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Bark Beetle Responses to Stand Structure and Prescribed Fire at Blacks Mountain Experimental Forest, California, USA: 5-Year Data

Abstract

Highly effective fire suppression and selective harvesting of large-diameter, fire-tolerant tree species, such as ponderosa pine (Pinus ponderosa C. Lawson) and Jeffrey pine (P. jeffreyi Balf.), have resulted in substantial changes to the structure and composition of interior ponderosa pine forests. Mechanical thinning and the application of prescribed fire are now commonly used to reduce fuel loads and restore late-seral conditions in interior ponderosa pine forests, but the propensity for some bark beetles (Coleoptera: Curculionidae, Scolytinae) to attack fire-injured trees has led to questions regarding how management objectives may be impacted by levels of delayed tree mortality attributed to bark beetle attack. In this study, we examined bark beetle responses to creation of mid-seral (low structural diversity; LoD) and late-seral (high structural diversity; HiD) forest structures at Blacks Mountain Experimental Forest, California, USA, during a five-year period following the application of prescribed fire (B). Twelve experimental plots (LoD and HiD), ranging in size from 77 ha to 144 ha, were split with one half randomly assigned prescribed fire, resulting in four treatments (HiD, HiD + B, LoD, and LoD + B). A total of 6877 pine and fir (Abies) trees (4.0 % of all trees) were killed by bark beetles (all bark beetle species combined). Most of the tree mortality was attributed to western pine beetle (Dendroctonus brevicomis LeConte) and mountain pine beetle (D. ponderosae Hopkins) in ponderosa pine, and fir engraver (Scolytus ventralis LeConte) in white fir (Abies concolor [Gord. & Glend.] Lindl. ex Hildebr.). The majority of bark beetle-caused tree mortality was recorded on HiD + B (41.5 % of trees that were killed by bark beetles) with 30.0 %, 23.5 %, and 5.0 % occurring on HiD, LoD + B, and LoD, respectively. Overall, a significantly higher percentage of bark beetle-caused tree mortality (of available trees) occurred on LoD + B (5.3 %) and HiD + B (4.8 %) compared to LoD (1.1 %). We observed no significant differences in levels of bark beetle-caused tree mortality (all bark beetle species combined) among diameter classes, but the level of western pine beetle-caused tree mortality was concentrated in the large tree component on burned split plots, which differs from results obtained two years after the application of prescribed fire. The amount of bark beetle-caused tree mortality increased substantially on unburned split plots compared to burned split plots following our initial survey (i.e., three years to five years after the application of prescribed fire). The implications of these and other results for the management and maintenance of late-seral conditions in interior ponderosa pine forests are discussed.

Introduction

Prior to Euro-American settlement, the interior ponderosa pine (Pinus ponderosa C. Lawson var. ponderosa) forest type in California, USA, was dominated by open and parklike stands of widely dispersed ponderosa and Jeffrey pines (P. jeffreyi Balf.), particularly on xeric sites. Frequent thinning of small-diameter (<19 cm diameter at breast height [dbh]) and fire-intolerant tree species by low intensity surface fires, and competitive exclusion of tree seedlings by understory grasses are believed to have maintained such conditions (Oliver 2000). Today, many of these forests are denser, have more small trees and fewer large trees, and are dominated by more shade tolerant and fire intolerant tree species, such as white fir (Abies concolor [Gord. & Glend.] Lindl. ex Hildebr.). These conditions are primarily a result of fire exclusion and past harvesting practices, and have led to heavy accumulations of surface and ladder fuels (Youngblood et al. 2004).

Large and severe wildfires in the western USA, particularly in California, have aroused public concern and manifested the need for well designed silvicultural treatments to reduce their extent and severity. Creating more fire resilient stands generally requires treatment of surface and ladder fuels, reductions in crown density, and maintenance of large diameter trees (Agee and Skinner 2005). A combination of thinning and prescribed fire has been shown to be highly effective for reducing the severity of wildfires in interior ponderosa pine forests. For example, Ritchie et al. (2007) studied the effects of fuel reduction treatments (including those detailed in this study) at Blacks Mountain Experimental Forest (BMEF), California, USA, on fire severity at the point of transition between treated and untreated stands that were later impacted by a wildfire. Tree survival was highest in areas that were both thinned and prescribed-burned. Survival in thinned-only areas was significantly greater than in untreated areas, but less than in areas that had received the combined thinned and prescribed-burned treatment (Ritchie et al. 2007).

Mechanical fuel treatments and the application of prescribed fire may also influence the amount and distribution of bark beetle- (Coleoptera: Curculionidae, Scolytinae) caused tree mortality at various spatial and temporal scales. For example, treatments may affect the health and vigor of residual trees; the size, distribution, and abundance of preferred hosts; and the physical environment within forest stands (Fettig et al. 2007). Carelessly implemented treatments may result in physical damage to residual trees, soil compaction, and increased rates of windthrow, increasing the likelihood of tree colonization by bark beetles, other subcortical insects, and root pathogens (Schweigkofler et al. 2005). Furthermore, tree volatiles released during harvest operations, including monoterpenes, are known to influence the physiology and behavior of bark beetles (Seybold et al. 2006) and colonization rates of trees by bark beetles in ponderosa pine forests (Fettig et al. 2006).

Forest health professionals have recognized that stand susceptibility to bark beetle attack can be manipulated by addressing factors related to tree density (Fettig et al. 2007), typically through mechanical thinning. For example, Sartwell (1971) presented data on radial growth and its relationship to mountain pine beetle (MPB; Dendroctonus ponderosae Hopkins) attack and suggested that slow growth was characteristic of nearly all pines killed by MPB in the Pacific northwest, USA. The author concluded that thinning to reduce tree competition and increase individual tree growth was critical to reducing stand susceptibility to bark beetle attack. Oliver (1995) reported maximum stand density index (SDI) for even-aged ponderosa pine stands in northern California was regulated by MPB and western pine beetle (WPB; Dendroctonus brevicomis LeConte) infestations. An SDI value of 230 defined a threshold for a zone of bark beetle-caused tree mortality within which endemic populations kill a few trees but net growth is positive. Maximum SDI was defined at 365. More recently, Hayes et al. (2009) suggested lower SDI thresholds based on models constructed from 44 stands throughout California. Fettig et al. (2007) discussed tree and stand factors associated with bark beetle infestations in ponderosa pine forests, and provided mechanistic explanations on the effectiveness of reducing stand density for preventing bark beetle infestations.

Bark beetles may attack and kill trees that were injured by fire but might have otherwise survived (Parker et al. 2006, Jenkins et al. 2008). These trees may then serve as a source of beetles and attractive semiochemicals (e.g., host volatiles and aggregation pheromones produced by many bark beetle species during host colonization) that attract other beetles into the vicinity, resulting in additional levels of tree mortality. The propensity for many species of bark beetles to attack fire-injured trees has led to questions regarding how the amount and distribution of bark beetle-caused tree mortality may be affected by the application of fuel reduction treatments. Maintenance of large trees, an important component of late-seral stands, is of primary concern. Kolb et al. (2007) suggested that growth of large ponderosa pines can be accelerated by thinning, but warned that prescribed fire may predispose trees to increased levels of delayed mortality, often caused by bark beetle attack. Recent work at BMEF showed increased vigor in thinned stands, but also an increase in mortality of large trees when thinning was followed by prescribed fire (Ritchie et al. 2008). However, the authors suggested that this is likely a one-time phenomenon occurring shortly after the application of prescribed fire.

In recent years, a substantial body of work has been published documenting the ecological effects of fuel reduction and forest restoration treatments in ponderosa pine forests. Most efforts have quantified short-term (<5 yr) responses obtained from small-scale (<10 ha) experimental plots. In the mid-1990s, an interdisciplinary team of scientists was formed at BMEF to conduct long-term ecological research designed “to increase our understanding of the effect of forest structural complexity on the health and vigor of an interior ponderosa pine ecosystem; to quantify the ecosystem’s responses to natural and human-induced disturbance; and to determine how these ecosystems can be restored and managed” (Oliver 2000). Our work is part of the resulting research project (Blacks Mountain Ecological Research Project) managed by the Pacific Southwest Research Station of the Forest Service and conducted at large spatial scales that represent reasonable management scenarios in these forests. The initial responses of vegetation (Ritchie et al. 2008, Zhang et al. 2008), invertebrates (Camann et al. 2008, Fettig et al. 2008, Gillette et al. 2008), and birds and small mammals (George and Zack 2008, Maguire et al. 2008) to treatments at BMEF have been described. An earlier publication (Fettig et al. 2008) described short-term responses of bark beetles captured during our initial survey conducted two years after the application of prescribed fire at BMEF, in which we reported that prescribed burns resulted in significant increases in the amount of bark beetle-caused tree mortality (all bark beetle species combined) and for WPB, MPB, and fir engraver (FE, Scolytus ventralis LeConte) individually. Our objectives for this study are to document the responses of bark beetles during a five-year period after the application of prescribed fire, and to compare trends in levels of tree mortality attributed to bark beetle attack between the two surveys.

Methods

Study Site and Treatment

We conducted this study at Blacks Mountain Experimental Forest (BMEF; 40°40′ N, 121°10′ W; 1700 m to 2100 m elevation) located on the Lassen National Forest, California, USA. The climate at BMEF is characterized by hot, dry summers and cold, moist winters. Mean daily temperatures range from − 9 °C to 29 °C. The growing season is about 120 days. Annual precipitation averages 50.8 cm with most falling as snow between October and May. The overstory at BMEF is dominated by ponderosa pine with varying amounts of white fir, incense cedar (Calocedrus decurrens [Torr.] Florin) and Jeffrey pine, depending on elevation and aspect. Prior to treatment, stands were dominated by two age cohorts consisting of 300 yr to 500 yr old pines and incense cedars, and a dense understory and mid-story of 50 yr to 100 yr old pines, white firs, and incense cedars (Oliver and Powers 1998, Zhang et al. 2008).

We established twelve experimental plots, ranging in size from 77 ha to 144 ha, to create two distinct forest structural types: mid-seral stage (low structural diversity; LoD) and late-seral stage (high structural diversity; HiD). We randomly assigned structures to four experimental plots (two replicates per structure per block) within each of three blocks. Blocking allowed us to allocate variation to differences in tree species composition associated with elevation gradients and year of treatment. We created LoD by removing larger overstory trees and small understory trees, leaving only trees of intermediate size. We created HiD by thinning smaller trees and retaining larger trees (i.e., thinning from below). Following harvesting, we treated half of each plot with prescribed fire in fall (Table 1). Fettig et al. (2008) describe in detail the climatic conditions, fuel moistures, and other parameters that occurred during prescribed burns (Table 2). The BMEF is included in grazing allotments; therefore, we fenced 6 of 12 experimental plots in an attempt to exclude grazing and permit analyses of the effect of grazing on several variables of interest to the interdisciplinary team. We viewed grazing as having little or no influence on bark beetle activity (Fettig et al. 2008) and was ignored in our analyses. Furthermore, grazing was inconsistent (i.e., concentrated in areas around water) and some cattle gained access to ungrazed plots. Oliver (2000) and Zhang et al. (2008) provide a comprehensive account of treatments. Prior to treatment, the average stand density of proposed experimental units was 32.4 m2 ha−1 ± 1.5 m2 ha−1 (SEM) of basal area. There were 871 trees ha−1 ± 58 trees ha−1 (SEM) composed of 57 % ponderosa and Jeffrey pines and 36 % white fir (Zhang et al. 2008). Immediately following treatment, average values for LoD and HiD were 10 m2 ha−1 and 25 m2 ha−1 of basal area, and 282 trees ha−1 and 513 trees ha−1, respectively. Species composition ranged from 49 % ponderosa and Jeffrey pines and 42 % white fir (HiD + B), to 61 % ponderosa and Jeffrey pines and 29 % white fir (LoD + B) (Zhang et al. 2008).

Table 1 Evaluation dates, Blacks Mountain Experimental Forest, California, USA.
Table 2 Parameters of prescribed burn treatments applied to 12 experimental units at Blacks Mountain Experimental Forest, California, USA.

Data Collection and Analyses

We conducted a 100 % cruise (census) on each experimental plot to locate dead and dying pine and fir trees by presence of crown fade, an irreversible symptom of tree mortality. Because both incense cedar and western juniper (Juniperus occidentalis Hook.) are minor components of BMEF (i.e., representing <7.7 % and <0.1 % of trees, respectively), these species generally are not attacked and killed by bark beetles and we ignored them. Because of the large spatial scale of this study, we staggered our treatment implementation and our initial (1998 to 2002; Fettig et al. 2008) and subsequent (2002 to 2005) surveys across multiple years (Table 1). We identified all recently killed pine and fir trees >19 cm dbh, and tallied and identified the causal agent of mortality. We recorded tree species, dbh, crown color, colonizing bark beetle species, presence of wood borers (Coleoptera: Cerambycidae, Buprestidae), and ranking of burn severity (1 to 4, based on external measures of bole char and bark consumption; Fettig et al. 2008, 2010b). We removed a section of bark approximately 625 cm2 on each recently killed tree with a hatchet at 2 m in height on at least two aspects to determine if any bark beetle galleries were present in the phloem or cambium. The morphology, distribution and orientation of galleries, and the host species are commonly used to distinguish among bark beetle species (Furniss and Carolin 1977). Bark removal also served as a means of separating mortality tallied during the first survey from that of the second survey. Fettig et al. (2008) provide a comprehensive description of methods.

On 26 September 2002, a wildfire (Cone Fire) ignited on the Hat Creek Ranger District, Lassen National Forest, USA, adjacent to BMEF and burned portions of 3 of the 12 experimental plots with variable intensities (Ritchie et al. 2007). One plot (plot 46) suffered low- to moderate-severity wildfire throughout much of the plot, and therefore was excluded from our analyses. Two other plots were partially burned by the Cone Fire. Only the burned split was impacted in plot 41 and the affected portion (i.e., as noted by the presence of bole char, crown scorch, or both on trees) amounted to about 8.7 % (of 55.6 ha). About 27 % (of 51.6 ha) of the burned split and 28.2 % (of 64.5 ha) of the unburned split were affected in plot 43. We subtracted the affected areas from the original area and adjusted the composition data accordingly. Cruises for both plots 41 and 43 were originally scheduled for 2002 (i.e., the summer of the wildfire), and were conducted as planned (Table 1).

Our primary variables of interest were the mean percentage of trees killed by 1) all causes, 2) all bark beetle species across all tree species, 3) individual bark beetle and host tree species, and 4) individual bark beetle species by host tree species within diameter class (19 cm to 29.2 cm, 29.3 cm to 39.3 cm, 39.4 cm to 49.5 cm, 49.6 cm to 59.7 cm, and >59.7 cm dbh). The precise role of each bark beetle species in causing tree mortality on our experimental plots is uncertain as fire-damaged trees in interior ponderosa pine forests are susceptible to colonization by several species of bark beetles. In some cases (e.g., WPB), trees must have green phloem and live buds to permit new needle growth for colonization and brood production to occur (Fischer 1980, Parker et al. 2006), but in others (e.g., Ips spp.) attacks may occur on dead and dying trees (Furniss and Carolin 1977). Accordingly, in reference to WPB, MPB, Jeffrey pine beetle (JPB; D. jeffreyi Hopkins) and FE, we attributed tree mortality to one of these species if evidence of colonization was found despite the potential presence of other bark beetle species. On occasion, we found WPB and MPB infesting the same tree; in these situations, we attributed tree mortality to WPB (Fettig et al. 2008). We attributed tree mortality to Ips spp. only when evidence of WPB, MPB, or JPB colonization was absent. We attributed mortality to WPB, MPB, JPB, FE, and Ips only when burn severity rankings on individual trees were ≤3 (i.e., we estimated that prescribed fire did not directly kill the tree based on external measures of fire severity; Fettig et al. 2008). Red turpentine beetle (RTB; D. valens LeConte) attacks are typically not considered to be a significant threat to tree health, and therefore, we did not attribute tree mortality to RTB attacks.

We analyzed the data cumulatively over the five-year period (i.e., since the application of prescribed fire). We used a randomized complete block experimental design with split plots (Oliver 2000) with three blocks, two treatments (HiD and LoD), and two replicates per treatment (n = 12). Half of each plot was treated with prescribed fire. Due to the imbalance in the number of plots (i.e., due to the Cone Fire) we used the Satterthwaite approximation method to estimate the appropriate degrees of freedom. We performed an analysis of variance on each response variable at α = 0.05 (SAS Institute, Cary, North Carolina, USA). If a significant treatment effect was detected, we used Tukey’s multiple comparison test (Tukey’s HSD) for separation of treatment means requiring more than one comparison.

Results

During this survey, we examined 173 220 pine and fir trees for mortality and bark beetle activity. Of these, 100 433 (58.0 %) were ponderosa pine, 55 396 (32.0 %) were white fir, and 17 391 (10.0 %) were Jeffrey pine. The number of trees within individual dbh classes (prior to adjustments for the Cone Fire) is reported by Fettig et al. (2008).

Overall Tree Mortality

A total of 12 568 (7.3 % of all trees) pine and fir trees died during the five-year period, most of which (41.3 % of dead trees; 5193 trees) occurred on HiD + B with 37.6 %, 18.1 %, and 3.0 % occurring on LoD + B, HiD, and LoD, respectively. These figures represent mortality attributable to all sources (i.e., fire, bark beetle, mechanical, pathogen, and unidentified sources, and their many interactions). Overall, a significantly higher percentage of tree mortality occurred on LoD + B (15.2 % of all trees) than HiD (4.0 %) and LoD (1.5 %) (P = 0.003; Figure 1). In addition, a significantly higher percentage of trees died on HiD + B (9.1 %) than LoD. No other significant differences were observed (Figure 1). Treatment effects varied among dbh classes, but generally the highest levels of mortality were observed in LoD + B (P < 0.020, all cases) except in the larger dbh classes (54.7 cm and >59.7 cm) where few large diameter trees occurred (Figure 1). In the 54.7 cm dbh class, a significantly higher percentage of tree mortality occurred on HiD + B (7.1 %) than HiD (1.2 %) (P = 0.031; Figure 1). In the >59.7 cm dbh class, a significantly higher percentage of tree mortality occurred on HiD + B (8.4 %) compared to HiD (1.2 %) and LoD + B (0 %) (P = 0.024; Figure 1). Approximately 79 % (9916 trees) of all tree mortality occurred on burned split plots, primarily HiD + B.

Figure 1
figure 1

Mean percentage of trees killed by all sources by diameter class (mid-point of 10 cm diameter classes shown except for largest diameter class), and for all trees by treatment (LoD, low structural diversity; HiD, high structural diversity; B, prescribed burn) five years after the application of prescribed fire, Blacks Mountain Experimental Forest, California, USA. Means (+ SE) followed by the same letter within groups are not significantly different (Tukey’s HSD; P > 0.05).

Bark Beetle Responses

During this study, WPB and MPB were found colonizing ponderosa pine; JPB was found colonizing Jeffrey pine; and FE was found colonizing white fir. We also found pine engraver (PE, Ips pini [Say]) and, to a much lesser extent, emarginate ips (I. emarginatus LeConte and I. latidens [=Orthotomicus] LeConte) colonizing ponderosa and Jeffrey pines. A total of 6877 pine and fir trees (4.0 % of all trees) were killed by bark beetles (all bark beetle species combined), the majority of which (41.5 % of trees killed by bark beetles; 2857 trees) occurred on HiD + B with 30.0 %, 23.5 %, and 5.0 % occurring on HiD, LoD + B, and LoD, respectively. Overall, a significantly higher percentage of bark beetle-caused tree mortality occurred on LoD + B (5.3 % of all trees) and HiD + B (4.8 %) than LoD (1.1 %) (P = 0.022; Figure 2). No other significant differences were observed (Figure 2). Similar effects were observed in the two smallest dbh classes (P < 0.021, both cases; Figure 2), and in the 24.1 cm dbh class Hi + B (4.4 %) and HiD (4.1 %) had significantly higher levels of bark beetle-caused tree mortality than LoD (1.1 %) (Figure 2). No significant treatment effect was observed in the 44.5 cm dbh class (P = 0.728). A significantly higher percentage of bark beetle-caused tree mortality occurred on HiD + B compared to LoD + B in the two largest dbh classes (P < 0.041, both cases) (Figure 2). In addition, significantly higher levels of bark beetle-caused tree mortality occurred on HiD + B (5.3 %) than LoD + B (0.9 %) in the >59.7 cm dbh class (Figure 2). Approximately 65 % (4472 trees) of all bark beetle-caused tree mortality occurred on burned split plots, primarily HiD + B.

Figure 2
figure 2

Mean percentage of trees killed by bark beetles (all bark beetle species combined) by diameter class (mid-point of 10 cm diameter classes shown except for largest diameter class), and for all trees by treatment (LoD, low structural diversity; HiD, high structural diversity; B, prescribed burn) five years after the application of prescribed fire, Blacks Mountain Experimental Forest, California, USA. Means (+ SE) followed by the same letter within groups are not significantly different (Tukey’s HSD; P > 0.05).

Western Pine Beetle

Among all plots, 1.2 % of ponderosa pines (1189 trees) were killed by WPB, of which 61.5 % (731 trees) occurred on HiD + B, with 21.7 %, 13.4 %, and 3.4 % occurring on HiD, LoD + B, and LoD, respectively. Overall, a significantly higher percentage of WPB-caused tree mortality occurred on HiD + B (2.8 %) than HiD (0.8 %) and LoD (0.5 %) (P = 0.018; Figure 3). No other significant differences were observed. In the two smallest dbh classes, a significantly higher percentage of WPB-caused tree mortality occurred on HiD + B than LoD (P < 0.031, both cases), and in the 34.3 cm dbh class, LoD + B (2.7 %) had significantly higher levels of WPB-caused tree mortality than LoD (0.7 %) (Figure 3). No significant treatment effect was observed in the 44.5 cm dbh class (P = 0.232). A significantly higher percentage of WPB-caused tree mortality occurred on HiD + B (4.3 %) than HiD (0.5 %) and LoD + B (0 %) in the 54.7 cm dbh (P = 0.011; Figure 3). In the largest dbh class, significantly higher levels of WPB-caused tree mortality occurred on HiD + B (5.2 %) than any other treatment (P = 0.020; Figure 3). Approximately 75 % (891 trees) of all WPB-caused tree mortality occurred on burned split plots, primarily HiD + B. Most of this activity was concentrated in the largest dbh class (P = 0.041; Figure 4).

Figure 3
figure 3

Mean percentage of ponderosa pines killed by western pine beetle (WPB) by diameter class (mid-point of 10 cm diameter classes shown except for largest diameter class), and for all trees by treatment (LoD, low structural diversity; HiD, high structural diversity; B, prescribed burn) five years after the application of prescribed fire, Blacks Mountain Experimental Forest, California, USA. Means (+ SE) followed by the same letter within groups are not significantly different (Tukey’s HSD; P > 0.05). Asterisk denotes no ponderosa pines existed in this dbh class within this treatment.

Figure 4
figure 4

Colonization rates of available pines among diameter classes (mid-point of 10 cm diameter classes shown except for largest diameter class) on burned split plots for western pine beetle (WPB), mountain pine beetle (MPB), Ips spp. (Ips), and all bark beetle species combined, Blacks Mountain Experimental Forest, California, USA. Means (+ SE) followed by the same letter within groups are not significantly different (Tukey’s HSD; P > 0.05).

Mountain Pine Beetle

Among all plots, 1.4 % of ponderosa pines (1415 trees) were killed by MPB, of which 40.7 % (576 trees) occurred on HiD + B with 32.7 %, 21.8 %, and 4.8 % occurring on HiD, LoD + B, and LoD, respectively. Overall, a significantly higher percentage of MPB-caused tree mortality occurred on HiD + B (2.1 %) and LoD + B (1.8 %) than LoD (0.5 %) (P = 0.046; Figure 5). No other significant differences were observed (Figure 5). No significant treatment effects were observed in the 24.1 cm, 44.5 cm, and 54.7 cm dbh classes (P > 0.101). In the 34.3 cm dbh class, a significantly higher percentage of MPB-caused tree mortality occurred on HiD + B (2.8 %) and LoD + B (2.7 %) than LoD (0.6 %) (P = 0.018; Figure 5). In the largest dbh class, a significantly higher percentage of MPB-caused tree mortality occurred on HiD + B (0.4 %) than any other treatment (P < 0.001; Figure 5). Approximately 62 % (884 trees) of all MPB-caused tree mortality occurred on burned split plots, primarily HiD + B. Most of this activity was concentrated in the smaller dbh classes (P < 0.001; Figure 4).

Figure 5
figure 5

Mean percentage of ponderosa pines killed by mountain pine beetle (MPB) by diameter class (mid-point of 10 cm diameter classes shown except for largest diameter class), and for all trees by treatment (LoD, low structural diversity; HiD, high structural diversity; B, prescribed burn) five years after the application of prescribed fire, Blacks Mountain Experimental Forest, California, USA. Means (+ SE) followed by the same letter within groups are not significantly different (Tukey’s HSD; P > 0.05). Asterisk denotes no ponderosa pines existed in this dbh class within this treatment.

Jeffrey Pine Beetle

A total of 35 Jeffrey pines (0.2 % of all Jeffrey pines) were killed by JPB and therefore statistical analyses were not conducted. Most of this mortality (85 %) occurred on burned split plots.

Ips spp.

Among all plots, 0.4 % of all pines (471 trees) were killed by Ips spp., of which 65.4 % (308 trees) occurred on LoD + B with 33.3 % and 1.1 % occurring on HiD + B and LoD, respectively. Only one tree was killed by Ips on HiD. Overall, a significantly higher percentage of Ips-caused tree mortality occurred on LoD + B (1.3 %) and HiD + B (0.5 %) than LoD (<0.1 %) and HiD (<0.1 %) (P = 0.038; Figure 6). A similar effect was observed in the smallest dbh class (P = 0.015; Figure 6). In the 34.3 cm dbh class, a significantly higher percentage of Ips-caused tree mortality occurred on LoD + B (1.4 %) than LoD (0.9 %) and HiD (0 %) (P = 0.002; Figure 6). No significant treatment effects were observed in the three largest dbh classes (P > 0.247, all cases). Approximately 99 % (465 trees) of all Ips-caused tree mortality occurred on burned split plots, primarily LoD + B. Most of this activity was concentrated in the smaller dbh classes (P < 0.001; Figure 4).

Figure 6
figure 6

Mean percentage of pines killed by Ips spp. (Ips) by diameter class (mid-point of 10 cm diameter classes shown except for largest diameter class), and for all trees by treatment (LoD, low structural diversity; HiD, high structural diversity; B, prescribed burn) five years after the application of prescribed fire, Blacks Mountain Experimental Forest, California, USA. Means (+ SE) followed by the same letter within groups are not significantly different (Tukey’s HSD; P > 0.05). Asterisk denotes no pines existed in this dbh class within this treatment.

Fir Engraver

A total of 3767 white firs (6.8 % of all white firs) were killed by FE, of which 36.4 % (1372 trees), 35.5 %, 22.1 %, and 6.0 % occurred on HiD + B, HiD, LoD + B, and LoD, respectively. A significantly higher percentage of FE-caused tree mortality occurred on LoD + B (10.1 %) than LoD (1.6 %) (P = 0.005; Figure 7). No other significant differences were observed among treatments. In the smallest dbh class, a significantly higher percentage of FE-caused tree mortality occurred on LoD + B (9.4 %) and HiD (7.1 %) than LoD (1.4 %) (P = 0.008; Figure 7). No significant treatment effects were observed in the 34.3 cm, 44.5 cm, and largest dbh classes (Figure 7). A significantly higher percentage of FE-caused tree mortality occurred on HiD + B (10 %) than HiD (4.2 %) in the 54.7 cm dbh class (P = 0.002; Figure 7). Approximately 58.5 % (2203 trees) of all FE-caused tree mortality occurred on burned split plots, primarily Hi + D.

Figure 7
figure 7

Mean percentage of white firs killed by fir engraver (FE) by diameter class (mid-point of 10 cm diameter classes shown except for largest diameter class), and for all trees by treatment (LoD, low structural diversity; HiD, high structural diversity; B, prescribed burn) five years after the application of prescribed fire, Blacks Mountain Experimental Forest, California, USA. Means (+ SE) followed by the same letter within groups are not significantly different (Tukey’s HSD; P > 0.05). Asterisks denote no firs existed in this dbh class within these treatments.

Discussion

One of the primary objectives at BMEF was to create stand conditions that serve as a model of resiliency to wildland fire and insects for interior ponderosa pine forests. The effectiveness of fuel reduction treatments to increase the resiliency of interior ponderosa pine forests to fire is well established (e.g., Agee and Skinner 2005, Ritchie et al. 2007), but may have unintended effects. For example, a common concern is that fire-injured trees may serve as breeding substrates for bark beetles that then attack adjacent trees or adjacent untreated areas at elevated levels. However, this has not been well documented in the scientific literature (Parker et al. 2006), perhaps due to the limited spatial and temporal scales of many studies. To that end, large numbers of severely stressed trees could provide abundant host material and once this resource has been exhausted (i.e., within one to two years following prescribed burns), bark beetles may attack and kill trees that might otherwise have survived. Breece et al. (2008) reported that prescribed burns increased bark beetle attacks on ponderosa pines over a three year period. Eighty percent of all bark beetle-attacked trees were colonized during the first year following the application of prescribed fire. During the second and third years, 8 % and 12 % were colonized, respectively (Breece et al. 2008). Fettig et al. (2010a) reported that the amount of bark beetle-caused tree mortality was evenly distributed two and four years after the application of thinning and burning treatments in the southern Cascades, USA. In the central Sierra Nevada, USA, it was reported that 38 %, 42 %, and 20 % of bark beetle-caused tree mortality occurred during the first, second, and third years following prescribed fire, respectively (Fettig et al. 2010b). The authors commented that all of the delayed mortality attributed to bark beetle attack may not have been captured during the three-year period. We found that 37 % and 63 % of bark beetle-caused tree mortality occurred during our initial (one to two years after prescribed fire) and second surveys (three to five years), respectively, an effect similar (when adjusted for temporal scale) to that observed in a mixed conifer forest in California (Fettig et al. 2010a). Unexpectedly, a large percentage of the total bark beetle-caused tree mortality that was recorded on HiD and LoD (84 % and 88 %, respectively) occurred three to five years after the application of prescribed fire to adjacent split plots. However, an unexpectedly low percentage of the total bark beetle-caused tree mortality recorded on LoD + B (38 %) occurred three to five years after the application of prescribed fire. The effect of the Cone Fire on levels of bark beetle-caused tree mortality detected during our surveys conducted in 2004 and 2005 (i.e., 2 yr to 3 yr after the event; Table 1) cannot be determined. While fire-injured trees may serve as breeding substrates for bark beetles that then may attack trees in adjacent areas at elevated levels in subsequent years (Parker et al. 2006), much of the area affected by the Cone Fire was a stand replacing event. Furthermore, of the nine plots that were not directly affected by the Cone Fire, all except one (plot 45) are >1 km from affected areas and several are separated by a large meadow (Oliver 2000). We therefore feel that the Cone Fire likely had little effect on levels of bark beetle-caused tree mortality observed on our plots.

Levels of bark beetle-caused tree mortality increased substantially on unburned split plots compared to burned split plots following our initial survey. For example, in the HiD structure, we observed a 420 % increase in the number of trees killed by bark beetles (333 versus 1731) between surveys on unburned split plots compared to a 50 % increase (1139 versus 1718) on burned split plots. Similarly, in the LoD structure, we observed a >600 % increase (41 versus 301) on unburned split plots compared to a 38 % decrease (998 versus 616) on burned split plots. Trends were similar for all individual bark beetle species (i.e., WPB, MPB, and FE) except Ips spp. where 96 % of all Ips-caused tree mortality was detected during our initial survey (Fettig et al. 2008). Ips spp. generally colonize slash, saplings, and weakened trees. Outbreaks are often short-lived, but increase in duration and extent when suitable host material is plentiful and populations grow sufficiently to colonize apparently healthy trees (Furniss and Carolin 1977). While concerns regarding Ips-population increases in logging slash and subsequent colonization of trees are important, prescriptions at BMEF included removal and utilization of slash following thinning. Accordingly, little or no slash was left within experimental plots to serve as breeding substrates for Ips beetles (Kegley et al. 1997, DeGomez et al. 2008). The response of Ips spp. to fuel reduction treatments at BMEF appears relatively short-lived and limited to fire-injured trees in the two smallest dbh classes (94 % of all cases) (Fettig et al. 2008). However, the apparent shift in WPB, MPB, and FE activity from burned to unburned split plots three to five years after the application of prescribed fire is concerning, and warrants further monitoring of these plots. The effect is likely due to unburned areas not benefiting from the positive effects of prescribed fire (e.g., on growing space three to five years after treatment) that affect tree vigor and thus susceptibility to bark beetle attack (Fettig et al. 2007). Unburned areas (split plots) were adjacent to burned areas (split plots) that likely served as an important source of beetles once the abundance of fire-injured trees that were highly susceptible to bark beetle attack declined. Interestingly, Fettig et al. (2006) observed a similar effect (albeit on a much smaller spatial scale) in another study examining the effects of chipping of unmerchantable trees on bark beetle activity in ponderosa pine stands. In that study, chipping following thinning increased plot risk to bark beetle attack in the short-term by the production of large amounts of attractive monoterpenes. In the longer term, it decreased hazard by increasing the amount of growing space allocated to each residual tree. Surveys along the perimeter of chipped plots revealed large numbers of recently attacked trees in untreated areas that did not benefit from the positive effects of thinning, but suffered a similar level of risk associated with high levels of monoterpenes beneath the forest canopy.

Concerns about maintaining large-diameter pines, particularly in reference to delayed mortality attributed to bark beetle attack on HiD + B, were expressed among the interdisciplinary team. Larger trees are more fire-resistant (Peterson and Ryan 1986, Harrington 1993, Thies et al. 2005). Furthermore, tree species with thicker bark (e.g., ponderosa pine) are less likely to suffer fire-induced damage than thinner barked species (e.g., white fir) (Ryan and Reinhardt 1988, Ryan and Frandsen 1991). In the central Sierra Nevada, Regelbrugge and Conard (1993) reported decreased probability of post-fire mortality in ponderosa pine with increasing tree diameter and height. We observed no significant differences in the proportion of pines killed by bark beetles (all bark beetle species combined) among dbh classes following prescribed fire, but for WPB, MPB, and Ips spp. individually, an effect was observed (Figure 4). For MPB and Ips spp., mortality was aggregated in the smallest dbh classes, which agrees with their host preference for smaller diameter trees in interior ponderosa pine forests (Furniss and Carolin 1977). However, for WPB, significantly higher levels of tree mortality (5.3 %) were observed for the largest dbh class, which differs from data from our initial survey (Fettig et al. 2008). The MPB frequently colonizes smaller diameter ponderosa pines in California, while WPB attacks are concentrated in the larger diameter classes (Miller and Keen 1960), which likely explains the different effects observed for these species. During our initial survey, 71 large dbh (>59.7 cm) ponderosa pines were killed by WPB on HiD + B and 9 on HiD. At the end of five years, 220 large dbh ponderosa pines had been killed by WPB on HiD + B and 53 on HiD. If this level continues, it will likely interfere with management objectives at BMEF that are aimed at maintaining large trees for fire resiliency and late-seral forest conditions. Only one tree was killed by WPB on LoD plots in the two largest dbh classes, which is surprising given the beetle’s preference for larger diameter trees (50.8 cm dbh to 81.3 cm dbh; Person 1928, Miller and Keen 1960), and may be an artifact of only 359 and 30 trees occurring within the 54.7 and >59.7 cm dbh classes on LoD plots, respectively.

We caution the reader that bark beetle responses to silvicultural treatments, specifically prescribed burns, are heavily influenced by location as this dictates the bark beetle assemblages that are present (e.g., WPB does not occur throughout the entire range of ponderosa pine) and associated population levels during and after fire. In brief, interior ponderosa pine forests are recognized for the diversity of tree-killing bark beetle species inhabiting them, most notably WPB, MPB, JPB, PE, and FE. In Fettig et al. (2008), we emphasized that results pertained to short-term responses under certain abiotic and bark beetle population pressures. At the beginning of this study, tree mortality attributed to WPB, MPB, JPB, and FE in northeastern California declined to what was considered “background levels” following an earlier period of elevated activity. In 2001, activity of these species, with exception of JPB, began to slightly increase in portions of northeastern California (Fettig et al. 2008) and remained at somewhat elevated levels by the time our surveys were completed in 2005 (Forest Service 2010). One of the best recognized inciting factors in tree and forest health decline is deficiency in moisture availability, which affects host tree vigor and thus defensive mechanisms, and the fitness and survivorship of insect herbivores such as bark beetles (Fettig et al. 2007). Much of northern California emerged from four years of drought with above normal precipitation in 2005. To that end, drought may have been an important factor influencing bark beetle activity during our second surveys, particularly for FE, which readily responds to drought stress in white fir (Berryman and Ferrell 1988). Interestingly, we observed a 123 % increase in FE-caused tree mortality three to five years following the application of prescribed fire (i.e., the period of time generally coinciding with drought) compared to a 32 % increase in bark beetle-caused tree mortality in the pine component during the same period of time. Accordingly, we feel that the responses documented here occurred during bark beetle population pressures commonly observed in these forests. However, a more significant test of the resiliency of these treatments and resulting structures will occur if bark beetle populations build in the region.

Slightly more than half (6877 trees) of tree deaths were attributed to delayed mortality resulting from bark beetle attack. The relatively low amount (4.0 %) of bark beetle-caused tree mortality is not surprising given the residual stand densities achieved in this study for both LoD and HiD (Oliver 2000, Zhang et al. 2008), and the strong relationship between stand density and tree and stand susceptibility to bark beetle attack. To that end, we observed no difference in the percentage of bark beetle-caused tree mortality between LoD and HiD except in the smallest dbh class (Figure 2), suggesting that the resiliency of these two structures to bark beetle infestations were similar over the five-year period. In Fettig et al. (2008), we concluded that near-term (two years following the application of prescribed fire) levels of bark beetle-caused tree mortality (1.5 %) would not interfere with management objectives, particularly considering that almost half was represented by FE infesting white fir and treatments were designed to promote ponderosa and Jeffrey pines over white fir. At that time, we issued a caveat that the effects of RTB attacks and prescribed fire on individual tree health and on subsequent levels of tree mortality may not be realized for some time. Eighty-one percent of all RTB-attacked trees over the five-year period were recorded during our initial survey, and we saw no evidence of trees that were attacked by RTB dying in the absence of attacks by other bark beetle species or confounding factors (e.g., mechanical damage).

Fettig et al. (2008) reported that wood borer attacks directly contributed to tree mortality in numerous fire-injured trees at BMEF during the second year after the application of prescribed fire, and that in the absence of such attacks tree mortality would likely not have occurred. However, this effect was not observed during our second survey or in a similar study conducted in the southern Cascades (Fettig et al. 2010a). While a body of literature is building concerning evidence of wood borers directly contributing to tree mortality in fire-injured trees (Parker et al. 2006), the effect, when it occurs, seems to be short-lived, occurring immediately after disturbance.

Two years after the application of prescribed fire at BMEF, we (Fettig et al. 2008) concluded that concerns regarding delayed mortality of large diameter ponderosa and Jeffrey pines at BMEF were unfounded. However, we issued a caveat that this relationship could change in the future, particularly if significant root damage had occurred as a result of the application of prescribed fire. The increased rates of mortality attributed to WPB attacks in the large tree component following prescribed fire (Figures 3 and 4) is concerning and may be occurring as a result of greater basal duff accumulations resulting in increased fire intensity and fine root injury, and thus reduced tree vigor. Swezy and Agee (1991) reported lethal heat loads occurred in >75 % of samples at the soil interface and at 5 cm soil depth following the application of prescribed fire in old-growth ponderosa pines at Crater Lake National Park, USA. Similar effects have been observed in other pine systems (Sullivan et al. 2003). We intend to continue monitoring these plots in the future to determine the longer-term effects of these treatments and the resulting stand structures on the amount of bark beetle-caused tree mortality, particularly in the large tree component.

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Acknowledgements

We thank R. Borys, C. Dabney, M. Patterson, and C. Stoll (formerly Pacific Southwest Research Station, PSW) for technical assistance, and J. Baldwin (PSW) for aid with statistical analyses, specifically concerning the Satterthwaite approximation method. We acknowledge the numerous contributions of M. Ritchie, C. Skinner, and others at the Redding Laboratory (PSW) to the maintenance of scientific infrastructure at BMEF. We also thank our numerous colleagues on the Blacks Mountain Experimental Forest Ecological Research Team for their participation in this large-scale study. Special thanks to C. Hayes (PSW) and two anonymous reviewers for their critiques, which greatly improved earlier versions of this manuscript. This research was supported, in part, by ecosystem management research funds received through the Washington Office of the Forest Service, a US President’s Early Career Award for Scientists and Engineers (to C.J.F.), and the Pacific Southwest Research Station. This article was written and prepared by US government employees on official time and it is, therefore, in the public domain and not subject to copyright.

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Fettig, C.J., McKelvey, S.R. Bark Beetle Responses to Stand Structure and Prescribed Fire at Blacks Mountain Experimental Forest, California, USA: 5-Year Data. fire ecol 6, 26–42 (2010). https://doi.org/10.4996/fireecology.0602026

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