Canadian Journal of Forest Research
Western spruce budworm effects on throughfall N, P, C fluxes and soil nutrient status in the Pacific Northwest
Journal: Canadian Journal of Forest Research
Manuscript ID cjfr-2018-0523.R2
Manuscript Type: Article
Date Submitted by the 02-Jun-2019 Author:
Complete List of Authors: Arango, Clay; Central Washington University, Biological Sciences Ponette-González, Alexandra; University of North Texas System, Department of Geography and the Environment Neziri, Izak; Central Washington University, Biological Sciences Bailey, Jen;Draft University of North Texas System, Department of Geography and the Environment
Keyword: herbivory, coniferous forest, outbreak insect, climate change, lepidoptera
Is the invited manuscript for consideration in a Special Not applicable (regular submission) Issue? :
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1 Western spruce budworm effects on throughfall N, P, C fluxes and soil nutrient status in
2 the Pacific Northwest
3
4 Clay Arango1*, Alexandra Ponette-González2, Izak Neziri1, Jennifer Bailey2
5
6 1Department of Biological Sciences, Central Washington University, 400 E University Ave,
7 Ellensburg, Washington 98926-7537, USA
8
9 2Department of Geography and the Environment, University of North Texas, 1155 Union Circle
10 #305279, Denton, Texas 76203, USA
11 Draft
12 *Corresponding author: [email protected], (509) 963-3163, fax (509) 963-2730
13
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14 Abstract
15 Western spruce budworm (Choristoneura freemani) is the most widely distributed insect
16 herbivore in western North American coniferous forests. By partially or completely defoliating
17 tree crowns, budworms influence fluxes of water, nutrients, and organic carbon from forest
18 canopies to soils and, in turn, soil chemistry. To quantify these effects, throughfall water,
19 inorganic nitrogen (N), phosphorus (P), and dissolved organic carbon (DOC) concentrations and
20 fluxes, and soil N and P concentrations were measured in coniferous forest sites with high and
21 background levels of budworm herbivory. Throughfall N and P concentrations and fluxes
22 increased at high budworm sites during and/or immediately after larval stage budworm feeding,
23 indicating reduced uptake and/or greater leaching from canopies as a result of budworm activity.
24 Annual throughfall N fluxes (<67-71 g NDraft ha-1 yr-1) and soil N concentrations were low regardless
25 of herbivory level. In contrast, throughfall P was considerably greater at sites with high (2174 g
26 P ha-1 yr-1) compared to background (1357 g P ha-1 yr-1) herbivory, and this was reflected in
27 nearly 3-fold higher soil P concentrations at high budworm sites. Our findings suggest that by
28 altering throughfall chemistry and soil N:P, budworms could influence elemental export from
29 watersheds.
30
31 Keywords: herbivory, coniferous forest, outbreak insect
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32 Introduction
33 Western spruce budworm (Christoneura freemani, hereafter WSB) is the most widely
34 distributed and destructive defoliator in western North American coniferous forests and a major
35 agent of forest disturbance (Fellin and Dewey 1982). In Washington State, insect forest damage
36 affected ~222,500 hectares in 2014, killing approximately 2.4 million trees by 2016, with nearly
37 all defoliation (~40,500 hectares) caused by WSB (WADNR 2016). The area affected by insects
38 is often similar to that burned by wildfires (156,000 hectares in 2014), underscoring the role of
39 insects in forest change.
40 Leaf-feeding canopy herbivores, such as WSB, influence the quantity and chemical
41 composition of water delivered to the soil in throughfall (water that falls through the canopy to
42 the soil; Stadler et al. 2001). First, throughDraft partial or complete defoliation, herbivores often
43 increase throughfall water flux (Michalzik 2011). Second, by fragmenting and damaging
44 foliage, herbivores stimulate leaching of organic and inorganic solutes into throughfall
45 (Michalzik 2011). Third, canopy herbivores deposit frass to canopies and soils during feeding
46 (Hunter 2001). Carbon-(C-) and nitrogen-(N-) rich frass in canopies and the litter layer can be
47 readily leached during the first seasonal rains (Hollinger 1986), increasing dissolved organic C
48 (DOC) and N fluxes into soils (Michalzik 2011). Increased DOC fluxes have, in turn, been
49 shown to fuel leaching of DOC and dissolved organic nitrogen (DON) from the forest floor
50 (Michalzik et al. 2001) and to promote microbial immobilization of N and phosphorus (P)
51 (Michalzik and Stadler 2005). DOC and frass-derived nutrients have also been found to
52 accelerate soil N transformation rates (Huber 2005). For example, in coniferous forests with
53 pine bark beetles, greater inorganic soil N levels after outbreak contributed to more rapid N
54 mineralization and nitrification (Griffin and Turner 2012). Reduced nutrient uptake by
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55 weakened trees (Frost and Hunter 2007) and tree death in stands experiencing herbivory can lead
56 to increased soil N and P as well via reduced assimilatory uptake and increased litter
57 decomposition (Mikkelson et al 2013). Both increased transformation rates and reduced nutrient
58 uptake in forests experiencing outbreak conditions have the potential to increase nutrient
59 leaching losses as much as 30 fold (Houle et al. 2009). While it is clear that herbivores can alter
60 throughfall chemistry and nutrient cycling in forest soils, ecosystem responses are highly
61 variable, and the magnitude and direction of herbivore impact is dependent on the system and the
62 environmental context (Hunter 2001).
63 Despite the persistent role of WSB as a disturbance agent in western North America, we
64 know of no research that has investigated WSB ecosystem effects in seasonally dry coniferous
65 forests in this region. WSB is a native, Draftoutbreak lepidopteran whose larvae feed mostly upon
66 freshly grown needles of Douglas fir and grand fir trees (Alfaro 2014). As an endemic
67 defoliator, WSB always exists at background levels (i.e., during non-outbreak years). However,
68 during outbreak years, these herbivores can reach densities high enough to defoliate tree crowns
69 within a season and to completely strip trees of their needles during a multi-season outbreak
70 (Zhao 2014). Dendrochronological analysis of the past three centuries shows that historic
71 outbreaks occur at the end of regional droughts, last up to ten years, and tend to be synchronized
72 across broad areas (Flower et al. 2014).
73 The WSB life cycle is tied closely to the seasonality of western coniferous forests (Nealis
74 2012). With warming spring temperatures in mid- to late-May, budworm larvae emerge from
75 hibernacula about two weeks prior to budburst and begin dispersing through the canopy. After
76 budburst in late May to mid-June, the larvae begin mining into buds and new needles mostly at
77 the top of the tree crown and on the fringes of branches. Needles begin to turn an orange color
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78 as the larvae feed and grow. After 30-40 days of continuous feeding in June and July, sixth
79 instar larvae construct a pupal case on the underside of the branches from which adults emerge
80 after about 2 weeks in mid-July to early August. Adults disperse to mate and lay eggs, dying
81 shortly thereafter. Larvae emerge from eggs after about 10 days whereupon they immediately
82 seek shelter in the bark by building hibernacula to protect them over the winter.
83 Coupled with decades of fire suppression that have increased forest basal area and
84 density, projections of an ever-warming climate suggest that drought stress in overstocked
85 western coniferous forests (Dalton et al. 2013) will further amplify favorable conditions for
86 WSB. Future WSB outbreaks are predicted to increase in frequency, intensity, and spatial extent
87 (Bentz et al. 2010), with unknown ecosystem effects. Given these projected changes in WSB
88 disturbance, we quantified throughfall andDraft soil chemistry under canopies with high and
89 background levels of WSB herbivory. Specifically, we hypothesized that net throughfall N, P,
90 and DOC fluxes would be higher under canopies with high compared to background levels of
91 WSB herbivory due to increased throughfall water fluxes and accelerated canopy leaching. We
92 also hypothesized that soil inorganic N and P concentrations would mirror the patterns observed
93 in net throughfall. We anticipate that these findings will shed light on how changes in WSB
94 populations could affect nutrient cycling in western coniferous forests under future climate
95 change scenarios.
96
97 Study Area and Site Selection
98 We conducted this study in the Okanogan-Wenatchee National Forest and the Teanaway
99 Community Forest, both located in the rain shadow (east slope) of the Cascade Range in central
100 Washington State. This region is characterized by a continental climate with dry summers (May-
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101 September) and wet winters (October-April). Long term climate data from Blewett Pass, a
102 weather station about 4 km from our study sites, shows that mean annual precipitation is 894 mm
103 with ~86% falling between October and April, mostly as snow from November through March
104 (National Oceanic and Atmospheric Administration (NOAA) 2018). Although peak
105 precipitation is concentrated between October and April, total precipitation at any individual site
106 in this mountainous region can vary considerably by aspect and elevation. Mean monthly
107 temperature is -1.8°C during the winter (Dec, Jan, Feb) and 15.2°C during the summer (Jun, Jul,
108 Aug) (NOAA 2018).
109 Due to the seasonal summer drought, the Cascade Range is dominated by conifer
110 forests, but the varied aspect and elevation of this mountainous landscape generates transitional
111 ecotones characterized by diverse speciesDraft assemblages (Omernick 1987). Our study area is
112 dominated by Douglas fir (Pseudostuga menzeiseii), grand fir (Abies grandis), and ponderosa
113 pine (Pinus ponderosa) with lesser amounts of western larch (Larix occidentalis) and lodgepole
114 pine (Pinus contorta) depending on microclimate.
115 The budworm outbreak we studied began in approximately 2005 and decreased to near
116 background levels by 2017 (Department of Natural Resources (DNR) 2018). Thus, our study,
117 from 2015 to 2016, captured the declining phase of the outbreak. Forest management agencies
118 qualitatively characterize outbreak by estimating defoliation during aerial and ground detection
119 surveys (USFS 1999), and they predict moderate defoliation in the coming year if 35 or more
120 adults are collected in pheromone traps (Cory et al. 1982) at the end of the growing season. We
121 used these projections to select sites in two watersheds with predicted high (Swauk Creek,
122 Okanogan-Wenatchee National Forest) and background (North Fork Teanaway River, Teanaway
123 Community Forest) levels of budworm canopy herbivory (Figure 1). At each site, tree damage
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124 was determined in the field by visually assessing degree of tree crown defoliation as well as the
125 prevalence of discolored dead needles and actively feeding larvae. Herbivory levels were then
126 confirmed by counting dispersing adults captured in pheromone traps after the feeding season in
127 2015 and examining Insect and Disease Detection Survey Annual Summary Maps. The latter
128 data showed no evidence of defoliation by WSB in the background herbivory watershed, while
129 there were many areas with high (>50% of susceptible foliage in polygon defoliated) and some
130 areas with low (≤50% of susceptible foliage in polygon defoliated) defoliation in the high
131 herbivory watershed (Figure 1).
132 Aside from herbivory level, the watersheds were comparable with respect to topography
133 and precipitation, although the high herbivory watershed was slightly higher in elevation and
134 drier (Table 1). In addition, the sites wereDraft relatively well interspersed. Sites in the background
135 herbivory watershed were located 3-6 km distant from each other, with the exception of
136 Moonbeam and Jack, which were <1 km from each other. Sites in the high herbivory watershed
137 were 3-8 km distant from each other.
138
139 Methods
140 Throughfall Sampling
141 We measured throughfall water, nutrient, and DOC fluxes in the study watersheds. On
142 25 June 2015, we established four throughfall sampling sites in the high herbivory watershed and
143 four sites in the background herbivory watershed. In each of the eight sites, three throughfall
144 collectors were installed (n=24 throughfall collectors total). Throughfall collectors remained
145 deployed from 25 June to 8 November 2015 and from 1 May through 19 September 2016.
146 Collectors were removed during winter due to snowpack and inability to access the sites.
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147 Throughfall collectors were constructed after Ponette-González et al. (2010). Each
148 throughfall collector consisted of a 20-cm diameter polyethylene funnel (324 cm2) securely
149 attached atop a 1-m length cellular core pipe hammered into the ground. Tygon tubing
150 connected to the bottom of the funnel was woven through a small hole drilled into the side of the
151 core pipe and connected to a heavy-duty polyethylene 4 L container, which was secured to the
152 pipe with wire. The tubing was positioned inside the lip of the collection container with a wrap
153 of parafilm so that throughfall collected in the funnel could drip securely into the container
154 below. A small polywool filter was placed within the funnel to prevent debris from entering the
155 samples and from inhibiting water flow. In each site, collectors were established below groups
156 of individual trees of the species P. menzeisii or A. grandis with field-verified presence or
157 absence of budworm activity. One throughfallDraft collector was placed beneath each tree, half way
158 between the trunk and the dripline. In addition, two bulk rainfall collectors (collectors that
159 remain open between sampling) were established in each study watershed in an open area
160 without canopy cover (n=4 bulk collectors).
161 Throughfall and rainfall were collected on an event basis, where an event was defined as
162 rainfall sufficient to produce 40 mL of throughfall. The first rainfall event with sufficient
163 volume for analysis was 11 September 2015; after this date samples were collected to first
164 snowfall on 8 November 2015 (n=4 events). After spring snowmelt, event-based sampling
165 resumed on 8 May 2016 and continued until 19 September 2016 (n=6 events). Thus, our
166 sampling captured 10 rainfall events over the course of the two deployment seasons (Figure 2).
167 Sample collection occurred within 48 hours of each rain event. During each collection, we
168 poured throughfall or rainfall into a 1 L graduated cylinder previously rinsed with deionized
169 water to measure total water volume. Each sample was then transferred to an acid washed 1 L
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170 HDPE bottle and stored on ice in a cooler for transport to the lab at Central Washington
171 University. After collection, we rinsed throughfall and rainfall samplers with deionized water,
172 replaced the polywool, and applied new parafilm to seal the 4-L collection jug. Upon return to
173 the lab, samples were refrigerated at 4°C and filtered through a Pall AE GFF, 1 µm pore size
174 filter within 24 hours of collection. We retained at most 240 mL of filtered sample from each
175 site, and those subsamples were frozen until water chemistry analysis could be performed.
176 Soil Sampling
177 We also collected one composite soil sample (n=7 sample events) within a 10 m radius of
178 each throughfall collector approximately every two months during the study period, but not
179 during winter due to snowpack (n=1 composite sample per throughfall collector, n=3 throughfall
180 collectors per site). For each soil sample,Draft we removed the O horizon and used a metal 1 cm
181 diameter soil corer to collect three cores of no more than 15 cm depth from the A horizon. For
182 each throughfall sampler, these three soil cores were collected into a single composite sample in
183 a clean Ziploc bag for transport to the laboratory. Given that we could not process soil samples
184 within two days of collection (Hart et al. 1994), samples were frozen (-12° C) until preparation
185 for soil extraction following the recommendation of Hart and Firestone (1989). To prepare the
186 soil samples for extraction, large pieces of organic matter and gravel were removed, and the
187 composite sample taken from near each throughfall collector was homogenized by sieving
188 through a 2 mm mesh. Two 10 g subsamples were taken from each soil sample, one for
+ - 189 ammonium (NH4 -N) and nitrate (NO3 -N) extraction using 2M KCl solution (Keeney and
190 Nelson 1987) and one for inorganic P extraction using a Bray P1 solution (Bray and Kurtz 1945).
191 We added 75 mL of each extraction solution to make two separate slurries, which were then
192 shaken on a rotary shaking table for 1 hour. After shaking, samples were centrifuged at 4000
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193 rpm for 10 minutes, and the supernatant was filtered (Pall AE GFF, 1 µm pore size) and frozen
194 until analysis could be performed.
195 Canopy Leaching Experiment
196 In summer 2016, we deployed additional samplers to test the relative importance of
197 leaching versus dry deposition to throughfall P concentrations. For this experiment, we
198 constructed artificial trees, consisting of a 2 m length of 19 mm x 38 mm lumber with 2-20 cm
199 lengths of artificial coniferous trees attached through the upper 20 cm of the lumber. The
200 artificial coniferous trees were leached in nanopure water (18.2 MΩ) for 24 h and rinsed with
201 nanopure water prior to construction. The artificial trees acted as a control for atmospheric
202 deposition of P whereas live trees would yield P from atmospheric deposition and from biotic
203 leaching (Runyan et al. 2013). On 1 MayDraft 2016, artificial trees were placed at each replicate
204 location within each of the eight throughfall sampling sites described above, anchored by a
205 plastic cable tied to a piece of rebar driven into the ground. We collected the first sample on 8
206 July after 18 dry days. We collected the second sample on 20 August 2016 after 31 dry days. To
207 collect the sample, we removed a length (about 10 cm) of artificial tree and a length of a branch
208 from a nearby grand fir tree, each of which would fit into an acid-washed 50 mL centrifuge tube.
209 Upon return to the laboratory, we immediately added 25 mL of nanopure water to each
210 centrifuge tube and shook the sample for 45 minutes to rinse and leach any compounds present
211 (Runyan et al. 2013). Subsequently, samples were filtered (Pall AE 1 µm nominal pore size) and
212 frozen for later inorganic P analysis. Phosphorus concentration was normalized by dry mass of
213 artificial or live tree (mg P/mg DM) for comparison.
214 Chemical Analyses
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+ 215 For all throughfall, rainfall, and soil samples we measured nitrogen as ammonium (NH4 -
- - - 3- 216 N) and nitrate + nitrite (NO3 -N + NO2 -N; hereafter just NO3 -N), and orthophosphate (PO4 -P),
217 and for throughfall and rainfall samples, we also measured dissolved organic carbon (DOC). For
3- 218 the canopy leaching experiment, we only measured PO4 -P. We used the phenol hypochlorite
+ 219 method to measure NH4 -N (Solorzano 1969), the cadmium reduction method (Environmental
- 220 Protection Agency (EPA) 1993) to measure NO3 -N, and the ascorbic acid method to measure
3- 221 PO4 -P (Murphy and Riley 1962); all of these nutrients were measured using a discrete-sample
222 water analyzer (Seal AQ1, Seal Analytical; Mequon, Wisconsin, USA) with EPA equivalent
223 methods. Soil samples were run separately from throughfall and rainfall samples with standards
224 using the appropriate extraction matrix (KCl or Bray P1) to account for any possible changes in
225 absorbance due to the different soil extractionDraft solutions used for N and P extraction. For DOC,
226 samples were acidified to pH < 2.0 to purge inorganic carbon before measuring via infrared
227 methods (American Public Health Association (APHA) 1995) on a total organic carbon analyzer
228 (TOC-L Total Organic Carbon Analyzer, Shimadzu; Kyoto, Japan).
229 Flux Calculations and Statistical Analysis
230 For each sample event, volume-weighted mean (VWM) throughfall concentrations were
231 calculated per site using the formula:
232
233 ∑(푐표푛푐푖 ∗ 푝푟푒푐푖푝푖)/∑푝푟푒푐푖푝푖
234
235 where conc is the solute concentration (mg/L), precip (L) is the sample volume, and i is the
236 collector.
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+ - 237 Bulk deposition and throughfall dissolved NH4 -N and NO3 -N (hereafter dissolved
3- 238 inorganic N; DIN), PO4 -P, and DOC fluxes for each collector and sample event were calculated
239 by multiplying concentration (mg/L) by water volume (L) and incorporating the surface area of
240 the collector (324 cm2). Deposition and fluxes were expressed in kg/ha. Net throughfall water,
241 nutrient, and organic carbon flux (NTF) were calculated as:
242
243 NTF = TF – BR
244
245 where TF is throughfall water (mm) or throughfall nutrient or organic carbon flux (kg/ha) and
246 BR is bulk rainfall (mm) or bulk rainfall deposition (kg/ha). We considered negative net
247 throughfall water flux (NTF < 0) to indicateDraft canopy water interception. For nutrient and organic
248 carbon fluxes, negative net throughfall flux indicates greater canopy uptake than the sum of dry
249 deposition and canopy leaching, whereas positive net throughfall flux (NTF > 0) indicates
250 greater dry deposition and canopy leaching than canopy uptake. Net throughfall cannot separate
251 the contribution of dry deposition from that of canopy herbivory to fluxes but represents an
252 integrated measure of these processes.
253 We analyzed differences in throughfall and net throughfall water, nutrient, and organic
254 carbon fluxes, and soil chemistry using linear mixed effects models (function lme in R package
255 lme4) with alternate variance structures when appropriate (Zuur et al. 2009). Each model had
256 two interacting main effects, budworm herbivory level (n=2, high versus background) and
257 sample event (n=10) as a factor, and a random effect of each throughfall sampler nested within
258 site (Table 2). Generally, the models required log normalization so that residuals met model
259 assumptions. If we found significant main effects and/or interactions, we performed pairwise
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260 analysis of estimated marginal means (functions emmeans and pairs in R package emmeans) to
261 describe an effect size of how herbivory influenced the response variable on a per sample event
262 basis. A post-hoc power analysis (pwr.f2.test function from the R package pwr) indicated that
263 our models had enough power (≥0.80) to detect a medium (>0.15 Cohen’s f2) effect size in the
264 interaction term, and most of our models had significant interactions. We analyzed the canopy
265 leaching experiment using a two-factor linear model (budworm = high or background; tree type
266 = artificial or real) with no interaction term followed by Tukey’s Honestly Significant
267 Differences test. Bulk deposition and throughfall nutrient and organic carbon fluxes were
268 summed for the entire study period and annualized by dividing by the total number of sample
269 days (n=277) and then multiplying by 365. All statistical analyses were conducted using R 3.5.1
270 (R Core Team, 2018) with α = 0.05, butDraft we additionally report pairwise comparisons of
271 estimated marginal means with α = 0.10.
272
273 Results
274 Budworm effects on water fluxes
275 Due to a significant interaction, throughfall water flux differed between budworm
276 herbivory levels during one sample event only (Table 2; LME, p<0.0001). On 21Jul16, a highly
277 localized rainstorm caused sites with background budworm herbivory to have elevated
278 throughfall water flux (Figure 3A). Net throughfall water fluxes were mostly negative (i.e.,
279 canopy water interception) and a significant herbivory level x sample event interaction effect
280 (LME, p<0.0001) indicated that budworm effects also differed among sample events (Figure 3B,
281 Table 2). There was no consistent pattern, however. There was more canopy interception in the
282 high budworm sites during two events (one post feeding on 11Sep15 and one during feeding on
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283 13Jul16) but more canopy interception in the background budworm sites during one event
284 (29Oct15). During large storm events (>30 mm rainfall), canopy water interception was similar
285 at high and background herbivory sites, indicating that differences in throughfall between
286 budworm herbivory levels during these storms were driven by greater rainfall rather than by
287 changes in canopy leaf area.
288 Budworm effects on throughfall concentrations
3- 289 Only throughfall VWM PO4 -P concentrations exhibited significant differences between
3- 290 levels of budworm herbivory (LME, p=0.0014), but VWM DIN, PO4 -P, and DOC
291 concentrations all had significant interactions between budworm herbivory level and sample
292 event (Figure 4, Table 2; LME, p=0.01, p=0.0001, p=0.0002 respectively). Throughfall DIN
293 concentrations were elevated at the highDraft budworm sites in the first rainfall after budworm
294 feeding in 2015 (i.e., 11Sep15) and during and after feeding in 2016 (21Jun16 to 19Sep16), but
295 they were higher in the background budworm herbivory sites slightly before or at the beginning
3- 296 of feeding on 8May16 (Figure 4A). Throughfall PO4 -P was higher in the high budworm
297 herbivory sites after feeding on 15Sep15 and during budworm feeding on 13Jul16 and 21Jul16
298 (Figure 4B). Concentrations of DOC in throughfall were elevated at the high herbivory sites on
299 8Nov15 and 13Jul16 (Figure 4C).
300 Budworm effects on throughfall nutrient fluxes to soil
301 Differences in throughfall fluxes between sites with high and background herbivory were
302 driven by significant interactions (Table 2) between herbivory level and sample event for DIN
3- 303 (LME, p<0.0001), PO4 -P (LME, p=0.0058), and DOC (LME, p=0.0004). Throughfall DIN
304 fluxes were generally higher at the high than at the background budworm sites during post
305 feeding in 2015 as well as during peak and near the end of feeding in 2016 (Figure 5A). High
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3- 306 budworm sites similarly showed a greater PO4 -P fluxes during and after the 2016 feeding
307 period (Figure 5B). There was no clear pattern for DOC, with higher fluxes in the high
308 budworm sites on 8Nov15 and in the background budworm sites on 11Oct 15 and 21Jul15
309 (Figure 5C).
310 Net throughfall nutrient fluxes
311 Net throughfall DIN fluxes were mostly negative in high and background budworm sites,
312 indicating greater canopy uptake of DIN than inputs via dry deposition and canopy leaching
3- 313 (Figure 6A). In contrast to N, net throughfall PO4 -P and DOC fluxes were generally positive,
314 indicating greater dry deposition and canopy leaching of these nutrients than net uptake by the
315 canopy (Figures 6B and 6C). We found significant differences in net throughfall DIN (LME,
3- 316 p<0.0001) and PO4 -P (LME, p=0.0001)Draft fluxes between levels of canopy budworm herbivory,
317 but again, this pattern was driven by significant interactions between budworm level and sample
3- 318 event (LME, p<0.0001 for DIN; p=0.03 for PO4 -P). Net throughfall DOC fluxes also exhibited
319 a significant interaction effect (LME, p=0.0004; Table 3). Differences in net DIN uptake
320 between sites with high and background herbivory were especially pronounced during the 2016
3- 3- 321 feeding period (8May16-21Jul16). Net PO4 -P showed the same patterns as throughfall PO4 -P
322 with generally elevated net P fluxes throughout the 2016 sample period and significantly higher
323 fluxes during two samples events, one at the beginning of and one during peak feeding. For net
324 DOC, there were two pulses during large rain storms (8Nov15 and 21Jul16).
325 Dry P deposition versus canopy leaching
3- 326 The average amount of PO4 -P collected from real trees was more than 30-fold higher
327 than that collected from artificial trees (Figure 7; ANOVA, p<0.0001). Assuming that artificial
328 trees collected only dry deposition, this indicates significant P leaching from real trees affected
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329 by canopy budworm herbivory. Although real and artificial trees in high budworm sites released
330 more phosphorus than those in background budworm sites (ANOVA, p<0.0001), the difference
331 between real and artificial trees was more pronounced at the high budworm sites. This suggests
332 greater relative P leaching under high herbivory.
333 Soil nutrients
+ - 334 Neither soil NH4 -N nor NO3 -N differed by budworm herbivory level (LME, p>0.05),
335 but both solutes varied by sample event (Figure 7, Table 2; LME, p<0.05 for both solutes), with
- 336 greater soil NO3 -N concentrations found on the last sample date of each year. Throughout the
+ - 337 study, NH4 -N was nearly five times higher than NO3 -N, which was near detection level (10 mg
- -1 338 NO3 -N L ) until the last sample event in November 2016. In contrast, soil inorganic P averaged
339 about three times higher in the high comparedDraft to the background budworm sites throughout the
340 study (LME, p=0.0005) and varied by date (LME p=0.04). There were no significant
341 interactions for soil nutrients. The relatively low inorganic N and high inorganic P led to very
342 low soil molar N:P ratios in all the study sites, with significantly lower N:P ratios in sites with
343 high budworm herbivory (data not shown, LME p<0.0001).
344
345 Discussion
346 Feeding activity enhances net throughfall N and P fluxes
347 Our findings show that WSB herbivory led to elevated net throughfall N and P fluxes
348 under Douglas fir and grand fir stands in Central Washington at the end, and during the declining
349 stages, of a multi-year (2005-2016 state-wide) outbreak cycle. In contrast to our hypothesis,
350 canopy water interception and, in turn, the amount of water delivered to the forest floor as
351 throughfall, did not differ consistently between high and background budworm sites, or during
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352 large storm events when dissimilarities in leaf area should exacerbate absolute differences in
353 canopy water interception (Ponette-González et al. 2010). It is possible that our visual
354 observations and the coarse aerial/ground detection survey data overestimated defoliation levels
355 in the high budworm herbivory sites, in which case we would expect differences in water flux
356 between Douglas fir stands with high and background herbivory to be minor (Schowalter 1999)
357 or undetectable. Compensatory growth following defoliation (Piene and Eveleigh 1996) and/or
358 enhanced needle longevity (Doran et al. 2017) could also explain the lack of difference in canopy
359 water interception between watersheds with high and background herbivory.
360 In contrast to water fluxes, we found that DIN concentrations and net DIN fluxes
361 increased during and/or immediately after larval stage budworm feeding. Tree canopies at the
362 high budworm herbivory sites retained substantiallyDraft less incoming DIN (67%) compared to
363 background budworm sites (85%). In fact, the high budworm sites showed declining DIN
364 retention over the course of the 2016 feeding season, approaching nearly zero net retention at the
365 height of the feeding period. These results indicate a shift to lower uptake and/or greater
366 leaching of N from the canopy at the high herbivory sites. That this pattern of decreasing DIN
367 was not mirrored at the background herbivory sites further supports the premise that budworms
368 accelerated canopy N cycling. Similar to our findings, le Mellec and Michalzik (2008) found
369 that canopy herbivory increased throughfall total N flux to the forest floor of infested plots and
370 that differences between infested and uninfested plots were most pronounced during peak
371 feeding.
372 To our surprise, net throughfall DOC fluxes did not increase with herbivory as we
373 hypothesized and as reported in previous studies (IM-Arnold et al. 2016). We recorded two
374 pulses of net throughfall DOC, one on 8 November 2015 at sites with high and one on 21 July
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375 2016 at sites with background herbivory, respectively. Because these were also the largest rain
376 storms sampled, we speculate that rainfall-driven canopy leaching was a relatively more
377 important driver of DOC fluxes than the effects of budworm activity during the study period. A
378 caveat of this interpretation is that only four sites were sampled within each treatment and, in
379 some cases, there was considerable measurement variability. Detecting the impacts of endemic
380 densities of herbivores on nutrient cycling can be challenging (Hunter et al. 2003).
381 Defoliator abundance and canopy damage are critical because of the mechanisms by
382 which herbivorous insects alter nutrient fluxes from canopies to soils. In cases where water flux
383 is unaffected, herbivores modify throughfall chemical concentrations by accelerating canopy
384 leaching and depositing frass that can be leached into water droplets (Hunter 2001). While the
385 impacts of folivores on C and N fluxes areDraft relatively well studied (e.g., Lovett et al. 2002), few
386 studies investigate the influence of leaf-feeding insects on throughfall P, a critical limiting
387 nutrient in many terrestrial ecosystems (Vitousek et al. 2010). Those that have show mixed
388 results. Seastedt et al. (1983) reported a minor increase in net throughfall P under black locust
389 trees affected by herbivory, while Hollinger (1986) found that P fluxes declined from pre-
390 outbreak to outbreak conditions in California. In contrast, our high budworm sites exhibited
391 significantly higher net P fluxes during peak budworm feeding. These mixed results could be
392 due to a variety of factors, including differences in defoliation level, precipitation regime, and
393 the insect-plant system studied. Seastedt et al. examined throughfall chemistry responses to low
394 (i.e., nominal) levels of defoliation, which likely contributed to the minor levels of canopy P
395 leaching they observed. In California, Hollinger suggested that the lack of P leaching was likely
396 the result of the seasonal nature of precipitation. Another possibility is that our high budworm
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397 herbivory sites had higher foliar P concentrations (Kemp and Moody 1984)––possibly due to
398 defoliation (Piene 1980)––and as a result higher P leaching rates.
399 On an annualized basis, high budworm sites received nearly two times as much P in
400 throughfall (2174 g P ha-1 yr-1) as background budworm sites (1357 g P ha-1 yr-1) (Table 3). Our
401 experiment showed that dry P deposition to artificial trees was minimal, albeit 3-fold greater in
402 the high budworm herbivory watershed. This is most likely due to the higher elevation and drier
403 precipitation regime at the high compared to the background budworm watershed.
404 Notwithstanding these differences in dry deposition, we measured a 30-fold increase in P rinsate
405 from real trees and a 4-fold difference in P rinsate between real trees at the high and background
406 sites. These results support our hypothesis for accelerated P leaching from damaged foliage
407 and/or dissolution of frass into water at Draftthe high budworm herbivory sites.
408 Lower soil N:P ratios under stands with high budworm herbivory
409 With the exception of urban areas, the Pacific Northwest region generally receives very
410 low DIN in wet deposition relative to other parts of the US (National Atmospheric Deposition
411 Program (NADP) 2018). We measured bulk DIN deposition rates of <500 g N ha-1 yr-1,
412 corresponding to the lowest category for NADP in 2015 and 2016 (NADP 2018) when this study
413 was conducted. Our data show that budworms can influence canopy N dynamics by making
414 canopies less retentive, but the overall DIN fluxes to the forest floor are still low at just 67-71 g
415 N ha-1 yr-1 (Table 3). Consistent with low wet atmospheric N deposition and minimal inputs via
+ - 416 throughfall, we measured very low soil NH4 -N and NO3 -N pools overall with no clear link to
417 budworm activity or seasonality. Pacific Northwest Douglas fir forests like our study system
418 have systemic N limitation whereby trees respond more strongly to N inputs compared to
419 moisture despite the dry summers (Gessel et al. 1990). Although we sampled the end of a
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420 budworm outbreak, climate change projections anticipate longer and more frequent insect
421 outbreaks. Under these projected scenarios, budworm herbivory could drive even greater fluxes
422 of N to soils than we observed, which could fundamentally change the N limitation status of
423 Pacific Northwest forests where budworms are active.
424 Similar to N in rainfall, P in rainfall was very low at <150 g P ha-1 yr-1, within the values
425 reported by Tipping et al. (2014) for North America. In contrast to DIN though, we observed
426 high inorganic P inputs to soils in sites with high budworm activity during two events, and soils
427 in high budworm sites had significantly more P than soils in background budworm sites
428 throughout the study. The significantly higher inorganic P in soils of high budworm sites
429 resulted in very low molar N:P ratios (<1 for all dates and times, data not shown). N:P ratios for
430 soil (13:1) and soil microbes (7:1) tend Draftto be well constrained at a global level (Cleveland and
431 Liptzin 2007), thus budworm herbivory could intensify the potential for soil N limitation via
432 excess P addition despite less N retention in the canopy, changing patterns of terrestrial N
433 fixation similar to changes that occur during succession following disturbance (Vitousek et al.
434 2013). A longer growing season expected under future climate change scenarios could further
435 intensify demand for N. Lower N:P ratios of soils during this budworm outbreak implies that
436 soil microbes could quickly immobilize excess N that results from herbivory, possibly
437 decreasing N export from soils to streams (Bengtsson et al. 2003). Like the forests of this region,
438 streams are also considered to be N-limited, in part due to the extirpation or severe truncation of
439 the N subsidy formerly delivered by returning anadromous salmonids (Naiman et al. 2002).
440 Moreover, N limitation implies a limited role for nitrification to transform N to more easily
- - 441 exported NO3 -N, consistent with the very low soil NO3 -N levels we observed. Therefore,
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442 decreasing low soil N:P ratios further via increased budworm herbivory could influence
443 downstream ecosystem function by reducing already low solute export rates.
444 Budworms have an outbreak cycle that lasts multiple years (Flower et al. 2014), and the
445 summers of 2015 and 2016 were the last years of the outbreak cycle we studied (DNR 2018).
446 Nevertheless, we still documented budworm effects on throughfall fluxes during the feeding
447 season and following feeding as rain began to fall. These effects included less N retention in the
448 canopy significantly greater throughfall P fluxes, and lower N:P ratio in soils with high budworm
449 activity. Budworm outbreaks are predicted to become larger and longer lasting due to the
450 interaction between climate change and a legacy of fire suppression (Mote and Salathe 2010).
451 Our results suggest that budworms are likely to play an important role in altering the internal
452 nutrient dynamics of coniferous forests Draftin western North America by translocating nutrients from
453 the canopy to soils during outbreaks, and these effects will magnify under future global change
454 scenarios.
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455 Acknowledgements
456 Sarah Clark, Michael Dallas, Natalie Levesque, Caitlin Moynihan, Emily Schuetz,
457 Stephen Simpson, and Toni Heay Stewart provided assistance in the field and/or in the lab. We
458 are grateful for Sally Entrekin’s comments and the comments of three anonymous reviewers that
459 improved this manuscript. Ali Scoville, Bob Hall, and AJ Reisinger provide used feedback on
460 the presentation of our statistical analyses. This research was supported by funding from the
461 National Science Foundation (DEB 1540679).
Draft
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646
Draft
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647 Tables
648 Table 1: Characteristics of the study sites where throughfall and rainfall were measured in
649 watersheds with high and background levels of western spruce budworm herbivory.
650 2015 2016 Elevation Rainfall Site Aspect Rainfall Rainfall (m asl) (mm)b (mm)a (mm)a Background Jack 963 347 981 918 Jungle 824 54 1067 1016 Moonbeam 973 137 981 918 Standup 903 93 1186 1121 223 Avg 916 158 1054 993 High Blue 1055 188 967 914 Billy Goat 984 Draft158 898 888 Hovey 1050 57 884 831 164 Hurley 978 50 884 831 Avg 1017 113 908 866 651 a Precipitation estimates were derived from National Atmospheric Deposition Program (NADP) 652 total annual precipitation maps; estimates are from the Parameter-elevation Regression on 653 Independent Slopes Model (PRISM). 654 655 b Rainfall measured at Standup and Hovey during the study period from 11 September to 8 656 November 2015 and from 8 May to 19 September 2016 (n=277 days). 657
658
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659 Table 2: Statistical model p-values, numerator degrees of freedom, and denominator degrees of
660 freedom for main effects. Bolded values are significant (p<0.05). TF = throughfall, NTF = net
3- 661 throughfall, VWM = volume weighted mean, DIN = dissolved inorganic nitrogen, PO4 -P =
662 orthophosphate, DOC = dissolved organic carbon.
Budworm Sample Event Interaction Response Variable p-value (numerator df, denominator df) log(TF water flux+1) (mm) 0.6579 (1, 22) <0.0001 (9, 146) <0.0001 (9, 146) net NTF water flux (mm) 0.3474 (1, 22) <0.0001 (9, 140) <0.0001 (9, 140) log(TF VWM DIN+1) (mg L-1) 0.6805 (1, 6) 0.0029 (9, 52) 0.0103 (9, 52) 3 -1 log(TF VWM PO4 P +1) (mg L ) 0.0014 (1, 6) <0.0001 (9, 52) 0.0001 (9, 52) TFVWM DOC (mg L-1) 0.4310 (1, 6) <0.0001 (9, 52) 0.0002 (9, 52) log(TF DIN flux+1) (kg ha-1) 0.7832 (1, 22) <0.0001 (9, 141) <0.0001 (9, 141) 3- -1 log(TF PO4 -P flux+1) (kg ha ) 0.0009 (1, 22) <0.0001 (9, 142) 0.0058 (9, 142) log(TF DOC flux+1) (kg ha-1) 0.7171 (1, 22) <0.0001 (9, 140) 0.0004 (9, 140) log(NTF DIN flux+1) (kg ha-1) <0.0001 (1, 22) <0.0001 (9, 141) <0.0001 (9, 141) 3- -1 log(NTF PO4 -P flux+1) (kg ha ) Draft0.0001 (1, 22) <0.0001 (9, 142) 0.0319, (9, 142) log(NTF DOC flux+1) (kg ha-1) 0.4459 (1, 22) <0.0001 (9, 140) 0.0004 (9, 140) + -1 log(Soil NH4 -N+1) (mg g ) 0.7365 (1, 22) 0.0001 (6, 121) 0.9597 (6, 121) - -1 log(Soil NO3 -N+1) (mg g ) 0.6987 (1, 22) 0.0143 (6, 100) 0.0507 (6, 100) 3- -1 Soil PO4 -P (mg g ) 0.0005 (1, 22) 0.0404 (6, 122) 0.0528 (6, 122) 663 664
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666 Table 3: Annualized bulk deposition (mean ± SE) and throughfall flux for background and high
667 budworm sites measured from 11 September to 8 November 2015 and from 8 May to 19
3- 668 September 2016 (n=277 days). DIN = dissolved inorganic nitrogen, PO4 -P = orthophosphate,
669 DOC = dissolved organic carbon
Background High Bulk Deposition (g ha-1 yr-1) + NH4 -N 258 ± 69 100 ± 8.4 - NO3 -N 187 ± 44 111 ± 59 DIN 445 ± 26 211 ± 67 3- PO4 -P 107 ± 74 144 ± 68 DOC 18231 ± 1979 18128 ± 11173 Throughfall Flux (g ha-1 yr-1) + NH4 -N 35 ± 5.4 38 ± 5.6 - NO3 -N 30 ± 2.7 34 ± 2.8 DIN 65 ± 4.4 71 ± 5.0 3- PO4 -P Draft1357 ± 417 2174 ± 646 DOC 105298 ± 8205 158587 ± 32689 670
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672 Figure Captions
673 Figure 1 Watersheds with high and background levels of western spruce budworm herbivory as
674 determined by annual detection surveys (https://www.fs.fed.us/foresthealth/applied-
675 sciences/mapping-reporting/gis-spatial-analysis/detection-surveys.shtml). Throughfall water and
676 chemistry were sampled at four sites within each watershed. Inset map: State of Washington with
677 location of study watersheds (in black box).
678
679 Figure 2 Accumulated precipitation at Blewett Pass Snotel site. Shaded area represents time
680 when the throughfall and bulk precipitation collectors were deployed. Small arrows indicate rain
681 events that we sampled.
682 Draft
683 Figure 3 (A) Estimated marginal means (+1 standard error) of throughfall water flux (water that
684 falls through the canopy to the soil) and (B) net throughfall water flux (throughfall – rainfall) in
685 sites with high and background levels of western spruce budworm herbivory. ’*’ (p<0.05) and
686 ‘+’ (p<0.10) indicate significant pairwise differences between high and low budworm activity
687 per sample event. Bar on top of figure indicates time of budworm feeding (black=feeding, grey=
688 not feeding). The gap between 8Nov15 and 8May16 indicates when samples were not collected
689 due to winter snowpack.
690
691 Figure 4 (A) Estimated marginal means (+1 standard error) of throughfall volume weighted
692 mean (VWM) dissolved inorganic nitrogen (DIN) concentration, (B) throughfall VWM
3- 693 orthophosphate (PO4 -P) concentration, and (C) throughfall VWM dissolved organic carbon
694 (DOC) concentration in sites with high and background levels of western spruce budworm
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695 herbivory. ’*’ (p<0.05) and ‘+’ (p<0.10) indicate significant pairwise differences between high
696 and low budworm activity per sample event. Bar on top of figure indicates time of budworm
697 feeding (black=feeding, grey= not feeding). The gap between 8Nov15 and 8May16 indicates
698 when samples were not collected due to winter snowpack.
699
700 Figure 5 (A) Estimated marginal means (+1 standard error) of throughfall dissolved inorganic
3- 701 nitrogen (DIN) flux, (B) throughfall orthophosphate (PO4 -P) flux, and (C) throughfall dissolved
702 organic carbon (DOC) flux in sites with high and background levels of western spruce budworm
703 herbivory. ’*’ (p<0.05) and ‘+’ (p<0.10) indicate significant pairwise differences between high
704 and low budworm activity per sample event. Bar on top of figure indicates time of budworm
705 feeding (black=feeding, grey= not feeding).Draft The gap between 8Nov15 and 8May16 indicates
706 when samples were not collected due to winter snowpack.
707
708 Figure 6 (A) Estimated marginal means (+1 standard error) of net throughfall (NTF) dissolved
3- 709 inorganic nitrogen (DIN) flux, (B) net throughfall orthophosphate (PO4 -P) flux, and (C) net
710 throughfall dissolved organic carbon (DOC) flux in sites with high and background levels of
711 western spruce budworm herbivory. Different letters indicate sample events that are
712 significantly different Bar on top of figure indicates time of budworm feeding (black=feeding,
713 grey= not feeding). The gap between 8Nov15 and 8May16 indicates when samples were not
714 collected due to winter snowpack.
715
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3- 716 Figure 7 Orthophosphate (PO4 -P) leached from real and artificial trees in plots with high and
717 background levels of western spruce budworm herbivory. Different letters indicate treatments
718 that are significantly different.
719
+ 720 Figure 8 (A) Estimated marginal means (+1 standard error) of soil ammonium (NH4 -N), (B) soil
- 3- 721 nitrate (NO3 -N), and (C) orthophosphate (PO4 -P) in sites with high and background levels of
722 western spruce budworm herbivory. ’*’ (p<0.05) and ‘+’ (p<0.10) indicate significant pairwise
723 differences between high and low budworm activity per sample event Bar on top of figure
724 indicates time of budworm feeding (black=feeding, grey= not feeding). The gap between Nov
725 15 and Apr 16 indicates when samples were not collected due to winter snowpack.
726 Draft
727
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728 Figures
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732 Figure 2
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735 Figure 3
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738 Figure 4
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741 Figure 5
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744 Figure 6
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747 Figure 7
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750 Figure 8
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