Patterns of woodboring beetle activity following fires and bark beetle outbreaks in montane forests of California, USA

AIC: Akaike’s information criterion

BA: basal area

BBO: bark beetle outbreak

BUPR: buprestids

CERA: cerambycids

CWHR: California Wildlife Habitat Relationships

DBH: diameter at breast height

DOY: day of year

EPN: Eastside Pine (forest)

MECO: Melanophila consputa

MONO: Monochamus species

RdNBR: Relative differenced Normalized Burn Ratio

SE: standard error

SMC: Sierran Mixed Conifer (forest)

QAIC: quasi-AIC (see AIC)

VRP: variable radius plot

Forests in the western United States have entered a period of intensifying disturbance. Drought years have become increasingly common and severe, and projected changes in precipitation are expected to exacerbate these trends (Seager et al. 2007; Diffenbaugh et al. 2015). Severe drought and increasing temperatures leave forests drier and more susceptible to bark beetle (Coleoptera: Scolytinae) outbreaks triggered by the lower resistance of drought-stressed trees (Bentz et al. 2010; Williams et al. 2013; Kolb et al. 2016; Marini et al. 2017). The frequency and area of large fires (Littell et al. 2009; Dennison et al. 2014), as well as fire-season length (Westerling et al. 2006; Westerling 2016), have increased in recent decades across the western US, where forest fires are expected to become more common in response to predicted changes in climate, especially in montane forests (Abatzoglou and Williams 2016). Larger and more severe wildfires have been facilitated by increased forest density, fuel loading, and the loss of forest heterogeneity resulting from fire exclusion (Miller et al. 2009a; Taylor et al. 2014; Boisramé et al. 2016). Over the next two decades, as many as 11 states in the western US might see the average annual area burned increase by 500 percent (Kitzberger et al. 2017), and temperature-driven models indicate that there is potential for extensive bark beetle outbreaks in forests of the western US throughout this century (Raffa et al. 2008; Bentz et al. 2010). The increase in tree mortality resulting from these large and frequent disturbances is expected to provide abundant habitat for woodboring beetles that typically colonize trees killed or weakened by other agents (Cranshaw et al. 2000). Many woodboring beetles aid decomposition and nutrient recycling in disturbed forest stands (Jacobsen et al. 2015; Kahl et al. 2017). Understanding how woodborers respond to different types of forest disturbance and stand characteristics will be important for predicting the response of forest communities to changes in the disturbance regime.

Woodboring beetle families Cerambycidae and Buprestidae contain hundreds of species, many of which are common and widespread throughout coniferous forests of the western US. As reviewed in Furniss and Carolin (1977), most woodboring beetles are considered secondary forest pests because they generally colonize trees already killed or weakened by disturbance, including wildfire, blowdown, timber harvest activities, and outbreaks of primary forest pests such as “primary” bark beetles that can kill healthy trees. In trees colonized by bark beetles, the spatial distribution of woodborer and bark beetle larvae is sometimes disjunct, and there is some evidence of competition for phloem between species within trees (Coulson et al. 1976; Foelker et al. 2018), as well as some evidence that woodboring larvae consume bark beetle larvae (Dodds et al. 2001). Larval woodborers feed within the cambium layer during early development (Fig. 1) and later tunnel into sapwood and heartwood. Their tunneling introduces fungi that reduce the economic value of harvested timber but also facilitate the decomposition of dead wood, recycling nutrients and reducing fuel loads (Harrington 1993; Lowell and Cahill 1996). The life cycle of both cerambycid and buprestid beetles is dominated by the larval stage, which typically lasts one year but can extend to several years (Kariyanna et al. 2017). Adults emerge and fly during the warmer months, seeking out recently dead or weakened trees. Most woodborers are unable to attack healthy trees and are more often attracted to trees that release chemical signals of stress (Ikeda et al. 1980; Kelsey and Joseph 2003; Miller 2006). Some woodborers locate burned trees by sensing heat or smoke (Schmitz et al. 1997; Schütz et al. 1999; Álvarez et al. 2015). Females deposit eggs in bark crevices, under bark scales, or in small niches cut into the bark, and have been observed to avoid oviposition where bark beetle activity is high (Gardiner 1957).

Research on woodboring beetles has concentrated primarily on native species that cause damage to wood products (Álvarez et al. 2015) or exotic and invasive species that disrupt entire ecosystems (Aukema et al. 2010; Aukema et al. 2011). Much less is known about how native woodboring beetles colonize and use trees damaged by fire and primary bark beetles, and how woodborers respond to the timing and severity of disturbance, forest composition and structure, host tree attributes, and interspecific competition (Costello et al. 2011; Costello 2013; Brin and Bouget 2018). Given the observed potential for ecological and economic impacts of many woodborer species, characterizing these responses could inform management of disturbed forests. Management of post-disturbance forests often involves multiple objectives, such as recovering the economic value of dead trees (Eklund et al. 2009), mitigating hazards associated with disturbance, facilitating reforestation (Collins and Roller 2013; Bohlman et al. 2016), reducing fuel levels and the risk of type conversion (Coppoletta et al. 2016), and providing a diversity of wildlife habitats (White et al. 2015). Achieving these objectives might be facilitated by manipulating woodborer impacts on woody fuel decomposition, nutrient cycling, and succession, as well as increasing the availability of woodboring larvae as prey for early-seral habitat specialists. For example, the black-backed woodpecker (Picoides arcticus Swainson, 1832) is a species of management interest (Saracco et al. 2011; Siegel et al. 2018) that feeds primarily on the larvae of woodboring beetles (Villard and Beninger 1993; Murphy and Lehnhausen 1998). Relative to bark beetles, woodborer larvae are larger and provide a more concentrated prey resource (Powell 2000).

In these montane forests of California, the range of woodborer response to forest conditions suggests that managing for a diversity of disturbance types and severity will support a diverse woodborer community. The different taxa and life stages that we sampled appeared to respond in characteristic ways. Buprestid adults were much more frequent in stands that burned at high severity, a pattern especially evident in Melanophila consputa, which is known for its attraction to heat and smoke (Schmitz et al. 1997; Schütz et al. 1999). In contrast, cerambycids that sense heat and smoke (Monochamus spp.; Álvarez et al. 2015) were more frequent in transects burned at low and moderate severity. Larval activity responded more often to features that might represent the capacity for larval growth and development. Disturbance type, however, appeared to have a relatively minor influence on the activity of most woodborers in our study. Except where burn severity was high, cerambycids and buprestids of both stages appeared to occur with similar frequency in burned and BBO sites.

The scale and timing of each analysis should be considered when interpreting an appropriate scope of inference. Burned sites were distributed throughout the Sierra Nevada, while BBO sites were clustered in the northern half of this region, and disturbance type also covaried with disturbance age, because large bark beetle outbreaks had just begun to develop as we initiated the study. We also sampled during a period of extended drought in our study region, when primary and secondary forest pests were likely to be responding in complex ways to stress on the forests in this system (Kolb et al. 2016). Regardless of the unique features of this study, strong patterns were revealed in our data that are likely relevant to many dry coniferous forests that occur throughout the western US. Evidence for these relationships should be investigated for generality in other regions, and across a diversity of fire ages and outbreak sites, to facilitate inferences currently constrained by the naturally high covariance and spatial clustering in values of key predictor variables.

The strong patterns we found, however, can help inform where—and at what scale—management actions might be most effectively applied. Within a disturbed patch of forest, the deliberate retention of pockets or islands of snags with specific characteristics might limit the negative ecological impacts of a reduction in woodborer activity. For example, larval woodborers were more common in particular quadrats where bark beetles were absent, suggesting that the local retention of woodborers might be enhanced by prioritizing retention of trees less heavily infested by bark beetles. Our observations that woodborer larvae were more abundant in smaller pine snags with more southerly exposure, that Xylotrechus Chevrolat 1860 adults were more common where the dominant host tree taxon was true fir, and that Monochamus species were more common where burn severity was low to moderate, can further inform design of snag retention patches that might effectively promote successional processes and provide prey for wildlife species like the black-backed woodpecker (Siegel et al. 2018; Tarbill et al. 2018; Stillman et al. 2019).

The design of post-fire forest management strategies often involves balancing multiple economic and ecological objectives with clear links to woodborer dynamics. Taken together, our results suggest a dynamic process controlling local initiation and accumulation of woodborer activity that likely leads to temporal trends in the faunal communities supported by woodborer larvae and adults as prey and agents of tree decomposition. Beginning with the environment inhabited by larvae, our finding of higher larval activity on smaller trees and southern bark aspects suggests that there are microclimatic effects on larval development. Smaller-diameter trees have thinner bark and lower thermal mass of sap and heartwood, as well as lower sapwood volume (water holding capacity), leading to more rapid desiccation of phloem tissue (Six et al. 2002; Hayes et al. 2008). Southern bole aspect would intensify these microclimatic conditions, likely creating a more favorable environment for rapid larval development. An inverse relationship between tree size and larval woodborer activity was also noted by Powell (2000) in the northern Rocky Mountains. Hanks (1999) suggested that phloem-boring larvae might be especially vulnerable to moisture conditions because they are embedded in the tissues of their host plant, and sensitivity to bark moisture could account for the association between drought stress and host susceptibility that is typical for many phloem-boring insects, especially cerambycid beetles. Thatcher (1961) reported that drought conditions can enable a larger number of cerambycid larvae to survive in systems in which adults attempt to attack living trees. Subcortical and sapwood temperatures in trees and logs exposed to sunlight can be 5 to 30 °C greater than shaded portions, and these warmer temperatures can decrease the larval development time of phloem-feeding woodborers by 1 to 2 yr (Graham 1925, Savely 1939, Hosking 1977; reviewed in Kariyanna et al. 2017). Alternatively, these patterns could be related to adult female selection of oviposition locations, although we found no evidence of such selectivity in the literature.

The inverse relationship between bark beetle and larval woodborer activity that we observed was likely due to competition for phloem resources or other factors affecting adult or larval microhabitat partitioning among beetle species. Because woodboring beetles typically colonize dead and dying trees, such as those killed or weakened by bark beetles, the nutrient-rich phloem resource can be greatly reduced by bark beetle activity prior to colonization by woodboring beetles, limiting its availability for larval woodborer development (Coulson et al. 1979; Powell 2000). Thus, variation in the availability of phloem for woodboring beetle larvae could depend on the timing of bark beetle attack versus woodborer colonization. For example, the western pine beetle (Dendroctonus brevicomis Erichson 1836) produces more than one generation per year (Miller and Keen 1960), which might allow these bark beetles to attack trees outside the peak flight period for woodboring beetles. If bark beetles attack trees in the weeks or months before woodborer dispersal, the resource might be depleted before woodborer eggs hatch. If bark beetles attack trees after woodborer dispersal, changes in the resource during the ensuing year might combine with resource depletion to make trees unattractive or unsuitable for some woodboring beetles. Alternatively, tree size can play a role in competition with bark beetles such as the fir engraver (Scolytus ventralis LeConte 1868), which tends to attack higher on the bole of larger-diameter trees, leaving the lower portions of the bole available for woodboring beetles (Ferrell and Hall 1975). In addition, because primary bark beetles avoid fire-killed trees, especially those with foliage completely scorched or consumed, the phloem resource in fire-killed trees might often be more available for woodboring beetle colonization (Parker et al. 2006). In contrast, competition with secondary bark beetles, such as Ips De Geer 1775 spp., in fire-killed trees may occur in the upper boles and limbs of larger trees and all portions of smaller trees. This is especially true for secondary bark beetles attacking thicker-barked tree species such as those sampled in this study. However, Costello (2013) found that no bark beetles emerged from fire-killed tree bolts, even though outbreaks of the mountain pine beetle (Dendroctonus ponderosae Hopkins 1902) were abundant near the fires in that study.

Costello (2013) also found significantly more woodborers emerging from fire-killed trees than from trees killed by bark beetles, in contrast with our finding of similar woodborer activity in burned and BBO stands. This contrast could reflect a difference in the timing of post-fire monitoring between their study of fires 1 to 2 yr old and our study of fires 1 to 8 yr old. Furthermore, bark beetle activity was common in our burned sites, occurring in over 30% of bark samples. Some snags in older burned sites might have been killed by bark beetles attacking fire-stressed trees in the years following a burn (Davis et al. 2012), reducing the distinction between burned and BBO sites as a food resource for woodboring beetles. Potential similarities in post-disturbance ecology of burned and BBO sites warrant further investigation.

Fire age also helped explain the pattern of adult captures. For all but one taxon modeled in burned sites, adult capture numbers declined with fire age, as expected if older fires were less attractive to dispersing adults or less productive as source habitats for these species. Data from buprestids (especially M. consputa) and cerambycids (Monochamus species) also supported a positive interaction between fire age and squared ignition date. Although these species were more common at burned sites with intermediate ignition dates, those fires did not appear to act as long-term sources of these species in the landscape. This dynamic process should lead to temporal trends in the faunal communities woodborers support. Indeed, black-backed woodpecker abundance in burned forests in California generally peaks within five years after fire, followed by sharp declines (Tingley et al. 2018).

Our results also suggest that there are competing influences of disturbance age. Adult woodborers might cue on current signs of a snag-producing disturbance, but signs of larval activity accumulate over time, so we might expect stronger or opposing effects of disturbance age on adult captures when compared with larval activity. In fact, we found a positive effect of fire age on the combined activity of larval woodborer species, but negative effects of fire age on captures of adult M. consputa, X. longitarsus, and Monochamus species, as well as all adult buprestids combined. Negative effects of fire age were most evident for M. consputa, as expected for members of a genus notably adapted for detecting fires in progress. Emerging adults of species that are attracted to fire might often emigrate, seeking fires in progress and reducing local recruitment that would otherwise tend to obscure a negative relationship between fire age and adult captures. We found negative effects of fire age on both buprestids (especially M. consputa) and cerambycids (Monochamus species) despite allowing up to eight years of post-fire recruitment, suggesting that there is low incidence of local recruitment in these species. Other woodborers less attracted to heat and smoke might be more likely to colonize trees that succumb to post-fire processes (Angers et al. 2011), such as bark beetle attack.

Finally, the effects of fire ignition date deserve special consideration in further analyses of woodboring beetle ecology. We found surprisingly consistent support across life stages and taxa for negative linear and quadratic relationships between ignition date and woodborer activity, suggesting that there is a temporal peak and decline in the efficacy of woodborer colonization following forest disturbances. The relatively late-season King, Moonlight, and Rim fires provided important support for this apparent effect that should be investigated for generality. Throughout the western US, fire season has lengthened substantially in recent decades (Abatzoglou and Williams 2016; Westerling 2016), due to changes in climate as well as increasing anthropogenic ignitions, which often occur well outside the range of dates associated with lightning ignitions (Balch et al. 2017). This lengthening of the fire season is expected to continue and perhaps even accelerate in the coming decades (Yue et al. 2013, Westerling 2016), yielding more fires that ignite outside the optimal period for driving woodborer population growth and the ecosystem processes that depend on woodborers.

However, we caution that our analysis is based on only 11 fires with a relatively modest spread in ignition dates (5 July to 13 September). Only two of these fires ignited earlier than the interquartile range of lightning ignition dates reported by Balch et al. (2017) for an area that includes the Sierra Nevada, and only two others ignited later than the interquartile range. Furthermore, ignition date is at best a rough index of burn date for the specific areas that we surveyed with our transects. We therefore cannot pinpoint thresholds in ignition date that are likely to yield robust colonization by woodborers, nor can we state with certainty that the apparent effect of ignition date in this study does not stem from another process associated with site. Rather, we view the pervasive effect of ignition date in our models as a starting point for further research on this topic.

Woodboring beetle activity at our study sites was often similar between burned stands and unburned stands damaged by bark beetles, but varied dramatically with stand composition, burn severity, and the intensity of bark beetle activity associated with each type of disturbance. In burned sites, the relationship between woodborer activity and ignition date suggested that increased prevalence of wildfire might not increase woodboring beetle populations if fires increasingly occur outside the historical fire season. Effects of these patterns on ecological processes like snag longevity and woodpecker activity should be further explored to inform decisions about management of disturbed forest stands. The patterns observed in this study should be investigated for generality in other regions, and across a diversity of fires and bark beetle outbreaks, to facilitate inference in this system, which naturally exhibits high covariance and spatial clustering in the values of key predictor variables.