Canadian Journal of Forest Research

Western spruce budworm effects on throughfall N, P, C fluxes and soil nutrient status in the

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 , climate change,

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 ( 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

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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462 References

463 Alfaro, R.I., Berg, J., and Axelson, J. 2014. Periodicity of western spruce budworm in Southern

464 , Canada. For. Ecol. and Manage. 315: 72-79.

465 doi:10.1016/j.foreco.2013.12.026.

466 American Public Health Association (APHA). 1995. Standard methods for the examination of

467 water and wastewater. 19th edition. American Public Health Association, American Water

468 Works Association, and Water Environment Federation, Washington, DC.

469 Bengtsson, G., P. Bengtson, and K.F. Mansson. 2003. Gross nitrogen mineralization-,

470 immobilization-, and nitrification rates as a function of soil C/N ratio and microbial

471 activity. Soil Biology and Biochemistry 35(1): 143-154.

472 Bentz, B.J., Régnière, J., Fettig, C.J., Hansen,Draft E.M., Hayes, J.L., Hicke, J.A., Kelsey, R.G.,

473 Negrón, J.F., and Seybold, S.J. 2010. Climate change and bark beetles of the western

474 United States and Canada: direct and indirect effects. BioScience 60(8): 602-613. doi:

475 10.1525/bio.2010.60.8.6.

476 Bray, R.H., Kurtz, L.T. 1945. Determination of total, organic, and available forms of phosphorus

477 in soil. Soil Sci. 59(1): 39-45.

478 Chen, Z., Kolb, T.E., and Clancy, K.M. 2001. Mechanisms of Douglas-fir resistance to western

479 spruce budworm defoliation: bud burst phenology, photosynthetic compensation and

480 growth rate. Tree Physiology 21(16): 1159-1169. doi:10.1093/treephys/21.16.1159.

481 Chen, Z., Kolb, T.E., and Clancy, K.M. 2002. The role of monoterpenes in resistance of Douglas

482 firs to western spruce budworm defoliation. J. Chem. Ecol. 28(5): 897-920.

483 doi:10.1023/A:1015297315104.

23 https://mc06.manuscriptcentral.com/cjfr-pubs Canadian Journal of Forest Research Page 24 of 45

484 Cleveland, C.C., and D. Liptzin. 2007. C:N:P stoichiometry in soil: is there a “Redfield ratio” for

485 the microbial biomass? Biogeochemistry 85(3): 235-252. doi:10.1007/s10533-007-9132-

486 0.

487 Cory, H.T., G. E. Daterman, G. D. Daves, Jr., L.L. Sower, R.F. Shepherd, and C. J. Sanders.

488 1982. Chemistry and field evaluation of the sex pheromone of western spruce budworm,

489 Choristoneura occidentalis Freeman. Journal of Chemical Ecology 8:339-350.

490 Dalton, M.M., Mote, P.W., and Snover, A.K. (Editors.) 2013. Climate change in the Northwest:

491 Implications for our landscapes, waters, and communities. Island Press, Washington D.C.

492 Department of Natural Resources (DNR). 2018. Forest health highlights in Washington–2017.

493 Available from

494 https://www.dnr.wa.gov/publications/rp_fh_2017_forest_health_highlights.pdf.Draft [accessed

495 7 September 2018].

496 Doran, O., MacLean, D.A., and Kershaw Jr., J.A. 2017. Needle longevity of balsam fir is

497 increased by defoliation by spruce budworm. Trees 31: 1933-1944. doi: 10.1007/s00468-

498 017-1597-4.

499 Environmental Protection Agency (EPA). 1993. Determination of nitrate-nitrite nitrogen by

500 automated colorimetery. Method 353.2, Revision 2.0. Environmental Monitoring Systems

501 Laboratory, Office of Research and Development, Cincinnati, Ohio.

502 Fellin, D.G., and Dewey, J.E. 1982. Western spruce budworm. US Department of Agriculture,

503 Forest Service.

504 Flower, A., Gavin, D.G., Heyerdahl, E.K., Parsons, R.A., and Cohn, G.M. 2014. Drought-

505 triggered western spruce budworm outbreaks in the interior Pacific Northwest: A multi-

506 century dendrochronological record. For. Ecol. and Manage. 324: 16-27.

24 https://mc06.manuscriptcentral.com/cjfr-pubs Page 25 of 45 Canadian Journal of Forest Research

507 doi:10.1016/j.foreco.2014.03.042.

508 Frost, C.J., and Hunter, M.D. 2007. Recycling of nitrogen in herbivore feces: plant recovery,

509 herbivore assimilation, soil retention, and leaching losses. Oecologia 151(1): 42-53.

510 doi:10.1007/s00442-006-0579-9.

511 Gessel, S.P., R.E. Miller, and D.W. Cole. 1990. Relative importance of water and nutrient on the

512 growth of coast Douglas fir in the Pacific Northwest. Forest Ecology and Management

513 30, 327-340.

514 Griffin, J.M., and Turner, M.G. 2012. Changes to the N cycle following bark beetle outbreaks in

515 two contrasting conifer forest types. Oecologia 170(2): 551-565. doi:10.1007/s00442-

516 012-2323-y.

517 Hart, S.C., Stark, J.M., Davidson, E.A., Draftand Firestone, M.K. 1994. Nitrogen mineralization,

518 immobilization, and nitrification. pp.985-1019. In A. L. Page et al., eds., Methods of Soil

519 Analysis: Part 2, Microbiological and biochemical properties. Agronomy, A Series of

520 Monographs, no.5, Soil Science Society of America, Madison, Wisconsin USA.

521 Hart, S.C., Firestone, M.K. 1989. Evaluation of three in situ soil nitrogen availability assays.

522 Canadian Journal of Forest Research 19(2): 185-191. doi.org/10.1139/x89-026.

523 Hollinger, D.Y. 1986. Herbivory and the cycling of nitrogen and phosphorus in isolated

524 California oak trees. Oecologia 70(2): 291-297. doi:10.1007/BF00379254.

525 Houle, D., Duchesne, L., and Boutin, R. 2009. Effects of spruce budworm outbreak on elemental

526 export below the rooting zone: a case study for a balsam fir forest. Ann. For. Sci. 66(7):

527 707. doi.org/10.1051/forest/20090.

25 https://mc06.manuscriptcentral.com/cjfr-pubs Canadian Journal of Forest Research Page 26 of 45

528 Huber, C., 2005. Long lasting nitrate leaching after bark beetle attack in the highlands of the

529 Bavarian Forest National Park. J. Environ. Qual. 34(5): 1772-1779.

530 doi:10.2134/jeq2004.0210.

531 Hunter, M.D. 2001. Insect population dynamics meets ecosystem ecology: effects of herbivory

532 on soil nutrient dynamics. Agricultural and Forest Entomology 3(2): 77-84.

533 doi:10.1046/j.1461-9563.2001.00100.x.

534 Hunter, M.D., Linnen, C.R., and Reynolds, B.C. 2003. Effects of endemic densities of canopy

535 herbivores on nutrient dynamics along a gradient in elevation in the southern

536 Appalachians. Pedobiologia 47(3), 231-244. doi:10.1078/0031-4056-00187.

537 IM-Arnold, A., Grüning, M., Simon, J., Reinhardt, A.B., Lamersdorf, N., and Thies, C. (2016).

538 Forest defoliator pests alter carbonDraft and nitrogen cycles. Royal Society Open Science

539 3(10). doi:10.1098/rsos.160361.

540 Keeney, D.R., and Nelson, D.W. 1987. Nitrogen--Inorganic Forms, sec. 33-3, extraction of

541 exchangeable ammonium, nitrate, and nitrite. pp.648-9. In A. L. Page et al., eds.,

542 Methods of Soil Analysis: Part 2, Chemical and Microbiological Properties. Agronomy,

543 A Series of Monographs, no.9 pt.2, Soil Science Society of America, Madison,

544 Wisconsin USA.

545 Kemp, W.P, and Moody, U.L. 1984. Relationships between regional soils and foliage

546 characteristics and Western Spruce Budworm (Lepidoptera: ) oubreak

547 frequency. Environ. Ent. 13(5): 1291-1297. doi.org/10.1093/ee/13.5.1291.

548 le Mellec, A., and Michalzik, B. 2008. Impact of a pine lappet (Dendrolimus pini) mass outbreak

549 on C and N fluxes to the forest floor and soil microbial properties in a Scots pine forest in

550 Germany. Can. J. of For. Res. 38(7): 1829-1841. doi:10.1139/X08-045.

26 https://mc06.manuscriptcentral.com/cjfr-pubs Page 27 of 45 Canadian Journal of Forest Research

551 le Mellec, A., Meesenburg, H., and Michalzik, B. 2010. The importance of canopy-derived

552 dissolved and particulate organic matter (DOM and POM)–comparing throughfall

553 solution from broadleaved and coniferous forests. Ann. For. Sci. 67(4): 411.

554 doi:10.1051/forest/2009130.

555 Lovett, G.M., Christenson, L.M., Groffman, P.M., Jones, C.G., Hart, J.E., and Mitchell, M.J.

556 2002. Insect defoliation and nitrogen cycling in forests: laboratory, plot, and watershed

557 studies indicate that most of the nitrogen released from forest foliage as a result of

558 defoliation by insects is redistributed within the ecosystem, whereas only a small fraction

559 of nitrogen is lost by leaching. BioScience 52(4): 335-341. doi:10.1641/0006-

560 3568(2002)052[0335:IDANCI]2.0.CO;2.

561 Michalzik, B. 2011. Insects, infestations,Draft and nutrient fluxes. In Forest hydrology and

562 biogeochemistry. Edited by D.F. Levia, D. Carlyle-Moses, and T. Tanaka. Springer,

563 Dordrecht. pp. 557-580.

564 Michalzik, B., and Stadler, B. 2005. Importance of canopy herbivores to dissolved and

565 particulate organic matter fluxes to the forest floor. Geoderma 127(3-4): 227-236.

566 doi:10.1016/j.geoderma.2004.12.006.

567 Michalzik, B., Kalbitz, K., Park, J.H., Solinger, S., and Matzner, E. 2001. Fluxes and

568 concentrations of dissolved organic carbon and nitrogen–a synthesis for temperate

569 forests. Biogeochemistry 52(2): 173-205. doi:10.1023/A:1006441620810.

570 Mikkelson, K.M., Bearup, L.A., Maxwell, R.M., Stednick, J.D., McCray, J.E., and Sharp, J.O.

571 2013. Bark beetle infestation impacts on nutrient cycling, water quality and

572 interdependent hydrological effects. Biogeochemistry 115 (1-3): 1-21.

573 doi:10.1007/s10533-013-9875-8.

27 https://mc06.manuscriptcentral.com/cjfr-pubs Canadian Journal of Forest Research Page 28 of 45

574 Mote, P.W., and Salathe, E.P. 2010. Future climate in the Pacific Northwest. Climatic Change

575 102(1-2): 29-50. doi:10.1007/s10584-010-9848-z.

576 Mote, P.W., Li, S., Lettenmaier, D.P., Xiao, M., and Engel, R. 2018. Dramatic declines in

577 snowpack in the western US. Climate and Atmospheric Science 1(1). doi:

578 10.1038/s41612-018-0012-1.

579 Murphy, J., and Riley, J. 1962. A modified single solution method for the determination of

580 phosphate in natural waters. Anal. Chim. Acta 27: 31-36. doi:10.1016/S0003-

581 2670(00)88444-5.

582 Naiman, R.J., R.E. Bilby, D.E. Schindler, and J.M. Helfield. 2002. Pacific salmon, nutrients, and

583 the dynamics of freshwater and riparian ecosystems. Ecosystems 5:399-417.

584 National Atmospheric Deposition ProgramDraft (NADP). 2018. Available from

585 https://nadp.slh.wisc.edu/NTN/maps.aspx. [accessed 1 September 2018].

586 National Oceanic and Atmospheric Administration (NOAA). 2018. Available from

587 https://www.ncdc.noaa.gov. [accessed 7 September 2018].

588 Nealis, V.G. 2012. The phenological window for western spruce budworm: seasonal decline in

589 resource quality. Agricultural and Forest Entomology 14: 340-347. doi:10.1111/j.1461-

590 9563.2012.00574.x.

591 Omernick, J.M. 1987. Ecoregions of the conterminous United States. Annals of the Association

592 of American Geographers 77(1): 118-125.

593 Park, J.H., Kalbitz, K., and Matzner, E. 2002. Resource control on the production of dissolved

594 organic carbon and nitrogen in a deciduous forest floor. Soil Biol. Biochem. 34(6): 813-

595 822. doi:10.1016/S0038-0717(02)00011-1.

596 Piene, H. Effects of insect defoliation on growth and foliar nutrients of young balsam fir. Forest

28 https://mc06.manuscriptcentral.com/cjfr-pubs Page 29 of 45 Canadian Journal of Forest Research

597 Sci. 26(4): 665-673. doi.org/10.1093/forestscience/26.4.665.

598 Piene, H., and Eveleigh, E.S. Spruce budworm defoliation in young balsam fir: the ‘green’tree

599 phenomenon. The Canadian Entomologist 128(6): 1101-1107.

600 doi.org/10.4039/Ent1281101-6.

601 Ponette-González, A.G., Weathers, K.C., and Curran, L.M. 2010. Water inputs across a tropical

602 montane landscape in Veracruz, Mexico: synergistic effects of land cover, rain and fog

603 seasonality and interannual precipitation variability. Global Change Biology 16(3): 946-

604 963. doi:10.1111/j.1365-2486.2009.01985.x.

605 Runyan, C.W., Lawrence, D., Vandecar, K.L., and D'odorico, P. 2013. Experimental evidence

606 for limited leaching of phosphorus from canopy leaves in a tropical dry forest.

607 Ecohydrology 6(5): 806-817. doi:10.1002/eco.1303.Draft

608 Schowalter, T.D. 1999. Throughfall volume and chemistry as affected by precipitation volume,

609 sapling size, and defoliation intensity. Great Basin Nat. 59(1): 79-84.doi:

610 Schowalter, T.D., Sabin, T.E., Stafford, S.G., and Sexton, J.M. 1991. Phytophage effects on

611 primary production, nutrient turnover, and litter decomposition of young Douglas-fir in

612 western . For. Ecol. and Manage. 42(3-4): 229-243. doi:10.1016/0378-

613 1127(91)90027-S.

614 Seastedt, T.R., Crossley Jr, D.A., and Hargrove, W.W. 1983. The effects of low‐level

615 consumption by canopy on the growth and nutrient dynamics of black locust

616 and red maple trees in the southern Appalachians. Ecology 64(5): 1040-1048.

617 doi:10.2307/1937812.

618 Solorzano, L. 1969. Determination of ammonia in natural waters by the phenol hypochlorite

619 method. Limnol. Oceanogr. 14: 799-801.

29 https://mc06.manuscriptcentral.com/cjfr-pubs Canadian Journal of Forest Research Page 30 of 45

620 Stadler, B., Solinger, S., and Michalzik, B. 2001. Insect herbivores and the nutrient flow from the

621 canopy to the soil in coniferous and deciduous forests. Oecologia 126(1): 104-113.

622 doi:10.1007/s004420000514.

623 Swetnam, T.W., and Lynch, A.M. 1989. A tree-ring reconstruction of western spruce budworm

624 history in the southern Rocky Mountains. Forest Science 35(4): 962-986.

625 doi:10.1093/forestscience/35.4.962.

626 Tipping, E., Benham, S., Boyle, J.F., Crow, P., Davies, J., Fisher, U., Guyatt, H., Helliwell, R.,

627 Jackson-Blake, L., Lawlor, A.J., Monteith, D.T., Rowe, E.C., and Toberman, H. 2014.

628 Atmospheric deposition of phosphorus to land and freshwater. Environ. Sci.: Processes

629 Impacts 16(7): 1608-1617. doi:10.1039/C3EM00641G.

630 USFS, 1999. Aerial survey standards. UnitedDraft States Forest Service, Forest Health Monitoring

631 Program.Vitousek, P.M., Porder, S., Houlton, B.Z., and Chadwick, O.A. 2010. Terrestrial

632 phosphorus limitation: mechanisms, implications, and nitrogen–phosphorus interactions.

633 Ecological Applications 20(1): 5-15. doi.org/10.1890/08-0127.1.

634 Vitousek, P.M., Porder, S., Houlton, B.Z., and Chadwick, O.A. 2010. Terrestrial phosphorus

635 limitation: mechanisms, implications, and nitrogen–phosphorus interactions. Ecological

636 Applications 20(1): 5-15. doi.org/10.1890/08-0127.1.

637 Vitousek, P.M., D.N.L. Menge, S.C. Reed, and C.C. Cleveland. 2013. Biological nitrogen

638 fixation: rates, patterns, and ecological controls in terrestrial ecosystems. Philosphical

639 Transactions of the Royal Society B 368: 20130119. doi.org/10.1098/rstb.2013.0119

640 WADNR. 2016. Forest health highlights in Washington – 2015. Washington State Department of

641 Natural Resources, Forest Health Program.

642 Zhao, K., MacLean, D.A., Hennigar, C.R. 2014. Spatial variability of spruce budworm defoliation

30 https://mc06.manuscriptcentral.com/cjfr-pubs Page 31 of 45 Canadian Journal of Forest Research

643 at different scales. For. Ecol. and Manage. 328: 10-19. doi:10.1016/j.foreco.2014.05.020.

644 Zuur, A.F., Ieno, E.N., Walker, N.J., Saveliev, A.A., and Smith, G.M. 2009. Mixed effects models

645 and extensions in ecology with R. Springer, New York.

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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