp decline in N pool size.
Canopy of unattacked trees
Larger increases in pool size and flux rate in Douglas-fir; greater soil N in Douglas-fir overall
Increased NH4+, NO3
-, net N mineralization, & net nitrification
Soils
Greater increases in litter %N and N pool size in Douglas-fir; greater litter N in Douglas-fir overall
Increased litter mass, depth, and %N of fresh needle litter
Litter
Increased understory cover in beetle-killed forests of both types, with more cover overall in Douglas-fir.
Increased cover-canopy opening-increased light, H2O, nutrients
Understory cover
Similar effects; cooler soil temperatures in beetle-killed forests of both species
Decreased soil temperature-insulating litter input -shading by understory growth
Soil temperature
Hypothesized differences in beetle effects in Douglas-fir compared to lodgepole pine forests
Hypothesized bark beetle effects in Douglas-fir forests
Ecosystem component
72
Figure 2
0
20
40
60
80
100
120
Undisturbed Beetle-killed Undisturbed Beetle-killed
% C
over
Tree seedlings: n.s.Forbs: LP, DF, SPP, CLShrubs: SPPSedges: n.s.Grasses: DF, SPP, CL
Lodgepole pine Douglas-fir
0102030405060708090
Bas
al a
rea
(m2 h
a-1)
Dead: LP, DF, SPP, CL, S*C
Live: LP, DF, SPP, S*CA)
B)
73
Figure 3
0.0
0.5
1.0
1.5
2.0
2.5
3.0
3.5
4.0
4.5Li
tter d
epth
(cm
) A) Species: 0.002
Class: 0.057Species*Class: ns
0
500
1000
1500
2000
2500
3000
Litte
r mas
s (g
m-2
)
Species: nsClass: nsSpecies*Class: ns
B)
0.0
0.2
0.4
0.6
0.8
1.0
1.2
1.4
1.6
Fres
h ne
edle
litte
r %N
Species: 0.002Class: <0.001Species*Class: 0.045
C)
0.0
0.2
0.4
0.6
0.8
1.0
1.2
1.4
1.6
All
need
le li
tter %
N
Species: <0.001Class: <0.001Species*Class: ns
D)
0.0
0.2
0.4
0.6
0.8
1.0
1.2
1.4
1.6
PICO PSME
Tota
l litt
er %
N
Species: <0.001Class: 0.003Species*Class: 0.021
E)
0
5
10
15
20
25
30
35
PICO PSME
Tota
l litt
er N
poo
l siz
e (g
N m-2
)
Species: 0.002Class: 0.035Species*Class: ns
F)
74
Figure 4
0
2
4
6
8
10
12E
xtra
ctab
le N
H4+ (µ
g N
g s
oil-1
)A)
Species: <0.001Class: 0.001Species*Class: ns
0
2
4
6
8
10
12
Ext
ract
able
NO3
- (µg
N g
soi
l-1)
Species: <0.001Class: 0.053Species*Class: ns
B)
0
5
10
15
20
25
30
35
40
Net
N m
iner
aliz
atio
n (µ
g N
g s
oil-1 y
r-1)
Species: nsClass: 0.012Species*Class: ns
C)
0123456789
10
Net
nitr
ifica
tion
(µg
N g
soi
l-1 y
r-1)
Species: 0.027Class: 0.023Species*Class: ns
D)
0.0
0.1
0.2
0.3
0.4
0.5
0.6
0.7
PICO PSME
Nitr
ifica
tion
fract
ion
Species: nsClass: nsSpecies*Class: ns
E)
0
500
1000
1500
2000
2500
3000
3500
4000
4500
PICO PSME
Org
anic
N (µ
g N
g s
oil-1
)
Species: <0.001Class: nsSpecies*Class: ns
F)
75
Figure 5
0.0
0.2
0.4
0.6
0.8
1.0
1.2
1.4
1.6
PICO PSME
Fres
h fo
liar %
NSpecies: >0.001Class: 0.058Species*Class: ns
A)
0.0
0.2
0.4
0.6
0.8
1.0
1.2
1.4
1.6
PICO PSME
Tota
l fol
iar %
N
Species: nsClass: 0.0003Species*Class: ns
B)
0
20
40
60
80
100
120
140
160
PICO PSME
Can
opy
N p
ool (
kg N
ha-1
)
Species: ns Class: <0.0001Species*Class: 0.040
D)
02000400060008000
100001200014000160001800020000
PICO PSME
Live
can
opy
biom
ass
(kg
ha-1
)
Species: ns Class: <0.001Species*Class: 0.045
C)
76
CHAPTER 3
Does salvage logging alter ecosystem response to
bark beetle disturbance in lodgepole pine forests?
ABSTRACT
Unprecedented outbreaks of native bark beetles in conifer forests of western North
America have elicited multiple management strategies including salvage logging. Salvage has
the potential to impact a wider range of forest ecosystem components than beetle outbreak alone,
which may lead to different ecosystem responses in salvaged forests compared to beetle-killed
only. In this study we use a paired and replicated before-after-control-impact (BACI) design to
test the first-year effects of salvage on sapling density, understory cover, soil microclimate, and
litter and soil N cycling of lodgepole pine (Pinus contorta Dougl.) forests attacked by the
mountain pine beetle (Dendroctonus ponderosae Hopkins). Salvage harvest removed most basal
area and decreased the density of all tree saplings (height <1.4 m) by 50%, although lodgepole
pine sapling density was unchanged and appeared adequate for stand replacement. Understory
biotic cover declined with salvage harvest, graminoid and shrub cover was unchanged, and cover
of downed coarse wood increased. Mid-summer soil temperature increased by ~1 ºC at both the
litter-soil interface and 10 cm depth in salvaged plots, and the N concentration of needle litter
increased by 30%. Soil and resin bag N were largely unaffected by salvage except for extractable
soil NO3- which tripled in concentration but remained low at 1.5 µg N g soil-1. Collectively, the
initial effects of post-beetle salvage were modest. Beetle-induced declines in plant N uptake and
increases in litter input appeared to be the main drivers of change in litter and soil N cycling over
time, with little additional influence of salvage during the first year following harvest.
77
INTRODUCTION
Concurrent outbreaks of several native bark beetle species throughout western North
America in the past decade have led to concern regarding ecosystem function, timber production,
tourism, and forest fire (Flint 2006, Flint and Haynes 2006, McFarlane and Witson 2008). In
response, management strategies have utilized both preemptive and post-outbreak approaches to
mediate the impact of this disturbance. Preemptive strategies include disrupting or mimicking
beetle pheromone communication (Bentz et al. 2005, Progar 2005, Gillette et al. 2006, Jeans-
Williams and Borden 2006), and using silvicultural methods to reduce stand susceptibility
(Safranyik et al. 1999, Fettig et al. 2007). Both methods have had moderate successes over
limited temporal and spatial scales, yet landscape-scale suppression of beetle outbreak is unlikely
(Raffa et al. 2008). Post-outbreak remediation techniques include fuel reduction treatments and
salvage harvests. Because beetle and salvage logging disturbances affect different ecosystem
components, interactions of these disturbance types may result in unique consequences for
ecosystem structure and function (Paine et al. 1998) compared to bark beetles alone.
Canopy biomass is a strong control on the rates and retention efficiency of forest N
cycling (Prescott 2002). Standing dead wood contains a significant amount of N (Fahey et al.
1985, Edmonds and Eglitis 1989, Herrmann and Prescott 2008) that is exported from the
ecosystem by salvage. However, rates of N input from coarse wood are small and contribute
little to annual rates of soil N turnover (Laiho and Prescott 2004). Fine woody debris and needle
litter (needles, twigs, reproductive structures, and bark) have higher N content and faster
turnover than coarse wood, and thus contribute more to overall N cycling in conifer forests
(Laiho and Prescott 1999). Beetle-induced tree mortality initiates a pulse of N-enriched needle
litter as the canopy decays (Morehouse et al. 2008, Klutsch et al. 2009, Griffin et al. 2011).
78
Salvage inputs of litter and fine woody debris are variable depending on method and site
treatment (Smethurst and Nambiar 1990, Goodman and Hungate 2006), and in beetle-killed
stands would likely depend on the extent of canopy deterioration that occurred due to beetles
prior to salvage. Litter and slash material may serve as a temporary N sink if it remains on the
forest floor undisturbed rather than burned, however slash burning can cause a pulse of soil
inorganic N lasting up to three years (Lopushinsky et al. 1992). Slash effects on soil N also vary
with time since disturbance and forest type. Following an outbreak of spruce beetle
(Dendroctonus rufipennis Kirby) in Alaska, Goodman and Hungate (2006) found no effect of
salvage on soil inorganic N four years after harvest. In contrast, a variety of slash treatments
following non-salvage harvest of Monterey pine (Pinus radiata D. Don.) forests all yielded
elevated N mineralization in the first two years after harvest (Smethurst and Nambiar 1990).
Logging activity can have a strong influence on both abiotic and biotic properties of
forest soils. Machinery can cause soil compaction and increased bulk density (Blouin et al. 2005)
that reduces porosity and water holding capacity, and increases soil strength (Blouin et al. 2008).
Soil temperature may also be elevated by logging, though the magnitude is dependent upon site-
treatment methods (Smethurst and Nambiar 1990). In contrast, beetle-induced litterfall can
decrease soil temperatures in the first few years after attack (Griffin et al. 2011). In turn, these
abiotic changes contribute to shifts in microbial community structure (Chatterjee et al. 2008) and
mycorrhizal associations (Kranabetter et al. 2006). The net result of these changes on soil N
dynamics is unclear. Some studies report short-lived increases in mineralization rates following
compaction (Kranabetter et al. 2006); others report reductions in soil inorganic N pool sizes
(Blouin et al. 2005, Choi et al. 2005, Kamaluddin et al. 2005). Variability among results may be
79
due to differing degrees of compaction, variation in slash treatments, or variation among forest
types in nutrient content of ecosystem pools.
Beetle and salvage disturbances also differ in their potential impact on tree saplings (here
defined as stems <1.4 m high), which provide a source of advance regeneration that could
ultimately lead to canopy replacement. In lodgepole pine (Pinus contorta Dougl.) forests,
successful mountain pine beetle (Dendroctonus ponderosae Hopkins) attack is limited to canopy
and large sub-canopy trees (Safranyik and Carroll 2006). Thus, post-beetle sapling densities are
largely unchanged and are usually sufficient for stand replacement with species compositions
consistent with local substrate and site conditions (Shore et al. 2006, Astrup et al. 2008, Nigh et
al. 2008, Rocca and Romme 2009, Vyse et al. 2009). However, salvage logging can have direct
and indirect, positive and negative effects on sapling density (Jonasova and Prach 2004). Direct
effects include physical damage or mortality of existing seedlings and saplings (Donato et al.
2006). Indirect mechanisms involve the effects of soil compaction and nutrient levels on tree
seedling growth and survival (Blouin et al. 2005, Bulmer and Simpson 2005, Kamaluddin et al.
2005, Blouin et al. 2008) and alteration of microsite conditions such as bare mineral soil or nurse
logs necessary for successful seedling establishment (Jonasova and Prach 2004). A flush of
understory growth and community composition change are common responses to both beetle
(McCambridge et al. 1982, Stone and Wolfe 1996, McMillin and Allen 2000) and salvage
(Hanson and Stuart 2005, del Rio 2006) disturbances, and are likely driven by increased light,
water, nutrient, and substrate availability.
In turn, litter inputs and vegetative cover can affect soil N cycling by shading and cooling
soils, providing N sinks in biomass, or altering rates of N fixation. Studies of other forest
management techniques suggest that soil disturbance is a key driver of vegetation differences
80
among beetle-killed stands with and without salvage. In cut-and-leave treatments, beetle-killed
trees are felled by hand and left on-site, requiring no heavy machinery or log skidding that
disturbs soil. Following an outbreak of the southern pine beetle (Dendroctonus frontalis
Zimmerman) in loblolly pine (Pinus taeda (L.)), Coleman et al. (2008) showed few differences
in understory tree or ground vegetation responses between beetle-killed stands with and without
this treatment type.
In this study, we use field sampling and laboratory analyses to compare stand structure,
ground cover, soil temperature, litter quantity and quality, and soil N metrics in beetle-killed vs.
beetle-killed + salvage logged stands of lodgepole pine. We hypothesized that salvage would
reduce sapling density, and increase the cover of bare soil, coarse wood, and understory
vegetation. We also hypothesized litter mass, depth, and N pool size would increase as more
material was added to the forest floor following salvage. Expectations for soil temperature were
less clear, as canopy opening would tend to increase soil temperatures while additional inputs of
litter mass may serve to insulate soils and cool soil temperature. With the addition of soil
disturbance from logging in beetle-killed forests, we expected soil inorganic N pools and the
accumulation of N on buried resin bags to also increase relative to beetle disturbance alone.
METHODS
Study region and experimental design
All study sites were located within a 4 km2 area of the Green River Lakes region on the
Bridger-Teton National Forest in northeastern Wyoming, USA (Figure 1). Forests are dominated
by lodgepole pine, with minor components of subalpine fir (Abies lasiocarpa Hook.), Engelmann
spruce (Picea engelmannii Parry), and whitebark pine (Pinus albicaulis Engelmann) (Despain
81
1990). Mean temperature is 14.2 °C July and -10.2 °C in January, and the region averages 297
mm precipitation per year, mostly as snow (WRCC 2010a). Soils are relatively nutrient poor and
derived from andesitic substrates. Mountain pine beetle activity in the area peaked in 2005,
though affected forests were of a mix of unattacked trees and beetle-killed trees in the red and
gray stages of canopy decline (Griffin et al. 2011) at the time of initial sampling in 2007.
We used a paired and replicated before-after-control-impact (BACI) experimental design
(Underwood 1994) to test for an effect of salvage-logging on stand structure, ground cover, soil
microclimate, and N cycling parameters in litter and soils of beetle-killed lodgepole pine. Pre-
salvage structure, litter, and soil N pool data were collected in 2007 from 16 circular plots of 8 m
radius. Eight plots were located in beetle-killed stands designated to be salvage-logged, and
eight were located in paired similar beetle-killed stands <400m away not designated for salvage.
Salvage logging in the beetle + salvage plots occurred in summer 2009. Harvest included both
beetle-killed and unattacked trees, though >90% of the basal removed was beetle-killed. Cutting
was performed by feller-buncher, with de-limbing done at landing sites and slash returned to and
scattered in the plots. Post-salvage structure, litter, and soil N pool data were collected from all
16 plots in summer 2010, while soil temperatures and resin bag N accumulation were measured
continuously from summer 2008 through summer 2010.
Vegetation
All live and dead trees greater than breast height (1.4m) were identified to species and
measured for diameter at breast height (DBH). Beetle-killed trees were identified by the presence
of exit holes and pitch tubes on the bark exterior, boring dust at the base of trees, and distinctive
J-shaped Dendroctonus galleries underneath the bark. Trees <1.4m were tallied by species and
measured to the nearest 10-cm height class in the northeast quadrant of each 8-m radius plot
82
(50m2), as well as in two 50-m2 rectangular plots 15 m to the east and west. Ground cover of
litter, coarse wood, bare soil, and three plant functional groups (forbs, sedges, and graminoids
(grasses + sedges)) was visually estimated to the nearest 10% in ten 0.25-m2 circular microplots;
total cover was allowed to exceed 100% to account for multiple strata of ground vegetation.
Microplots were located within the central 5-m radius area of the plot using a stratified random
design of fixed distances (one at 0.5 m; two at 1.5 m, 2.5 m and 3.5 m; and three at 4.5 m) and
random bearings (in 10º increments) from the plot center.
Microclimate
Soil temperature was measured hourly in four of the eight plot pairs from June 2008
through September 2010 using three pairs of iButton datalogger probes (Maxim Integrated
Products Inc., Dallas Semiconductor, Sunnyvale, CA) per plot. Beetle + salvage plots were not
instrumented during the salvage period in summer 2009. One of each iButton pair was installed
at the litter-organic soil interface, and the second was installed at 10 cm soil depth. For each
depth, temperature probe data were summarized as follows. First, plot-level hourly temperatures
were determined by averaging hourly data from the three probes per plot. Plot-level daily mean
temperatures were then calculated from the hourly data, and averaged to determine treatment-
level daily means.
Litter quantity and quality
In each 0.25-m2 microplot, litter depth was recorded at three locations and a 400-cm2
sample of the litter layer was collected and oven-dried at 60 ºC. Plot-level litter depth and mass
were obtained by averaging values from the 10 microplots. Litter from each microplot was sorted
into three categories: fresh current-year needle litter, identified by bright red color and lack of
mottling on surface (Morehouse et al. 2008); all-needle litter (all ages of needles combined); and
83
total litter (all foliar litter components and woody litter components < 1.0 cm wide). Sorted litter
was composited by plot for each category, and ground to powder for C:N analysis on a Leco
CNS-2000 at the University of Wisconsin Soil and Plant Analysis Laboratory (UWSPAL 2010).
Soil chemistry, soil N pools, and resin bag N
One soil core was collected from each 0.25-m2 microplot (N = 10 per plot) using a 5-cm
diameter x 15-cm long PVC corer. Soils were sieved (2 mm mesh), weighed, and divided into
three subsamples: (1) 30 g oven-dried at 60 ºC for gravimetric percent moisture (2007 and 2010);
(2) 20 g extracted in 75 ml of 2M KCl for 2 hours, with the extract then filtered and frozen for
later analysis of NH4+ and NO3
- pools (2007 and 2010); (3) 20 g air-dried and bulked by plot for
soil texture and chemical analyses (2007 only). Air-dried soil was analyzed for pH, total N,
exchangeable Ca, Mg, and K, available P (Bray P1 extract), and organic matter at the University
of Wisconsin Soil and Plant Analysis Laboratory (UWSPAL 2010). Soil organic N was
determined by difference using total N and inorganic N values, and soil texture was determined
using the Bouyoucos hydrometer technique (Bouyoucos 1962). KCl extractions were analyzed
for [NH4+] and [NO3
-] using colorimetric methods on an Astoria Pacific II continuous flow
autoanalyzer.
Resin bags were constructed using 20 g of mixed bed ion exchange resin (J.T Baker
#JT4631-1) tied inside a piece of un-dyed nylon stocking material (Binkley et al. 1992). One
resin bag was incubated at 10 cm soil depth in each microplot (N = 10 / plot). Sampling occurred
in four summer periods (June-August in 2007 and 2008, June-October in 2009, and June-
September in 2010) and three winter periods (August-June in 2007-8 and 2008-9, and October-
June in 2009-10). Beetle + salvage plots were not measured during the salvage period in summer
84
2009. Upon retrieval, resins were removed from the nylon and extracted in 2M KCl and analyzed
in the same manner as the soil samples described above.
Statistical analyses
To test for similarity in pre-salvage site conditions between the beetle only and beetle +
salvage stands, we performed paired t tests (α < 0.05) on topographic (elevation, slope, aspect),
soil (texture, N, organic matter (OM), pH, and cations), vegetation (basal area, density of <1.4m
stems, and ground cover), and litter (quantity and quality) metrics measured in 2007. Potential
effects of post-beetle salvage logging on these variables were tested by calculating plot-level
changes between 2007 and 2010, and performing paired t-tests (α < 0.05) of this change in beetle
only vs. beetle + salvage plots. Soil temperatures were analyzed using mid-summer (July 1-
August 31) and mid-winter (January 1-February 28) data from before and after the salvage
period. For each soil depth, we first calculated the plot-level difference in mean daily
temperatures between like calendar days before and after salvage, then performed paired t-tests
for a salvage effect on these differences (N = 62 for summer; N = 59 for winter). Resin bag data
within each sampling period were calculated as the rate of N accumulation (µg N g resin-1 day-1)
and the mass of N collected (µg N g resin-1). Potential effects of post-beetle salvage logging on
resin bag N were tested using paired t-tests of the change over time in beetle only vs. beetle +
salvage plots using only the years immediately before and after salvage (2010-2008). All
variables were tested for normality and transformed when necessary to meet the assumptions of
statistical methods. Unless otherwise noted, reported variance measures are two standard errors.
RESULTS
Pre-salvage site conditions.
85
Beetle only and beetle + salvage plots were similar (P > 0.05) in all topographic,
vegetation, litter, cover, and soil metrics except for organic soil depth, which was deeper by 0.5
cm in beetle-only sites (Table 1). Mean elevation of all plots was 2518 ± 13 m, with a mean
slope of 14.3 ± 4º. Lodgepole pine averaged 97% of the total basal area (live + dead) of all
stands, and beetle mortality averaged 60% of lodgepole basal area (Table 1). The density of all
saplings was highly variable among all plots prior to salvage but was not significantly different
between beetle only and beetle + salvage plots (Table 1). Density of lodgepole pine saplings also
did not differ between beetle and beetle + salvage plots (Table 1), and was 61% and 62% of the
total density in each, respectively.
Vegetation and ground cover
As expected, salvage logging reduced total lodgepole pine basal area by 90%, which
included small amounts of both live and non-beetle killed dead basal area in addition to beetle-
killed trees (Figure 2a). Live unattacked basal area declined similarly in both beetle only and
beetle + salvage sites, and beetle-killed basal area increased in beetle-only plots (Figure 2a). In
the understory, salvage reduced total biotic cover, largely in response to a decline in forb cover,
and there was no effect on graminoid or shrub cover (Figure 2b). Ground cover of bare soil
increased after salvage from 0.1 to 8.4%, and coarse wood cover increased from 2.3 to 11.6 %
(Figure 2b).
Salvage logging had a discernable effect on sapling density despite high variability in
density change (both positive and negative) between 2007 and 2010 in both beetle only and
beetle + salvage plots (P = 0.045). Total sapling density in beetle + salvage plots declined on
average 1892 ± 1648 stems ha-1 down to a mean of 1675 ± 1001 stems ha-1. However, though the
densities of both 0-30cm and 30-140cm size classes declined in beetle + salvage sites, when
86
compared to changes in beetle only plots an effect of salvage was limited to the 30-140cm size
class (Figure 2a). For lodgepole pine saplings, there was no effect of salvage on either the total
density or the density of either the 0-30 or 30-140 cm size classes (Figure 3b).
Soil microclimate
Salvage logging had a warming effect on both mid-summer and mid-winter soil
temperatures at both the litter-soil interface and 10cm depth. Salvage increased mid-summer
litter-soil interface temperature by 0.8 ± 0.2 ºC, and 10 cm depth temperature by 1.1 ± 0.1 ºC (P
< 0.001 for both; Figure 3a). Mid-winter temperatures were 0.2 ± 0.05 and 0.3 ± 0.04 ºC warmer
at the litter-soil interface and 10 cm depth, respectively (P < 0.001 for both; Figure 3b). Mean
daily temperatures at the litter-soil interface peaked at 17 ºC in each of the three sampled
summers, and reached minimums of -3 and -9 ºC in the two sampled winters. At 10 cm depth,
temperatures peaked at 12-14 ºC in summers, and reached lows of 3 and 6 ºC in winters.
Litter quantity and quality
Salvage had little effect on litter amounts or N content, though several litter variables
changed over time across all plots. Litter mass increased by 37% (1600 to 2200 g m-2) between
2007 and 2010 but was unaffected by salvage (Figure 5a). Litter depth averaged 2.7 ± 0.2 cm and
did not change over time or with salvage (Figure 5b). Fresh (<1yr) needle litter %N was 0.73 ±
0.03 % across all sites and did not change with either time or salvage (Figure 5d). Though the
%N of all needle litter was unchanged in beetle-only sites, it increased following salvage (Figure
5e) from 0.72 ± 0.08 to 0.94 ±0.07% but was insufficient to cause a significant salvage effect in
total litter %N (Figure 5f). Total litter %N did, however, increase across all plots from 0.81 ±
0.06 to 0.98 ± 0.06% between 2007 and 2010 (Figure 5f). Increased litter mass and increased
87
total litter %N both contributed to a 50% increase over time in the litter N pool across all plots
from 1473 ± 148 to 2177 ± 295 g N m-2, with no additional effect of salvage (Figure 5c).
Soil and resin bag N
Pools of extractable soil inorganic N changed through time and with salvage. Soil NH4+
tripled across all plots from 5.7 ± 0.9 to 17.3 ± 0.9 µg N g soil-1, with no additional effect of
salvage (Figure 6a). In contrast, soil NO3- was unchanged in beetle-only plots through time but
almost tripled in beetle + salvage plots from 0.6 ± 0.2 to 1.6 ± 0.5 µg N g soil-1 from 2007 to
2010 (Figure 6b). However, this difference in soil NO3- was insufficient to cause an effect of
salvage on total extractable inorganic N, which was dominated by NH4+ and increased from 6.2 ±
1.0 to 18.9 ± 3.0 µg N g soil-1 across all plots between 2007 and 2010.
Resin bag NH4+ accumulation rate increased throughout the study in all plots (Figure 7a).
In the pre-salvage period, NH4+ accumulation rate increased in all plots from < 0.10 to ~0.20 µg
N g resin-1 day-1 between summer 2007 and summer 2008. NH4+ accumulation rate continued to
increase through the post-salvage period in all plots, reaching ~0.35 µg N g resin-1 day-1 in the
summer of 2010. Patterns of NH4+ mass collected paralleled those of accumulation rate (Figure
7b). There was no effect of salvage on the changes in seasonal or annual NH4+ accumulation
rates, or seasonal mass collected between the year before and the year after the salvage period (P
> 0.05). Annual mass of NH4+ collected was also unaffected by salvage (Figure 8). Within the
consistently increasing trend, resin bag NH4+ accumulation oscillated between higher rates in
summer and lower rates in winter (Figure 7a). Resin bag NO3- accumulation rate was consistent
throughout the study and almost always under 0.10 µg N g resin-1 day-1 during both summer and
winter (Figure 7b), with parallel patterns in total mass collected (Figure 7d). There was no effect
88
of salvage on seasonal or annual NO3-accumulation rates or seasonal mass of NO3
- collected (P >
0.05). Annual mass of NO3-collected was also unaffected by salvage (Figure 8).
DISCUSSION
In this study we tested the effects of post-disturbance salvage logging on the structure,
microclimate, and N cycling of bark beetle-killed lodgepole pine forests. Bark beetle disturbance
alone redistributes material among the canopy and forest floor, and influences both the
microclimate and N cycling of the litter-soil profile compared to undisturbed forests (Griffin et
al. 2011). We hypothesized that salvage logging of beetle-killed forests would cause further
canopy disturbance and litter inputs, as well as new disturbance to saplings, understory
vegetation, and the forest floor which together may in turn alter patterns of N cycling in the
litter-soil profile compared to beetle-kill alone. Salvage had the greatest impact on forest
structure, though there were substantive effects on ground cover, soil microclimate and litter and
soil N cycling as well. In addition to declines in total basal area, density of 30 to 140-cm high
saplings, and % cover of forbs and all plants, salvage caused an increase in mid-summer and
mid-winter soil temperatures, %N content in needle litter, extractable soil NO3-, and the percent
cover of coarse wood.
Tree sapling density and regeneration potential
Salvage-induced declines in the total sapling density observed here are consistent with
those found in other salvaged forests (Jonasova and Prach 2004, Keyser et al. 2009). However,
salvage did not affect the smallest size classes (0 to 30-cm high) of all saplings, or any size class
of lodgepole pine sapling. This may be due to the undisturbed forest floor in post-beetle forests
which may protect smaller stems from trampling or erosion compared to post-fire forests, and is
89
a common result following salvage (Vyse et al. 2009). The average density of stems <1.4m high
in salvaged plots of this study remained above 1600 stems ha-1, which is considered sufficient for
stand replacement (Vyse et al. 2009), and also remained dominated by lodgepole pine (>50% of
total density). Effects of salvage on sapling density are variable among forest types, pre-salvage
disturbance types, and salvage method. In a study of post-southern pine beetle (Dendroctonus
frontalis Zimmermann) salvage logging in loblolly pine (Pinus taeda L.), Coleman et al. (2008)
concluded that cut and leave treatments did not change the successional trajectory of beetle-
killed forests. Similarly, Peterson and Leach (2008) found no effect of post-windthrow salvage
on sapling densities in loblolly pine, though there was an effect on species composition. In
contrast, post-beetle (Ips typographus L.) clear-cut salvage logging in a Norway spruce (Picea
abies (L.) Karst.) forest reduced the density of spruce saplings and promoted early successional
understory species (Jonasova and Prach 2004, 2008).
Microclimate and ground cover
The salvage effect of 1 ºC increase in temperature below the litter and at 10-cm depth
observed here following beetle salvage is consistent in magnitude with remotely-sensed surface
temperatures in beetle-killed spruce and clear-cut forests of the Czech Republic, which showed a
difference of 2 ºC between beetle-killed and clear-cut forest. (Hais and Kucera 2008). However,
initial beetle disturbance has been shown to lower soil temperature compared to undisturbed
forests due to increased litter input (Byers 1984, Griffin et al. 2011). Litter mass did not increase
with salvage in this study and thus likely did not provide any additional cooling effect on soils.
The removal of tree boles and the decreases in biotic ground cover and in the density of stems
<1.4 m high may have reduced soil shading, and could explain how salvage increased soil
90
temperatures. Warmer temperatures may be associated with drier soils, which would have a
suppressing effect on soil N turnover.
Increased cover of coarse wood after salvage logging is likely to create more variability
in forest floor microsite conditions, influencing the rates of litter decomposition and subsequent
return of inorganic N to soils at fine spatial scales (Remsburg and Turner 2006). However,
similar changes are likely to occur in beetle-kill only forests as dead trees begin to fall, and over
time total coarse wood loading in beetle-kill only forests will likely be much higher than in
salvaged forests (Lewis 2009). Decreases in total biotic cover observed 1 year following salvage
are not likely to remain (Kurulok and Macdonald 2007), as a flush of understory growth spurred
by increased water, light, and nutrients is often observed following both beetle mortality (Stone
and Wolfe 1996) and post-beetle salvage harvest (Jonasova and Prach 2008). Though highly
variable and not significant in paired analyses, percent cover of bare soil showed an increasing
trend in salvaged sites and possibly contributed to significant decreases in biotic cover. However,
over time these areas are likely to be colonized by understory vegetation. With more mineral soil
substrate exposed as seedbed, colonization by pioneer species may alter understory community
composition in salvaged sites relative to beetle-kill alone. In a Colorado spruce-fir forest salvage
logged after windthrow, understory communities shifted to dominance by graminoids (del Rio
2006). Thus, despite initial differences in biotic cover after 1 year, prediction of long term
impacts of salvage on understory composition is difficult.
Litter and Soils
Several metrics of litter and soil N cycling changed similarly over time in all plots,
consistent with expectations of canopy deterioration beetle-killed forests (Griffin et al. 2011).
Total litter mass, %N, and litter N pool size were unaffected by salvage logging, but all increased
91
over the three years between sampling. These results suggest that either (1) most foliage had
fallen from beetle-killed trees in all plots prior to salvage, or (2) that the amount of salvage-
induced litterfall in beetle + salvage plots was equivalent to litterfall in beetle-kill only plots over
the same time period. In either scenario, salvage does not appear to increase the amount of litter
mass or N input which would already occur following beetle outbreak. Salvage did increase the
%N of all needle litter, although this was insufficient to influence the total litter N pool size.
This may be due to salvage-induced inputs of green needle litter from trees that survived the
outbreak and had elevated foliar N content (Griffin et al. 2011).
Similarly, patterns of change in soil and resin bag N of salvaged forests were mostly
consistent with those observed in beetle-killed only forests. Nitrogen availability is largely
determined by the balance between plant uptake and inputs of organic N to the forest floor.
Equivalent live basal areas between beetle only and beetle + salvage plots over time suggest that
canopy plant demand for N was not affected by salvage. Inputs of litter N suggest N supply was
also similar, which together may explain why few differences in soil N were observed.
Extractable soil NO3- pool was the only soil N metric that differed over time between beetle-only
and beetle + salvage plots, and though still low, approximately tripled in beetle + salvage plots
compared to beetle-kill alone. Increased soil NO3- could be due to reduced cover of forbs, which
preferentially take up NO3- over NH4
+ (Miller and Bowman 2002, Falkengren-Grerup et al.
2004).
An increase in NO3- over time was not detected with buried resin bags in either beetle
only or beetle + salvage plots, while resin bag NH4+ increased equally in both. With the
exception of increased (but still low) soil NO3-, salvage does not appear to substantially alter soil
N dynamics relative to the changes that would already occur in beetle-killed forests. Other
92
studies of tree mortality impacts on soil NO3- in Rocky Mountain conifers are consistent with
these results, and show that NO3- production and export remains low even after substantial tree
removal or death (Knight et al. 1991, Parsons et al. 1994, Prescott et al. 2003, Thiel and Perakis
2009). Beetle and salvage induced inputs of litter have high C:N ratios and may serve as N sinks
in the early stages of decomposition (Fahey et al. 1985, Remsburg and Turner 2006) and limit
the accumulation of N in soils.
CONCLUSIONS
Post-beetle salvage logging had detectable influences on ground cover, soil microclimate,
and N cycling through the litter-soil profile during the initial year following tree harvest.
However, the effects of salvage harvest were of less magnitude than we expected. Although
sapling density declined by 50%, the density of lodgepole pine saplings did not change and
appeared adequate for stand replacement. Salvage increased soil temperatures by ~ 1 ºC at both
the litter-soil interface and at 10cm depth, and increased the %N in all-needle litter but not in the
total litter N pool. With no effect on plant N demand (as inferred from basal area) or on litter N
inputs, salvage also did not have a substantial effect on soil N other than small increases in
extractable soil NO3-. Rather, N cycling in the litter and soils of both beetle only and beetle +
salvage sites changed over time in a manner consistent with that expected following bark beetle
outbreak with few modifications by salvage harvest. These data suggest that the harvest of
beetle-killed trees for wood products can have small initial consequences for many ecosystem
response variables. However, loss of snags and the future inputs of coarse wood to the forest
floor are likely to be pronounced over longer time frames.
93
ACKNOWLEDGEMENTS
We thank Emily Stanley, James Thoyre, and the Center for Limnology at UW for access
to laboratory equipment. For help in the field and the laboratory, we appreciate the assistance of
many University of Wisconsin students: Heather Lumpkin, Erin Mellenthin, Alex Rahmlow,
Amanda Rudie, Ben Ruh, Corey Olsen, Ryan Peasley, Greg Skupien, Lucille Marescot, Jaclyn
Entringer, Emily Ruebl, Ashton Mieritz, Shane Armstrong, and Nick Labonte.
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TABLES
Table 1. Pre-salvage characteristics for beetle-killed only and beetle-killed plus salvage-logged
sites. N = 8 sites per treatment. Boldface indicates significant paired t test at α = 0.05.
Disturbance type Site characteristic Beetle only Beetle + Salvage P Topography Elevation (m) 2510 ± 22 2526 ± 12 0.0899 Aspect (SW index) a -0.7 ± 0.4 -0.1 ± 0.6 0.2250 Slope (º) 14 ± 6 15 ± 5 0.7851 Soils Texture & Structure Sand (%) 61 ± 5 57 ± 5 0.3128 Silt (%) 26 ± 4 29 ± 3 0.1969 Clay (%) 13 ± 2 14 ± 2 0.8483 Bulk density (g cm-3) 0.69 ± 0.06 0.72 ± 0.07 0.5277 Organic soil depth (cm) 3.2 ± 0.3 2.7 ± 0.2 0.0137 Organic matter (%) 4.3 ± 0.8 4.7 ± 1.0 0.5424 Cations pH 5.4 ± 0.1 5.3 ± 0.2 0.4905 Ca (µg g-1; exch.) 1132 ± 293 1251 ± 361 0.5691 Mg (µg g-1; exch.) 131 ± 24 141 ± 34 0.5415 K (µg g-1; exch.) 187 ± 37 185 ± 25 0.8680 P (µg g-1; exch.) 25 ± 8 28 ± 5 0.4172 Nitrogen NH4
+ (µg N g-1) 5.3 ± 1.4 6.0 ± 1.1 0.3769 NO3
- (µg N g-1) 0.5 ± 0.3 0.6 ± 0.3 0.4116 Organic N (mg N g-1) 0.86 ± 0.25 0.94 ± 0.28 0.6978 Litter Mass (g m-2) 1614 ± 236 1668 ± 160 0.7698 Depth (cm) 2.7 ± 0.4 2.3 ± 0.4 0.1913 N pool size (g N m-2) 12.6 ± 2.2 13.9 ± 1.8 0.3794 <1 yr old needle %N 0.73 ± 0.07 0.69 ± 0.05 0.3402 All needles %N 0.74 ± 0.07 0.72 ± 0.08 0.6538 Total litter %N 0.78 ± 0.06 0.84 ± 0.10 0.303 P. contorta basal area (m2 ha-1) Live unattacked 8.3 ± 2.8 5.6 ± 2.8 0.0978 Live attacked 2.1 ± 2.2 0.6 ± 1.2 0.3050 Dead unattacked 5.0 ± 3.1 8.0 ± 4.3 0.2290 Dead attacked 22.4 ± 4.1 21.9 ± 5.0 0.8156 Sapling density (stems ha-1) All species 7133 ± 3380 3567 ± 1855 0.1010b
P. contorta 4383 ± 2854 2225 ± 1326 0.2798b
Ground cover (%) Bare soil 0.4 ± 0.5 0.1 ± 0.1 0.1705 Litter 42 ± 9 41 ± 10 0.7418 Coarse wood 2.6 ± 3.2 2.3 ± 1.7 0.8202 Total biotic 62 ± 11 60 ± 13 0.7322 Graminoid 16 ± 6 7 ± 4 0.0607 Shrub 7 ± 6 3 ± 3 0.0976 Forb 37 ± 11 49 ± 9 0.0664 a calculated as: cos(aspect-225)
b paired t test performed on log10 transformed data
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FIGURE LEGENDS
Figure 1. Study area, Green River Lakes Region of the Bridger-Teton National Forest.
Eight plot pairs were established in summer 2007, with one member inside a polygon designated
for salvage harvest and the other outside. All forests in the field of view are dominated by
lodgepole pine, and mountain pine beetle activity was extensive throughout the area at the time
of initial sampling (2007). Salvage harvests occurred in the summer of 2009, and post-salvage
data were collected in summer 2010.
Figure 2. Stand structure and ground cover change in beetle only and beetle + salvage
plots following logging. N = 8 pairs of plots. Error bars = 2 standard error. (A) Live and dead
basal area change (2010 – 2007) in attacked and unattacked lodgepole pine; (B) Ground cover
change in abiotic (bare soil, litter, and coarse wood) and biotic (total and by plant functional
group) classes. An asterisk indicates a significant difference at α = 0.05.
Figure 3. Change in sapling density in beetle only and beetle + salvage plots following
logging. N = 8 pairs of plots. Error bars = 2 standard error. (A) all saplings; (B) lodgepole pine
saplings. An asterisk indicates a significant difference at α = 0.05.
Figure 4. Average daily soil temperature in beetle only and beetle + salvage plots. Beetle +
salvage sites were not instrumented during the salvage period in summer 2009. Variable width of
each plotted time series = daily 95% confidence interval. Beetle-only data (black) is overlaid on
beetle + salvage data (gray). (A) temperature at the litter-organic soil interface; (B) temperature
at 10-cm soil depth. Salvage logging increased both mid-summer (July-August) temperature at
101
both soil depths by approximately 1 ºC, and mid-winter (January-February) temperatures by 0.2
ºC (P < 0.001; time periods used in analyses are shown in boxes).
Figure 5. Change in litter quantity and quality in beetle only and beetle + salvage plots
following logging. N = 8 pairs of plots. Error bars = 2 standard error. (A) litter mass; (B) litter
depth; (C) %N in <1yr old needle litter; (D) %N in all-needle litter; (E) %N in total litter; (F)
total litter N pool size. An asterisk indicates a significant difference at α = 0.05.
Figure 6. Change in soil inorganic N pool sizes in beetle only and beetle + salvage plots
following logging. N = 8 pairs of plots. Error bars = 2 standard error. (A) extractable NH4+; (B)
extractable NO3-; (C) total extractable inorganic N (TIN). An asterisk indicates a significant
difference at α = 0.05.
Figure 7. Resin bag N in beetle only and beetle + salvage plots. N = 8. Error bars = 2 standard
error. Beetle + salvage sites were not measured during the salvage period in summer of 2009. (A)
Resin bag NH4+ accumulation rate; (B) Resin bag NO3
- accumulation rate; (C) Mass of resin bag
NH4+; (D) Mass of resin bag NO3
-.
Figure 8. Change in annual mass of resin bag N in beetle only and beetle + salvage plots
following logging. N = 8 pairs of plots. Error bars = 2 standard error. (A) NH4+; (B) NO3
-; (C)
total inorganic N (TIN).
102
FIGURES
Figure 1
103
Figure 2
-28
-24
-20
-16
-12
-8
-4
0
4
8
Liveunattacked
Liveattacked
Deadunattacked
Deadattacked
Cha
nge
in b
asal
are
a (m
2 ha-1
)
Beetle onlyBeetle + Salvage
*
*
A)
-50
-40
-30
-20
-10
0
10
20
30
40
Bare soil Litter Coarsewood
Totalbiotic
Graminoid Shrub Forb
Cha
nge
in p
erce
nt c
over
(%)
*
* *
B)
104
Figure 3
-4000
-3000
-2000
-1000
0
1000
2000
3000
4000
Cha
nge
in to
tal d
ensi
ty ..
(sap
lings
ha-1
)
Beetle onlyBeetle + Salvage
A)
**
-4000
-3000
-2000
-1000
0
1000
2000
3000
4000
All stems<140 cm
0-30 cm 30-140 cm
Cha
nge
in lo
dgep
ole
pine
..de
nsity
(sap
lings
ha-1
) .
B)
105
Figure 4
-8.0-6.0-4.0-2.00.02.04.06.08.0
10.012.014.0
6/11
/200
87/
11/2
008
8/11
/200
89/
11/2
008
10/1
1/20
0811
/11/
2008
12/1
1/20
081/
11/2
009
2/11
/200
93/
11/2
009
4/11
/200
95/
11/2
009
6/11
/200
97/
11/2
009
8/11
/200
99/
11/2
009
10/1
1/20
0911
/11/
2009
12/1
1/20
091/
11/2
010
2/11
/201
03/
11/2
010
4/11
/201
05/
11/2
010
6/11
/201
07/
11/2
010
8/11
/201
09/
11/2
010
10 c
m d
epth
tem
p (C
) .
Beetle + Salvage
Beetle only
B)
-10.0-8.0-6.0-4.0-2.00.02.04.06.08.0
10.012.014.016.018.0
Litte
r-so
il in
terfa
ce te
mp
(C)
.
A)Salvage periodPre-salvage Post-salvage
106
Figure 5
0
200
400
600
800
1000
1200C
hang
e in
litte
r mas
s (g
m-2
) .
A)
-0.5
0.0
0.5
1.0
1.5
Cha
nge
in li
tter d
epth
(cm
) .
B)
0
2
4
6
8
10
12
14
Cha
nge
in li
tter N
poo
l (g
N m-2
).
C)
-0.1
0.0
0.1
0.2
0.3
0.4
Cha
nge
in <
1 yr
old
nee
dle
litte
r %N
.D)
-0.1
0.0
0.1
0.2
0.3
0.4
Cha
nge
in a
ll ne
edle
litte
r %N
.
E)
*
-0.1
0.0
0.1
0.2
0.3
0.4
Cha
nge
in to
tal l
itter
%N
.
F)
107
Figure 6
0
2
4
6
8
10
12
14
16
18
Cha
nge
in s
oil N
H 4
+ (µg
N g
soi
l -1)
.
A)
0.0
0.5
1.0
1.5
2.0
2.5
3.0
3.5
4.0
Cha
nge
in s
oil N
O 3
- (µg
N g
soi
l -1)
B)
*
0
2
4
6
8
10
12
14
16
18
Cha
nge
in T
IN (µ
g N
g s
oil -1
) .
C)
108
Figure 7
0.0
0.1
0.2
0.3
0.4
0.5
Rat
e of
NH
4+ acc
umul
atio
n .
(ugN
g re
sin
-1 d
ay -1
)
A)
Salvage period
Pre- Salvage
Post- salvage
0.0
0.1
0.2
0.3
0.4
0.5
Rat
e of
NO
3- a
ccum
ualti
on .
(ug
N g
resi
n -1
day
-1)
Beetle onlyBeetle + Salvage
B)
0
10
20
30
40
50
60
70
Mas
s of
NH
4+ acc
umul
ated
.
(ug
N g
resi
n -1
)
C)
0
10
20
30
40
50
60
70
Summer2007
Winter2007-08
Summer2008
Winter2008-09
Summer2009
Winter2009-10
Summer2010
Mas
s of
NO
3- a
ccum
ulat
ed .
(ug
N g
resi
n -1
)
Beetle onlyBeetle + Salvage
D)
109
Figure 8
-30
-20
-10
0
10
20
30
40
50
60
Cha
nge
in a
nnua
l mas
s of
resi
n .
bag
N (u
g N
g re
sin -1
)
Beetle only Beetle + Salvage
NH4+ NO3
- TIN
110
THESIS CONCLUSIONS
Disturbances can change and shape ecosystems by creating or reducing spatial heterogeneity,
redistributing biomass among ecosystem pools, altering resource availability and abiotic
conditions, or initiating community composition change (Bormann and Likens 1979, Pickett and
White 1985, Turner and Dale 1998). In Greater Yellowstone and across western North America,
bark beetles have thus shaped conifer forests as a native component of the disturbance regime
throughout the Holocene (Brunelle et al. 2008). However, recent outbreaks have reached an
unprecedented spatial extent and severity (Raffa et al. 2008), the frequency and severity of
outbreaks are expected to increase in a warming climate (Logan et al. 2010), and there is
growing concern over the impact of beetle outbreak on ecosystem services (Flint 2006) and
interactions with other disturbance types (Simard et al. 2011). In this dissertation I have sought
to fill large knowledge gaps regarding the impact of beetle disturbance on N cycling, the role of
forest type in determining ecosystem response, and the effect of forest management on post-
beetle ecosystem function. Following are the main conclusions from this body of work:
(1) Post-beetle patterns of N cycling change in the litter, soil, and vegetation of conifer
forests were consistent with the guild-specific impacts of bark beetles on canopy
vegetation. Two to four years after disturbance, both biotic and abiotic controls were
important factors in the increase of N content in litter and soils, with concurrent increases
in the foliar N of undisturbed vegetation. By thirty years after outbreak, soil N
availability returned to undisturbed levels though the canopy N pool remained depleted.
Overall impact on soil N cycling was less severe than seen in post-fire forests.
111
(2) Contrasting conifer forest types have qualitatively similarly responses to beetle
disturbance in the litter-soil profile, though the response of undisturbed vegetation
varies. Despite substantial differences in the pre-disturbance N dynamics of lodgepole
pine and Douglas-fir forests, both showed 25-50% increases in needle litter N content and
soil N pools and fluxes. However, only in lodgepole pine did increased soil N
availability lead to increased foliar N of unattacked canopy trees.
(3) Beetle-killed forests exhibit several mechanisms of N retention following
disturbance, partially mitigating the potential for N loss via soil NO3- leaching.
Minimal increases in nitrification rate and soil NO3- pools, increased N concentration in
undisturbed canopy biomass, and declines in foliar Mn suggest multiple pathways for N
retention in beetle-killed forests.
(4) Salvage logging in beetle-killed forests has few additional impacts on N cycling in
the first year after harvest. Salvage harvest did not increase litter inputs or soil N
cycling any more than would have occurred in un-salvaged beetle-killed forests.
However, substantial differences in coarse wood inputs and understory vegetation
composition are likely to develop over longer time scales.
(5) Salvage reduced the total density of saplings in beetle-killed forests, though residual
densities should be sufficient for stand replacement. Salvage may promote the
dominance of lodgepole pine in post-beetle forests, as lodgepole pine saplings were less
susceptible to density decline following salvage than other species.
112
REFERENCES
Bormann, F. H., and G. E. Likens. 1979. Pattern and process in a forested ecosystem: disturbance, development, and the steady state based on the Hubbard Brook ecosystem study. Springer-Verlag, New York. Brunelle, A., G. E. Rehfeldt, B. Bentz, and A. S. Munson. 2008. Holocene records of Dendroctonus bark beetles in high elevation pine forests of Idaho and Montana, USA. Forest Ecology and Management 255:836-846. Flint, C. G. 2006. Community perspectives on spruce beetle impacts on the Kenai Peninsula, Alaska. Forest Ecology and Management 227:207-218. Logan, J. A., W. W. Macfarlane, and L. Willcox. 2010. Whitebark pine vulnerability to climate- driven mountain pine beetle disturbance in the Greater Yellowstone Ecosystem. Ecological Applications 20:895-902. Pickett, S. T. A., and P. S. White, editors. 1985. The ecology of natural disturbance and patch dynamics. Academic Press, Orlando, FL. Raffa, K. F., B. H. Aukema, B. J. Bentz, A. L. Carroll, J. A. Hicke, M. G. Turner, and W. H. Romme. 2008. Cross-scale drivers of natural disturbances prone to anthropogenic amplification: The dynamics of bark beetle eruptions. Bioscience 58:501-517. Simard, M., W. H. Romme, J. M. Griffin, and M. G. Turner. 2011. Do mountain pine beetle outbreaks change the probability of active crown fire in lodgepole pine forests? Ecological Monographs 81:3-24. Turner, M. G., and V. H. Dale. 1998. Comparing large, infrequent disturbances: What have we learned? Ecosystems 1:493-496.
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