Older forests used by northern spotted owls functioned as fire refugia during large wildfires, 1987-2017


 Background

The northern spotted owl (Strix occidentalis caurina) is an Endangered Species Act-listed subspecies that requires forests with old-growth characteristics for nesting. With climate change, large, severe wildfires are expected to be more common and an increasing threat to spotted owl persistence. Understanding fire severity patterns related to nesting forest can be valuable for forest management that supports conservation and recovery, especially if nesting forest functions as fire refugia (i.e., lower fire severity than surrounding landscape). We examined the relationship between fire severity and nesting forests in 472 large wildfires (> 200 ha) that occurred rangewide during 1987–2017. We mapped fire severities (unburned-low, moderate, high) within each fire using relative difference normalized burn ratios and quantified differences in severity between pre-fire nesting forest (edge and interior) and non-nesting forest. We also quantified these relationships within areas of three fire regimes (low severity, very frequent; mixed severity, frequent; high severity, infrequent).
Results

Averaged over all fires, the interior nesting forest burned at lower severity than edge or non-nesting forest. These relationships were consistent within the low severity, very frequent and mixed severity, frequent fire regime areas. All forest types burned at similar severity within the high severity, infrequent fire regime. During two of the most active wildfire years that also had the largest wildfires occurring in rare and extreme weather conditions, we found a bimodal distribution of fire severity in all forest types. In those years, a higher amount—and proportion—of all forest types burned at high severity. Over the duration of the study, we found a strong positive trend in the proportion of wildfires that burned at high severity in the non-nesting forests, but not in the two nesting forest types.
Conclusions

Under most wildfire conditions, the microclimate of interior patches of nesting forests likely mitigated fire severity and thus functioned as fire refugia. With changing climates, the future of interior forest as fire refugia is unknown, but trends suggest these older forests can dampen the effect of increased wildfire activity and thus an important component of landscape plans focused on fire resiliency.


Results
Averaged over all res, the interior nesting forest burned at lower severity than edge or non-nesting forest.
These relationships were consistent within the low severity, very frequent and mixed severity, frequent re regime areas. All forest types burned at similar severity within the high severity, infrequent re regime.
During two of the most active wild re years that also had the largest wild res occurring in rare and extreme weather conditions, we found a bimodal distribution of re severity in all forest types. In those years, a higher amount-and proportion-of all forest types burned at high severity. Over the duration of the study, we found a strong positive trend in the proportion of wild res that burned at high severity in the non-nesting forests, but not in the two nesting forest types.

Conclusions
Under most wild re conditions, the microclimate of interior patches of nesting forests likely mitigated re severity and thus functioned as re refugia. With changing climates, the future of interior forest as re refugia is unknown, but trends suggest these older forests can dampen the effect of increased wild re activity and thus an important component of landscape plans focused on re resiliency.

Background
Wild re effects on individual species and wildlife communities can range from highly bene cial to strongly negative depending on species-speci c adaptability to disturbance and re characteristics such as vegetation type burned, size, return interval, seasonality, and severity (Smith 2000). For example, many wild res can be bene cial for some avian species (e.g., woodpeckers) because post-re conditions enhance forage and nesting opportunities (Hutto 2008), but wild re can remove many important habitat requirements for other species (e.g., greater sage-grouse Centrocercus urophasianus) (Coates et al. 2015, Foster et al. 2019. It is common within large wild res to have a mix of re severities, ranging from unburned-to-low severity to areas with nearly complete mortality of forest vegetation (high severity). For forest-adapted species, the effects of wild re are often more negative with increasing severity, such that low severity may be neutral or bene t a species, while high-severity re negatively affects the species (Fontaine and Kennedy 2012). At the population scale, negative effects of high severity wild re can be serious for forest wildlife facing extinction or extirpation. For example, wild res in Australia in 2020 burned critical habitat for as many as 100 threatened species (Pickrell and Pennisi 2020), and wild re is listed as one of the main threats to greater sage-grouse habitat, though rangewide habitat has been fragmented from other causes (USFWS 2015).
The northern spotted owl (Strix occidentalis caurina) inhabits coniferous forests of the Paci c Northwest of North America and for nesting (and roosting) it requires late-successional, multistoried, closed-canopy forests with large trees (Forsman et al. 1984, Wiens et al. 2014, Wilk et al. 2018. These older forests are also a critical component of foraging habitat and within these forests spotted owls have some partitioning of space use with barred owls, an important competitor and threat to spotted owls (Wiens et al. 2014, Jenkins et al. 2019b. Due primarily to loss of older forests from timber harvest, the northern spotted owl was listed as threatened in 1990 under the US Endangered Species Act and found to warrant reclassi cation to endangered in 2020 due continued population declines (USFWS 1990(USFWS , 2020. The older forests that are suitable for nesting by northern spotted owls are monitored as a component of the Northwest Forest Plan (NWFP) effectiveness monitoring program (e.g., Davis et al. 2016). Based largely on monitoring results, large wild res have been identi ed as one of the primary and increasing threats affecting spotted owl habitat , and the occurrence and extent of large wild res in the Western US is predicted to increase due to climate change (Westerling et al. 2006, Abatzoglou and Williams 2016, Davis et al. 2017, Wan et al. 2019).
High-severity re can reset forest succession and remove forest cover suitable for nesting by northern spotted owls ; resulting in negative effects on territory occupancy and survival (Clark et al. 2011, Clark et al. 2013, Rockweit et al. 2017. However, localized studies of wild re severity in landscapes with varying-aged forests and management history have shown that, compared to other forest types, old forests were at lower risk for high-severity wild re (Bradley et al. 2016, Zald andDunn 2018). Lesmeister et al. (2019) found that-for one mixed-severity wild re that burned in a drought year and during high re weather conditions-nesting forest with old forest characteristics had the lowest odds of burning at high severity compared to other forest types. However, it is unknown if those patterns were unique to that landscape and those weather conditions, and thus ndings may differ if data were used from many res over long time periods. Fire refugia is de ned as landscape elements that burn less often or severely than the surrounding landscape (Meddens et al. 2018); therefore, nesting forest may be considered re refugia if burning at lower severity is maintained over long time scales and many res.
Understanding the patterns of re severity as related to the different forest types and spatial patterns over the entire range of the northern spotted owl can provide valuable information on how to manage those forests for the species conservation and recovery. Forests used by spotted owls for nesting (i.e., nesting forest) are older forests with large trees and moderate to closed canopy (Forsman et al. 1984). We mapped nesting forest and non-nesting forest for each year of the three-decade study and quanti ed wild re severity in those forest types across large wild res. To elucidate the role nesting forest may have played as re refugia, our objectives were to: 1) examine the pre-re pattern of spotted owl nesting forest in relation to observed wild re severity; 2) compare wild re severity between spotted owl nesting forest and other forest types in the re perimeter; and 3) examine temporal trends in wild re severity in each forest type over the duration of the study. Compared to other forest types, the interior portions of nesting forest (> 30 m from an edge) typically have milder microclimates during summer with lower wind speeds and temperature, and higher humidity (Chen et al. 1995). Therefore, we hypothesized that interior forest would function as re refugia by burning at lower re severities compared to other forest types during large wild res.

Study area
We conducted our study of wild re severity within the USA portion of the range of northern spotted owls ( Fig. 1). Within this area, 472 large wild res (> 200 ha) occurred from 1987-2017 over the full range of re regimes extending across approximately 162 000 km² from western Washington to northwest California (Fig. 1a). Based on potential vegetation types, a diversity of forest ecosystems is common within this area, with old-growth conifer forests being the most common climax communities. The major biophysical driving variables of extent, structure, composition, and dynamics of these old-growth forests were climate, topography, soils, succession processes, and disturbance events Dyrness 1973, Oliver 1981). Historically, landform, soil conditions, and relatively stable regional climate resulted in somewhat predictable biotic communities, pathways of forest development, levels of ecosystem productivity, and spatial patterns of disturbance regimes (Franklin and Dyrness 1973).
Fire regimes describe the normal frequency and severity of wild res. Within the northern spotted owl range, re regimes include: low severity, very frequent; mixed severity, frequent; and high severity, infrequent (Fig. 1a, Reilly et al. 2017). We used historical re regime data from Spies et al. (2018) to explore the relationship between re severity and nesting forest rangewide and for each re regime ( Fig. 1b).

Forest type classi cation
Forests used by northern spotted owls for nesting and roosting are typically more than 125 years of age with average tree diameters at breast height > 50 cm (often high diversity of sizes and some trees are > 75 cm diameter) and multi-layered canopies with > 60% canopy cover (Davis et al. 2016). Here we refer to this forest type as nesting forest. We used maps of nesting forests that were generated by the NWFP monitoring program and have been used in many publications on spotted owl population dynamics and resource selection (e.g., Wiens et al. 2014, Dugger et al. 2016, Jenkins et al. 2019a. Within the study there were large areas not capable of developing into nesting forest, mainly due to soil type, plant association, or elevation (Davis and Lint 2005). Therefore, we restricted our classi cation of forest types to areas that possessed the abiotic and biotic characteristics to be capable of development into suitable nesting forest absent disturbances to reset successional stage (i.e., habitat capable, Fig. 1a).
Information on pre-re forest composition and structure is critical for examining relationships between forest types and wild re effects Krawchuk 2018, Lesmeister et al. 2019). We used open source software Maxent (Phillips et al. 2006, Phillips et al. 2017) to model nesting forest following NWFP monitoring methods within the habitat capable areas (Fig. 1a, Davis et al. 2011, Davis et al. 2016). Using Google Earth Engine (Gorelick et al. 2017), we applied nesting forest algorithms to a Landsat-based (30m pixel resolution) annual time series (1987-2017) of forest structure and species composition maps (Bell et al. 2021). The resulting annual maps of nesting forest spanned all years analyzed here. We classi ed the maps into binary maps of nesting forest and further processed using open source software GUIDOS (Soille and Vogt 2009) to produce maps relating to the spatial con guration of nesting forest. We classi ed the nesting forest pixels as either INTERIOR or EDGE forest (Fig. 2). The INTERIOR forest pixels were > 30m from NON-NESTING forest and the EDGE forest pixels were adjacent to ≥ 1 NON-NESTING forest pixel(s). The NON-NESTING pixels were forests not suitable for nesting and were primarily younger forests, thinned older forest, or pre-forest conditions (Davis et al. 2016). The smallest patch size of nesting forest that could contain an INTERIOR class was a 3x3 pixel con guration (0.81 ha), large enough to contain microclimates distinct from NON-NESTING forests (Heithecker and Halpern 2007).

Wild re data
Northern spotted owl territories are on average 700 ha (range 180 to 1 390 ha) in size (Dugger et al. 2016), so we focused on wild res that were ≥ 200 ha in size, large enough to impact > 25% of an average territory. These 472 wild res totaled 20 970 km², with 17 273 km² burned in habitat-capable forests (Fig. 3). Using wild res of ≥ 200 and analyzing a 30-year period allowed us to examine re severity encompassing varying forest cover types and arrangements, as well as temporal trends in severity.
We used a Landsat-based time series (1986-2017) of forest disturbance maps produced by the Landscape Change Monitoring System (LCMS; Healey et al. 2015) to measure extent and severity of wild re. These maps used forest disturbance data collected with TimeSync software  and an ensemble LandTrendr disturbance mapping algorithm (Cohen et al. 2018, Healey et al. 2018) to produce annual disturbance maps with magnitude quanti ed by relativized difference in the normalized burn ratio (RdNBR) (Miller and Thode 2007). We used Reilly et al. (2017) classi cations of re severity based on RdNBR within re perimeters for unburned-low (RdNBR < 235), moderate (RdNBR 235-649), and high (RdNBR ≥ 649) severity classes (Appendix 1).

Wild re selection ratios
We selected wild res with ≥ 50% of the forested area within their perimeters classi ed as habitat capable (n = 472; 17 273 km²) to compare re severity relationships in INTERIOR and EDGE to those in NON-NESTING forest types. Most of the wild res had > 90% of the area within their perimeter comprised of habitat capable forest. We used selection ratios (Manly et al. 2002) to compare wild re severity in our three forest types, taking into account the proportion of each forest type within each wild re perimeter (Moreira et al. 2001, Moreira et al. 2009. We de ned our selection ratios as the area burned:area available for burning (B/A) ratio. We estimated B/A for forest type i burning at severity class j (w ij ) by w ij = o ij / π i , where o ij = the proportion area burned at severity j that was forest type i, and π i is the proportion of forest type i available to burn (i.e., within wild re perimeter). Values for w ij = 1 indicated the forest type burned at a given severity in proportion to its availability, w ij > 1 indicated the forest type burned at a given severity greater than expected by chance, and w ij < 1 indicated the forest type burned at a given severity less than expected.
We calculated the mean B/A ratios and 95% con dence intervals for all 472 wild res rangewide and within areas of the three re regimes (low severity, very frequent; mixed severity, frequent; high severity, infrequent). We used the amount of overlap in 95% con dence intervals to evaluate for signi cant differences in B/A ratios for re severity and forest type combinations. For example, if con dence intervals for B/A ratios did not overlap 1, we considered the area in each forest type to have burned at a given severity more or less than expected by chance. Due to non-normal distribution of B/A ratios, we also conducted a Tukey post hoc comparison of contrasts between re severity and forest types.

Fire severity patterns and trends
For each of the three forest types, we calculated the annual proportion of area burned at each of the three re severities. We used linear regression to analyze long-term trends in yearly proportion of each forest type burning at high-severity re. We considered slope estimates with 95% con dence intervals not overlapping 0 to indicate strong evidence of a trend in average percent of high-severity re.
We examined normalized burned area frequency distribution patterns of observed re severity based on RdNBR by forest type using kurtosis and skew statistics for the four wild re seasons with the most area burned during our observation period: 1987, 2002, 2008, and 2017. We interpreted skewness values of > 1.0 or <-1.0 to indicate a substantially skewed distribution in RdNBR by forest type. Increasing positive skewness indicated greater frequency of a forest type burning at lower severity classes, while negative skewness indicated greater frequency of burning in higher severity classes. Higher kurtosis values in RdNBR indicated narrow distribution with a given severity and lower kurtosis suggested more at distribution over re severities (Thode et al. 2011, Sugihara et al. 2018).
Of the 472 res, 307 had all or a portion of the perimeter in the area of low severity-high frequency regime, 309 in mixed severity-moderate and high frequency, and 114 in high severity-low frequency re regimes. In the low severity-high-frequency regime, INTERIOR forest had higher odds of burning at low severity  Fig. 4B). The NON-NESTING forest had low odds of burning at unburned-low severity (B/A = 0.95, 95% CI = 0.93-0.96) but was more likely to burn at moderate (B/A = 1.06, 95%, CI = 1.05-1.08) or high severity (B/A = 1.05, 95% CI = 1.02-1.08; Fig. 4B).
For res in the high severity-low frequency re regime, the INTERIOR forest burned at high severity less than expected (B/A = 0.82, 95% CI = 0.70-0.93), but 95% CIs overlapped 1.0 at the two lower re severities (Fig. 4D). The EDGE forest burned at low odds of burning at high severity (B/A = 0.89, 95% CI = 0.80-0.98) and unburned-low severity (B/A = 0.91, 95% CI = 0.85-0.96), but high odds of burning at moderate severity (B/A = 1.09, 95% CI = 1.02-1.16). The 95% CIs for the NON-NESTING forest overlapped 1.0 for all three severity classes. A Tukey post-hoc comparison of B/A ratios among severity classes and forest types indicated that INTERIOR forest tended to burn at unburned-low severity compared to EDGE and NON-NESTING forests (Appendix 2).

Fire severity patterns and trends
The number of res and area burned varied greatly among years studied, with higher number of res corresponding with more area burned (Fig. 3) The proportion of area burned each year differed among years for all forest types (Fig. 5). For most years, the proportion of area burned at high severity was less than area burned at moderate or unburned-low severity (Fig. 5). All forest types had some evidence of an increasing linear trends in the average yearly percent of area burned at high severity (Fig. 5), but only in the NON-NESTING forest was there strong evidence of an increase (Fig. 5D). The slope estimate for NON-NESTING forest indicated a 0.7% (95% CI = 0.29-1.05%) annual increase in average area burned at high severity.
For each of the four largest wild re seasons, each burning over 200 000 ha of habitat-capable forests, the re severity frequency distribution patterns differed between forest types (Fig. 6). Frequency distributions for INTERIOR were consistently most positively skewed (2.3-3.3) and had greatest kurtosis (5.0-10.9) toward low severity, with most of the area burning at lower severities (Fig. 6). Albeit less pronounced than for INTERIOR, EDGE forest was positively skewed (1.1-2.6) and had greater kurtosis (1.1-2.6), exhibiting a low to moderate severity pattern (Fig. 6). Skew and kurtosis for EDGE was intermediate to INTERIOR and NON-NESTING. For NON-NESTING forest, skewness was moderately positive (0.8-1.4) and little kurtosis (-1.2-0.6), indicating a relative even distribution across the RdNBR spectrum (Fig. 6). Fire severity frequency distributions were the most bimodal during the 2002 and 2017 re seasons (Fig. 6). These were the years with two largest wild res during our study period (2002 Biscuit Complex and 2017 Chetco Bar Fire) and had the highest area burned per wild re (Fig. 3).

Discussion
Here we analyzed the likelihood of different forest types burning at different re severities during 472 large wild res that occurred over a span of 30 years throughout the range of northern spotted owls in the Paci c Northwest, USA. The spatial and temporal expanse of our dataset and the ability to generate annual maps of spotted owl nesting forest afforded us the ability to gain unprecedented insights into the function of nesting forest as re refugia. No other study of wild re and forests used by northern spotted owls match the large number of res, geographic extent, or number of years evaluated in this study. Large wild res are a severe threat to northern spotted owl habitat and populations (Clark et al. 2011, Davis et al. 2011, Rockweit et al. 2017), yet the issue has been debated in the scienti c literature, especially when also considering other spotted owl subspecies (Hanson et al. 2009, Spies et al. 2010, Ganey et al. 2017, Jones et al. 2020, Lee 2020. In addition to wild re, multiple other stressors play a role in degrading the prognosis for persistence of northern spotted owl populations (Dugger et al. 2016, Miller et al. 2018, Jenkins et al. 2019a, Wiens et al. 2019). We approached this study to better understand the long-term and broadscale patterns of risk that high-severity re pose to spotted owls and their habitat because the extent and frequency of wild res that is expected to increase with climate change (Davis et al. 2017, Halofsky et al. 2020. We observed consistent patterns of re severity in different forest types used by this old forest obligate and nesting forest played an important role as re refugia in the face of increasing wild re activity. Our ndings from broadscale and long-term data were similar to those from wild res (Douglas Complex) that burned in a mixed-ownership landscape of the Klamath-Siskiyou ecoregion of southwestern Oregon, USA (Zald andDunn 2018, Lesmeister et al. 2019). The Douglas Complex burned an area of 38 000 ha in mixed-severity with large patches of high-severity re. Older forests in late-successional reserves (i.e., suitable nesting forest) burned at lower severity despite having higher fuel loading than other forest types within the re perimeters . Ownership patterns were also a strong predictor of re severity for the Douglas Complex, where federally managed lands were primarily comprised of latesuccessional forest reserves that burned at lower severity compared to plantation forests on private timber industry lands (Zald and Dunn 2018). Those studies suggested that, in addition to contribution to northern spotted owl conservation, older forests functioned as re refugia and had an added bene t of buffering the effects of climate change-induced increases in wild re occurrence.
In our study, interior older forests burned at lower severity than other forest types, especially when compared to the non-nesting forest type that was primarily younger and open-canopied forests that were prone to burn at higher severities. Edges and fragmented nesting forest burned at intermediate severities, with edges presumably buffering interior forest from higher re severity in non-nesting forest. We hypothesize the mechanism driving these re severity patterns was the long-known relationship (see Hursh andConnaughton 1938, Countryman 1955) between differing microclimates of forests and susceptibility to high-severity wild re. In the Paci c Northwest, closed-canopy, structurally complex latesuccessional coniferous forests with high biomass (e.g., spotted owl nesting forest) maintain cooler, more temperate microclimates and provide an insulating effect on temperatures (Chen et al. 1995, Frey et al. 2016. Fire behavior and severity is largely driven by interactions of wind, humidity, temperature, fuels, and topography (Countryman 1964, Thompson and Spies 2009, Halofsky et al. 2011. Younger and opencanopied forests have hotter, drier, and windier microclimates and those conditions decrease dramatically over relatively short distances into the interior of older closed-canopy forests with high tree density (Chen et al. 1995, Heithecker and Halpern 2007, Arroyo-Rodríguez et al. 2016).
We found an increasing trend in proportion of annual area burned by high-severity re over the duration of our study, but the trend occurred most strongly in the non-nesting forest type, suggesting that the effects of climate change on the occurrence of high-severity wild res may be most pronounced in younger and open canopy forests. Interior nesting forest, functioning as re refugia, buffered potential negative effects of high-severity re during most years of the 30-yr period. Forests functioning as re refugia can support ecosystem resilience to disturbances as well as post re ecosystem recovery and biodiversity (Meddens et al. 2018). Our ndings are consistent with recent research that found higher extents and quality of re refugia in closed-canopy older forests compared to younger and more opencanopied forest cover types Krawchuk 2018, Andrus et al. 2021). Compared to other vegetation types, late-successional forests have a higher likelihood of burning at lower re severities (Meigs et al. 2020), even during high-re weather conditions during drought years ). These interior forests were re refugia during our observed timespan, but it remains unknown if they are ephemeral refugia or will function as persistent refugia into the future with a changing climate.
Mature forests have higher resiliency to re effects and climate variability, especially when not subject to fragmentation in a matrix of young ammable patches that can shift mature forests to an alternative steady state more prone to repeat high-severity re (Kitzberger et al. 2012). Similarly, examining forests in Australia, Duff et al. (2018) showed that older forests had higher resilience to drought conditions that increased ammability of vegetation, thus functioned as re refugia. Intact old forest with less fragmentation in Amazonian forests also function as refugia by ameliorating the effects of re (Silva Junior et al. 2018, Maillard et al. 2020. In some years with extremely large wild res there was a bimodal distribution in re severity in all forest types, potentially degrading the function of older forest as re refugia. The 2002 re season was dominated by the Biscuit Fire, which at over 200,000 ha was the largest re in our study. The 2017 re season had the greatest amount of area burned of the years we sampled and was dominated by the Chetco Bar Fire which burned over 190,000 ha. The bimodal patterns we observed in these 2 years were consistent with theorized re severity distributions when extremely large res (i.e., mega res), that occur infrequently, produce large patches of high-severity burns (van Wagtendonk and Fites-Kaufman 2006). The primary factors in the extent and severity of the 2002 and 2017 mega res was strong dry foehn winds with katabatic heating that carried westward from high-density air from higher elevations in the deserts east of the Cascade Mountains (Ustin et al. 2009, Halofsky et al. 2011. Extreme wind events occurring episodically are also primary predictor of spatial variation in large wild res in other regions (e.g., Moritz et al. 2010). These rare and extreme weather conditions have been the primary driver of the most well-known mega res during recorded history of the region, including the 1902 Yacolt burn, 1933 Tillamook burn, 1936 Bandon re, (Dague 1930, 1934, Martin et al. 1974, Herring and Greene 2001, Zyback 2004, Potter 2012. One or a few very infrequent, wind-driven crown res can shift severity distributions to more high-severity, creating a bimodal distribution (Thode et al. 2011, Cansler andMcKenzie 2014). Early evidence from recent mega res occurring in the most extreme years suggests there may be a further shift to a at distribution among re severities. In September 2020 ve mega res in Oregon burned about 329 000 ha in relatively at distribution of severity (low = 31%, moderate = 28%, high = 42%) during a sustained and historic wind storm that caused the record-setting re season (Antczak et al. 2020, Higuera and Abatzoglou 2020, R. J. Davis unpublished data, Mass 2020. In all of these mega res, extreme easterly foehn winds resulted in extraordinary re growth in all forest types regardless of management history.
Timber harvest remains the primary threat to forests used by northern spotted owls rangewide ), but on federal lands managed under the Northwest Forest Plan, the threat from wild re is now greater than the threat from timber harvest (Davis et al. 2016). These are concerning trends, especially considering that the extent and frequency of large wild res is expected to increase with climate change (Davis et al. 2017, Wan et al. 2019). Forest management plans-even some with stated goals to enhance northern spotted owl conservation-may seek to reduce wild re risk by thinning forest stands of all ages using practices that modify forest structure by increasing canopy base height, reducing crown contiguity and bulk density, and reducing forest fuels. These actions can degrade suitability of the forest for nesting by spotted owls and may decrease wild re severity in the short term (Agee and Skinner 2005, Martinson and Omi 2013, Kalies and Yocom Kent 2016, Prichard et al. 2020), but are less effective at reducing wild re extent and severity on a large scale beyond a short time window (Stone et al. 2003, Reinhardt et al. 2008, Barnett et al. 2016, Schoennagel et al. 2017. Converting older, closed-canopy forests that function as re refugia to more open, managed forests does not assure a dampening effect on wild re severity (Zald and Dunn 2018), due in part to the complex changes in the microclimate of forest stands after thinning. Fuel loads and arrangement are a component of the re environment, but forest thinning actions can alter microclimates to increase ammability. Variable retention harvesting, which aims to mimic natural forest disturbance regimes and retains old forest structure, including snags and logs, is becoming more commonplace (Franklin and Donato 2020). These silvicultural prescriptions may retain enough forest structure to function as edge nesting forest and thus less prone to high-severity re than non-nesting forest. These actions may be especially effective if the resulting landscape has extensive areas of interior nesting forest.

Conclusions
We present evidence that spotted owl nesting forest does not burn at higher severity than non-nesting forest types, so there is reason for optimism that forests used by spotted owls will persist with increased occurrence of wild re. We do not infer that our results trivialize the threat to spotted owls from large wild res; when mega res occur, spotted owl populations are negatively affected . Although high severity res have been an important ecological process in Paci c Northwest forests for at least 11 000 years with frequent res steadily increasing over the past 4 000 years (Walsh et al. 2015), contemporary mega res are beyond what is known to occur during recorded history. Extremely dry and windy conditions set the stage for large, high-severity res in moist coastal forests (Bessie andJohnson 1995, McKenzie et al. 2004) and if those weather events become commonplace with climate change, spotted owl persistence will be further threatened. Rather than minimize the risk from rare mega res, our results support the hypothesis and mounting evidence that under most re weather, nesting forests are at lower risk of high-severity wild re compared to the surrounding landscape. Nesting forest may be at increased risk to high-severity re when fragmented and surrounded by non-nesting forest (primarily younger forests). Our results support guidelines in the 2011 spotted owl recovery plan that describe targeting restoration projects outside of current spotted owl habitat (USFWS 2011). Projects that occur in non-nesting forest may accomplish many forest restoration goals aimed at improving the resiliency of late-successional forests to wild res and climate change without the negative impact on nesting forests while maintaining re refugia on the landscape.

Declarations
Ethics approval and consent to participate Not applicable Consent for publication Not applicable Availability of data and materials If this paper is accepted, we will make the datasets used in our analyses available in a public repository.

Competing interests
The authors declare that they have no competing interests.

Funding
Funding for this project was provided by USDA Forest Service, Paci c Northwest Research Station, Portland, OR, USA.
Author's contributions DBL developed the analysis concept, provided funding, interpreted statistics, and majority of manuscript writing; RJD developed the analysis concept, compiled data, gure creation, and contributed to manuscript writing; SGS performed GIS analysis, compiled and analyzed data, and contributed to writing the manuscript; ZY developed the original spatial datasets, assisted in interpreting statistics, and contributed to writing the manuscript. Figure 1 The range of the northern spotted owl range in the USA. Map a: habitat capable forest and coverage of large (> 200 ha) wild res from 1987-2017. Habitat capable forests were those areas with environmental conditions of elevation (below 2000 m) and soil types that without disturbance (e.g., timber harvest) could develop into suitable forest for nesting and roosting by spotted owls given time for succession.

Figures
Map b: the extent of three historical re regimes modi ed from Spies et al. (2018).

Supplementary Files
This is a list of supplementary les associated with this preprint. Click to download.