Adapting western North American forests to climate change and wildfires: 10 common questions

Fuel treatments and active forest management

Are the effects of fire exclusion overstated? If so, are treatments unwarranted and even counterproductive?

Concerns about forest thinning and other forms of active management are sometimes based on the assumption that contemporary conditions and fire regimes in dry pine and mixed‐conifer forests are not substantially departed from those maintained by uninterrupted fire regimes (Hagmann et al. ²⁰²¹). This perspective does not accurately reflect the breadth and depth of scientific evidence documenting the influence of over a century of fire exclusion. Support for the suggestion that ecological departures associated with fire exclusion are overestimated has repeatedly failed independent validation by multiple research groups (Hagmann et al. ²⁰²¹). In addition, these arguments fail to consider widespread Indigenous fire uses that affected landscape scale vegetation conditions linked to valued cultural resources and services, food security, and vulnerability to wildfires (Lake et al. ²⁰¹⁸, Power et al. ²⁰¹⁸). As is explored in the following sections, a number of forest management and treatment strategies are shown to be highly effective. Site conditions and history are always important considerations. Moreover, there is no one‐treatment‐fits‐all approach to forest adaptation.

Evidence from a broad range of disciplines documents widespread, multi‐regional 20^th‐century fire exclusion in interior forested landscapes of wNA (see a detailed reference list and discussion in Hagmann et al. ²⁰²¹). Collectively, these studies reveal extensive changes in tree density, species and age composition, forest structure, and continuity of canopy and surface fuels. Forests that were once characterized by shifting patchworks of forest and nonforest vegetation (i.e., grasslands, woodlands, and shrublands) in the early 20th‐century gradually became more continuously covered in forest and densely stocked with fuels (Fig. 4).

However, for over two decades, a small fraction of the scientific literature has cast doubt on the inferences made from fire‐scar based reconstructions and broader landscape‐level assessments to suggest that estimates of low‐ to moderate‐severity fire regimes from these studies are overstated. Hagmann et al. (in press) examine this counter‐evidence in detail and identify critical flaws in reasoning and methodologies in original papers and subsequent re‐application of these methods in numerous geographic areas. Subsequent research shows that studies relying on Williams and Baker (2011) methods for estimating historical tree densities and fire regimes overestimate tree densities and fire severity (see also Levine et al. ²⁰¹⁷). Moreover, established tree‐ring fire‐scar methods more accurately reconstruct known fire occurrence and extent. Other studies, also based on the methods of Williams and Baker (2011), conflate reconstructed low‐severity, high‐frequency fire regimes with landscape homogeneity. These interpretations disregard critical ecosystem functions that were historically associated with uneven‐aged forests embedded in multi‐level fine‐, meso‐ and broad‐scale landscapes. By extension, claims that low‐severity fire regimes are overestimated then imply that large, high‐severity fires were a regular occurrence prior to the era of European colonization. Such interpretations may lead to the conclusion that recent increases in high‐severity fire are still within the historical range of variability, and that there is no need of restorative or adaptive treatments (Hanson and Odion 2014, Odion et al. ²⁰¹⁴, Baker and Hanson ²⁰¹⁷).

Indeed, research from across wNA has shown that high‐severity fire was a component of historical fire regimes, and that fires of all severities are currently in deficit (Parks et al. ^2015b , Reilly et al. ²⁰¹⁷, Haugo et al. ²⁰¹⁹, but see Mallek et al. ²⁰¹³). However, reanalysis of the methods of Baker and others shows that their methods inherently overestimate fire severity and the frequency and area affected by high‐severity fire (Fulé et al. ²⁰¹⁴, Hagmann et al. ²⁰²¹). In addition, high‐severity patches in recent fires are less heterogeneous and more extensive than the historical range of variability for forests characterized by low‐ and moderate‐severity fire regimes (Stevens et al. ²⁰¹⁷, Hagmann et al. ²⁰²¹). Finally, research across wNA reveals key climate‐vegetation‐wildfire linkages, where fire frequency, extent, and severity all increase with increasing climatic warming, suggesting that observed trends in fire patterns are commensurate with predicted relationships with ongoing climate change (McKenzie and Littell ²⁰¹⁷, Parks and Abatzoglou ²⁰²⁰).

Another perspective on this debate contends that whether historical records can be agreed upon is of ancillary importance. Adaptive forest management and fuel reduction treatments are primarily aimed at increasing forest resilience and/or resistance to climate change, fire and other disturbances, which has positive societal and ecological impacts that do not require justification based on historical conditions, particularly given the no‐analog present and future that climate change presents (Freeman et al. ²⁰¹⁷). For example, the most concerning contemporary high‐severity fire events are associated with large patches of complete stand replacement (Miller and Quayle 2015, Lydersen et al. ²⁰¹⁶). In some cases, high‐severity fire events convert forests to shrubland and grassland assemblages as alternative stable states in uncharacteristically large patches (Falk et al. ²⁰¹⁹, Kemp et al. ²⁰¹⁹, Stevens‐Rumann and Morgan ²⁰¹⁹). As such, a critical forest management concern is that high‐severity wildfires are accelerating rates of vegetation change, forest conversion, and vulnerability of native habitats in response to a warming climate.

Is forest thinning alone sufficient to mitigate wildfire hazard?

While “thin the forest to reduce wildfire threat” is commonly cited in the popular media, the capacity for thinning alone to mitigate wildfire hazard and severity is not well supported in the scientific literature. Thinning treatments require strategic selection of trees to target fuel ladders and fire‐susceptible trees, along with a subsequent fuel reduction treatment (Jain et al. ²⁰²⁰). When thinning is conducted without accompanied surface fuel reduction, short and long‐term goals may not be realized.

Thinning from below reduces ladder fuels and canopy bulk density concurrently, which can reduce the potential for both passive and active crown fire behavior (Agee and Skinner 2005). For instance, Harrod et al. (2009) found that thinning treatments that reduced tree density and canopy bulk density and increased canopy base height significantly reduced stand susceptibility to crown fire compared to untreated controls. Furthermore, large‐diameter trees and snags that provide essential wildlife habitat and other ecosystem values can be retained and fuels can be deliberately removed around these structures using this approach (Lehmkuhl et al. ²⁰¹⁵). Where wood from treatments can be marketed, revenues from thinning help to sustain broader management goals on public lands. For example, some landscape restoration collaboratives seek to reinvest profits from commercially viable thinning to off‐set costs associated with more labor‐intensive manual thinning and prescribed or cultural burning needs (Schultz and Jedd 2012).

Some studies show that thinning alone can mitigate wildfire severity (e.g., Pollet and Omi ²⁰⁰², Prichard and Kennedy ²⁰¹⁴, Prichard et al. ²⁰²⁰), but across a wide range of sites, thin and prescribed burn treatments are most effective at reducing fire severity (see reviews by Fulé et al. ²⁰¹², Martinson and Omi ²⁰¹³, Kalies and Yocom Kent ²⁰¹⁶). On most sites, thinning alone achieves a reduction of canopy fuels but contributes to higher surface fuel loads. If burned in a wildfire, these fuels can contribute to high‐intensity surface fires and elevated levels of associated tree mortality (e.g., Stephens et al. ²⁰⁰⁹, Prichard and Kennedy ²⁰¹²). When trees are felled and limbed, fine fuels from tree tops and branches (termed activity fuels) are re‐distributed over the treatment area, thereby increasing surface fuel loads (Martinson and Omi 2013). Mechanical fuel reduction treatments of these activity fuels are possible, but in many locations, biomass removal and utilization (e.g., for bioenergy) after thinning treatments can be cost‐prohibitive due to long hauling distances and the economic and technological challenges of building new biomass facilities (Hartsough et al. ²⁰⁰⁸). Mastication equipment is sometimes used to shred understory trees and shrubs into smaller woody fragments, which are then redistributed and left on site (Kane et al. ²⁰⁰⁹). However, following mastication, surface fuels are temporarily elevated, and masticated stands that burn in wildland fires can cause deep soil heating from long‐duration smoldering combustion and elevated fire intensities (Kreye et al. ²⁰¹⁴).

Other unintended consequences of thinning without concomitant reduction in surface fuels can occur. For instance, decreasing canopy bulk density can change site climatic conditions (Agee and Skinner 2005). Wildfire ignition potential is largely driven by fuel moisture, which can decrease on drier sites when canopy bulk density is reduced through commercial thinning (e.g., Reinhardt et al. ²⁰⁰⁶). Reduced canopy bulk density can lead to increased surface wind speed and fuel heating, which allows for increased rates of fire spread in thinned forests (Pimont et al. ²⁰⁰⁹, Parsons et al. ²⁰¹⁸). Other studies show no effect of thinning on surface fuel moisture (Bigelow and North 2012, Estes et al. ²⁰¹²), suggesting that thinning effects on surface winds and fuel moisture are complex, site specific, and likely vary across ecoregions and seasons.

In summary, although the efficacy of thinning alone as a fuel reduction treatment is questionable and site dependent, there exists widespread agreement that combined effects of thinning plus prescribed burning consistently reduces the potential for severe wildfire across a broad range of forest types and conditions (Fig. 3; Fulé et al. ²⁰¹², Kalies and Yocom Kent ²⁰¹⁶, Stephens et al. ²⁰²¹). Given this broad consensus in the scientific literature, some authors suggest that forest thinning should be considered in the context of wildfire hazard abatement, ecological restoration and adaptation, and revitalization of cultural burning (Lehmkuhl et al. ²⁰⁰⁷, Hessburg et al. ²⁰¹⁵, Huffman et al. ²⁰²⁰). Where restoring resilient forest composition and structure and reducing future wildfire hazard are goals of management (Koontz et al. ²⁰²⁰), combined thinning and burning approaches will provide ecological and wildfire‐risk reduction benefits (Knapp et al. ²⁰¹⁷).

Can forest thinning and prescribed burning solve the problem?

Fire has been a tool that has been actively used for millennia. Indigenous burning practices maintained prairies, oak and pine savannas, riparian areas, mixed‐conifer, hardwood, and dry forests, and high mountain huckleberry and beargrass assemblages for food, medicine, basketry and other resources (Trauernicht et al. ²⁰¹⁵, Roos et al. ²⁰²¹). Following prolonged fire exclusion, many seasonally dry forest landscapes that were once frequently burned now are densely stocked with multi‐layered canopies that often require thinning prior to restoring fire (North et al. ²⁰¹², Ryan et al. ²⁰¹³). Prescribed burning on its own and in combination with mechanical thinning are essential fuel reduction treatments with demonstrated effectiveness in reducing fire severity, crown and bole scorch, and tree mortality compared to untreated forests (Safford et al., ^{2012a, 2012b}, Kalies and Yocom Kent ²⁰¹⁶). Thinning and burning in partnership with local Indigenous knowledge and practice can support culturally valued practices, traditions, livelihoods, and food and medicine security (Sowerwine et al. ²⁰¹⁹).

Although the use of prescribed burning, often in combination with mechanical thinning, has been shown to be highly effective at mitigating wildfire severity and increasing forest resilience to drought, insects and disease (Hood et al. ²⁰¹⁵), these treatments alone cannot address forest management challenges across wNA. Fuel reduction treatments are not appropriate for all conditions or forest types (DellaSala et al. ²⁰⁰⁴, Reinhardt et al. ²⁰⁰⁸, Naficy et al. ²⁰¹⁶). In some mesic forests, for instance, mechanical treatments may increase the risk of fire by increasing sunlight exposure to the forest floor, drying surface fuels, promoting understory growth, and increasing wind speeds that leave residual trees vulnerable to wind throw (Zald and Dunn 2018, Hanan et al. ²⁰²⁰). Furthermore, prescribed surface fire is difficult to implement in many current mesic forests since fire readily spreads into tree crowns via abundant fuel ladders and can result in crown fires. In other forest types such as subalpine, subboreal, and boreal forests, low crown base heights, thin bark, and heavy duff and litter loads make trees vulnerable to fire at any intensity (Agee 1996, Stevens et al. ²⁰²⁰). Fire regimes in these forests, along with lodgepole pine, are dominated by moderate‐ and high‐severity fires, and applications of forest thinning and prescribed underburning are generally inappropriate. However, landscape burning and maintenance of high elevation forests and meadows is part of cultural burning, and high‐intensity crown fire is used operationally on national forests and parks within the United States and Canada for landscape restoration objectives (Table 2).

Even where socially and ecologically appropriate, thinning and low‐intensity prescribed burning generally require repeated treatments to meet fuel reduction objectives. For example, without prior thinning, low‐intensity prescribed fire, on its own, may not consume enough fuel or cause enough tree mortality to change forest structure and reduce crown fire hazard (e.g., Lydersen et al. ^2019b ). In contrast, prescribed burns in heavy slash may result in high tree mortality. The first harvest entry into fire‐excluded stands often leaves high surface fuel loads and dense understories that require one or more prescribed burning treatments to reduce surface and ladder fuels (Goodwin et al. ²⁰¹⁸, Korb et al. ²⁰²⁰). Thus, it often takes multiple treatments and/or fire entries, as well as ongoing maintenance, to realize resilience and adaptation goals (Agee and Skinner 2005, Stevens et al. ²⁰¹⁴, Goodwin et al. ²⁰²⁰). Given the extent and variability of forest ecosystems that have experienced prolonged fire exclusion, active forest management can be only one tool to increase adaptation to climate and future fires.

Although thinning and prescribed burning have been shown to be highly effective, the current scale and pace of these treatments do not match the scale of the management challenge (Barnett et al. ^2016b , Kolden ^{2019, 2019}). Mechanical treatments are constrained by land management allocations and their enabling legislation (e.g., wilderness and roadless areas), operational constraints (e.g., steep slopes, distance to roads, costs), and administrative boundaries (e.g., riparian areas, areas managed for species of concern). In the central Sierra Nevada for example, these constraints, combined with large areas of non‐productive timberland that are unsuitable for commercial treatment due to steep slopes or distance from roads, left only 28% of the landscape available for mechanical thinning and prescribed burning treatments (North et al. ^2015a ). In the remaining area, prescribed burning alone and/or use of managed wildfires may be suitable replacement treatments (Boisramé et al. ²⁰¹⁷, Barros et al. ²⁰¹⁸). However, prescribed fire‐only treatments are frequently limited by cost, liability, air quality regulations, equipment availability, personnel capacity and training, and the need for ongoing maintenance treatments (Quinn‐Davidson and Varner ²⁰¹², Schultz et al. ²⁰¹⁹).

In light of these constraints, some researchers and managers have called for the expanded use of landscape‐scale prescribed burns and managed wildfires in addition to fuel reduction treatments as a promising approach to expand the pace and scale of adaptive management (Question 5). Increasingly collaborative restoration partnerships with Indigenous cultures can increase opportunities for re‐instating tribal stewardship practices (Lake et al. ²⁰¹⁸, Long and Lake ²⁰¹⁸). Under appropriate weather and safety conditions, and where infrastructure is not at risk, managed wildfire may serve as a useful and cost‐effective tool for reintroducing wildfire to fire‐excluded forests and achieve broad‐scale management goals.

Should active forest management, including forest thinning, be concentrated in the wildland urban interface (WUI)?

A question often asked by land managers is where to locate fuel treatments to maximize their advantage while minimizing adverse impacts. The 2000 National Fire Plan (USDA and USDI ²⁰⁰¹) and the 2002 Healthy Forests Initiative identified the need to reduce wildfire risk to people, communities, and natural resources. The 2003 Healthy Forests Restoration Act (HFRA, Congress.gov, ²⁰²⁰) then specified that >50% of fuel reduction funding be spent on projects within the Wildland Urban Interface (WUI), and it reduced environmental review within 2.41 km (1.5 miles) of at‐risk communities. The significant increase in homes lost and suppression dollars spent in the WUI in subsequent years (Mell et al. ²⁰¹⁰) has catalyzed extensive research on the WUI environment and population expansion into wildlands (Radeloff et al. ²⁰¹⁸). Subsequent studies demonstrating fuel treatment effectiveness in the WUI (Safford et al. ²⁰⁰⁹, Kennedy and Johnson ²⁰¹⁴) and spatial methods for optimizing WUI fuel treatments (Bar Massada et al. ²⁰¹¹, Syphard et al. ²⁰¹²) could be taken to suggest that most fuel reduction should be implemented in the WUI to protect homes and lives.

However, prioritizing the WUI‐only for fuel reduction treatments is often too narrow in scope to address broader landscape‐scale objectives. For example, Schoennagel et al. (2009) found that more than two‐thirds of the area within a 2.5 km radius of at‐risk communities was privately owned and unavailable for federally funded fuel treatments. This finding partly elucidates why most hazard reduction fuel treatments are implemented outside of HFRA designation. Fuel treatments on federal lands near communities may also be significantly more difficult, expensive, and risky to implement, while air quality regulations and associated risks create disincentives to treating near homes. Alternatively, agencies may be able to meet both annual prescribed burning accomplishment targets and ecological objectives in areas more distant from the WUI with fewer risks, less money, and fewer personnel (Kolden and Brown 2010, Schultz et al. ²⁰¹⁹). Further, there is increasing evidence that treating fuels across larger spatial extents in strategically planned wildland locations, rather than immediately adjacent to WUI, can indirectly reduce risk to communities (Smith et al. ²⁰¹⁶, Bowman et al. ²⁰²⁰). Benefits of this strategy include increased initial attack and short‐term suppression effectiveness, reduced crown fire potential and ember production, reduced smoke impacts to communities, and increased forest resilience (Ager et al. ²⁰¹⁰, Stevens et al. ²⁰¹⁶).

Fuel reduction treatments also can support cultural, ecological, ecosystem service, and management objectives beyond the WUI. For example, treatments that restore the ecological resilience of old‐growth forests and patches with large and old trees are critical to long term maintenance of wildlife habitats (Hessburg et al. ²⁰²⁰) of seasonally dry forests and terrestrial carbon stocks, and slowing the feedback cycle between fire and climate change (Hurteau and North 2009). Treatments in watersheds that are distant from the WUI and protect municipal and agricultural water supplies are critical to minimizing high‐severity fire impacts that can jeopardize clean water delivery (Bladon 2018, Hallema et al. ²⁰¹⁸). For example, post‐fire erosion and debris flows may cause more detrimental and longer‐term impacts to watersheds than the wildfires themselves (Jones et al. ²⁰¹⁸, Kolden and Henson ²⁰¹⁹).

Finally, treated areas outside the WUI can serve as defensible positions for fire suppression personnel that can be used to establish control lines or allow for more flexible suppression strategies, freeing up resources to protect WUI infrastructure or forests in another area (Thompson et al. ²⁰¹⁷), or can support rapid and organized evacuation when they are implemented along evacuation routes (Kolden and Henson 2019). Across complex landscapes, it is more effective in the long‐term to prioritize fuel treatments that maximize benefits across large areas and over long time frames, rather than constrain them to the WUI.

Can wildfires, on their own, do the work of fuel treatments?

The use of managed wildfires and co‐managing incidents (e.g., suppressing in some areas, and allowing other areas to burn) is increasingly promoted in the scientific literature (Stephens et al. ²⁰¹⁶, Moreira et al. ²⁰²⁰). Managed wildfires are particularly appropriate in backcountry areas where lack of road access, steep topography, firefighter safety concerns, or management designations limit opportunities for active management (Hessburg et al. ²⁰¹⁶, Huffman et al. ²⁰²⁰). However, in many cases the effects of fire exclusion on increased tree density, layering, surface fuels, and fuel ladders are extensive (Meyer 2015). Under these conditions, opportunities for cultural burning, prescribed burning, and managed wildfires are limited to days with low to moderate fire weather, and these windows of opportunity are shrinking under climate change (Westerling et al. ²⁰¹⁶).

For the past several decades, land managers have generally followed one of two strategies to respond to wildfires in wNA forests. First, most agencies in the United States and Canada have followed a policy of aggressive fire suppression, and this approach is increasingly used in Mexico (Stephens and Fulé 2005). Under this policy, a small fraction of fires that escape suppression (<3%) are responsible for over 90% of area burned, based on a 1992 to 2015 reference period (Abatzoglou et al. ²⁰¹⁸). Second, some land managers, including those managing national parks and wilderness areas, have designated large, remote areas where most wildfires are allowed to burn under moderate fire weather and fuel conditions (Huffman et al. ²⁰²⁰). These are termed managed wildfires, with the goal of restoring more characteristic fire regimes and landscape patterns in the context of incident‐specific objectives (Table 2).

In contrast, unplanned fires that escape suppression in fire‐excluded landscapes during extreme fire weather do not generally restore forest resilience. Landscapes that are consistently managed with active fire suppression typically have a greater area burned at higher severity than those managed to restore more resilient fire regimes (Stevens et al. ²⁰¹⁷, Rodman et al. ²⁰²⁰). In fire‐excluded forest landscapes, forest surface and canopy fuels tend to be highly elevated, and despite active fire suppression, forests may eventually burn under extreme fire weather, which is becoming more frequent as the climate warms. For example, Povak et al. (2020) found fire severity during the 2013 Rim Fire was higher in the Stanislaus National Forest, much of which had not burned for >80 yr, compared to Yosemite National Park where past burn mosaics existed. High‐severity burn patches in fires that escaped suppression are larger and less complex than in fires managed with less aggressive suppression tactics (Stevens et al. ²⁰¹⁷), and seed sources for forest regeneration are more often distant, yielding sparse or non‐existent tree regeneration (Shive et al. ²⁰¹⁸, Korb et al. ²⁰¹⁹, Stevens‐Rumann and Morgan ²⁰¹⁹). In dry pine and moist mixed‐conifer forests, subsequent shrub establishment can lead to a cycle of repeated high‐severity fires that perpetuates shrub dominance and a potentially long‐term shift in alternative stable states (Collins et al. ²⁰⁰⁹, Cocking et al. ²⁰¹⁴, Coppoletta et al. ²⁰¹⁶, Coop et al. ²⁰²⁰).

Where managers allow managed wildfires to burn under prescription, burned areas are typically smaller and have greater proportions of low‐ and moderate‐severity burn patches within the fire perimeter, and high‐severity patches are typically smaller (Parks et al. ²⁰¹⁴, Stevens et al. ²⁰¹⁷). Within low‐ and moderate‐severity burn patches, fuels are reduced, and forest structures resembling more typical historical conditions emerge (Holden et al. ²⁰⁰⁷, Huffman et al. ²⁰¹⁸, Stoddard et al. ²⁰²⁰). In some forests, this includes characteristic patterns of small tree clumps and interspersed openings (Fig. 4; Kane et al. ^{2014, 2019}, Jeronimo et al. ²⁰¹⁹). In fire‐excluded forests, a first entry with managed wildfire may not meet fuels reduction and management objectives unless allowed to burn at a severity that modifies stand structure (Huffman et al. ²⁰¹⁷). Fire resilient landscapes are generally created by burning and reburning, in which prior fires modify the spread, intensity, and severity of subsequent fires (Prichard et al. ²⁰¹⁷, Walker et al. ²⁰¹⁸, Yocom et al. ²⁰¹⁹, Koontz et al. ²⁰²⁰).

Promising strategies are emerging to delineate landscapes into operational units where decisions about applying managed fire can be considered before ignitions even occur (Thompson et al. ²⁰¹⁶, Dunn et al. ²⁰¹⁷). Managed wildfires are an important management tool and they are increasingly recognized as a vital component of adaptive management. However, relying solely on managed wildfires to achieve management objectives is not possible due to a number of factors that include current restrictions on the use of managed wildfire in the WUI or near other infrastructure, limited burn windows with moderate fire weather, and the potential negative consequences of allowing fire spread into nearby fire‐excluded areas with elevated fuel loads.

Is the primary objective of fuel reduction treatments to assist in future firefighting response and containment?

In a review of fuel treatment options for interior western United States forests, Reinhardt et al. (2008) recommend that the central objective of fuel reduction treatments should not be to halt fire spread or reduce ignitions. Rather, fuel reduction treatments could be implemented to modify fire behavior and mitigate fire effects (Safford et al., ^{2012a, 2012b, b} ), thereby reinforcing the initial resilience of the treated stand by further reducing fuels, introducing greater heterogeneity, and allowing firefighters to fight fires, as needed, using direct techniques (Stevens et al. ²⁰¹⁴, Kalies and Yocom Kent ²⁰¹⁶). Under adaptive management, fuel treatments are not designed to prevent or stop fires but to moderate fire behavior when fire inevitably returns (Calkin et al. ²⁰¹⁴). However, there is a frequent misconception that fuel treatments should facilitate suppression and limit the size of wildfires (Table 1; Cochrane et al. ²⁰¹², Schoennagel et al. ²⁰¹⁷).

The reasoning behind treating fuels to facilitate fire suppression activities is circular. If fuel treatments make suppression more successful, then wildland fuels continue to accumulate, creating even more hazardous conditions for the entire landscape. Inevitably, this makes subsequent suppression more difficult, and more areas will be burned in fewer, unmanageable events with greater ecological consequences (Collins et al. ²⁰¹⁰, Calkin et al. ²⁰¹⁵). This phenomenon has been described as “the wildland fire paradox” (Arno and Brown 1991). Rather than creating conditions where wildfire is easier to suppress, fuel treatments designed within a restoration or climate adaptation strategy are engineered to allow subsequent wildfires to burn without the need of full suppression tactics and to increase opportunities for prescribed or cultural burning.

Typical fuel reduction activities near communities illustrate the long‐term consequences of using treatments with the expressed objective of suppressing future wildfires. Near communities, fuel reduction treatments are often explicitly implemented to create conditions that enhance fire suppression efficacy in both the surrounding wildland and WUI (Moghaddas and Craggs 2007). Treatment locations are selected based on criteria that involve community protection (Fleeger 2008), suppression concerns (Finney 2001), and fuel hazards (Schmidt et al. ²⁰⁰⁸), at stand and landscape scales (Chung et al. ²⁰¹³). Suppression strategies are designed to use treated areas for burnout operations, anchor points for fire lines, and safe zones for firefighters. Some of the challenges associated with this approach are that burnout operations often burn at high severity (Backer et al. ²⁰⁰⁴), and most fire line and safe zone construction involves the cutting of live and dead trees and mineral soil exposure, all of which result in conditions that can facilitate the spread of invasive species where they are present or nearby, degrade archaeological‐heritage sites, and actually reduce ecological resilience (Davies et al. ²⁰¹⁰). Further, if insufficient area is treated on a landscape, the unexpected behavior of large wildfires will overwhelm the ability of small fuel treatments to facilitate effective suppression (Agee et al. ²⁰⁰⁰, Finney et al. ²⁰⁰¹). If fuel treatments are designed such that the next wildfire can be allowed to burn with limited or no suppression, then three economic and ecological objectives might be achieved: reduced suppression costs and actions; management of future wildfires as effective fuel treatment maintenance; and favorable ecological outcomes in areas treated before wildfire.

There is little doubt that fuel reduction treatments can be effective at reducing fire severity and achieving culturally and ecologically beneficial effects, if designed and implemented correctly (Stephens et al. ²⁰⁰⁹, Fulé et al. ²⁰¹²). However, fuel treatments intended only for crown fire hazard mitigation rarely constitute effective restoration (Stephens et al. ²⁰²⁰). As the pace and scale of fuel treatments increase, emphasis on resilient forest structure and composition, long‐term reduction of surface and canopy fuels, and adaptation to climate change are critical components of treatment objectives rather than creating conditions that are more conducive to fire suppression (Hessburg et al. ²⁰¹⁹).

Do fuel treatments work under extreme fire weather?

Although extreme fire behavior including strong winds and column‐driven fire spread can overwhelm individual treatments, there is strong scientific evidence that even under extreme weather conditions, fuel reduction treatments are effective at moderating fire severity across a range of forest types and wildfire events. For example, Walker et al. (2018) studied the 2011 Las Conchas fire in New Mexico that burned under extreme weather and found that sites that were previously prescribed burned exhibited higher conifer survival (i.e., lower severity fire) compared to sites that were not treated prior to the wildfire. Similarly, Yocom Kent et al. (2015) found that moderate‐ and high‐severity effects in the Rodeo‐Chediski Fire, which burned under extreme fire weather, were reduced from 76% in untreated areas to 57% in prescribed fire, and 38% in thin and burn treatments. Likewise, Povak et al. (2020) presented evidence that some treated areas experienced lower severity fire even under the most extreme fire growth period of the 2013 Rim Fire. Past wildfires also acted as short‐term barriers to fire spread and mitigated fire severity in mixed‐conifer forests of the interior western United States (Parks et al. ^2015a , Stevens‐Rumann et al. ²⁰¹⁶). Lastly, two studies in seasonally dry mixed‐conifer forests of north‐central Washington State found that thinning followed by prescribed burning was an effective treatment for mitigating wildfire effects under extreme weather conditions (Prichard and Kennedy 2014, Prichard et al. ²⁰²⁰). Results of these observational studies are also supported by numerous modelling studies indicating that fuel treatments reduce fire intensity and effects in dry conifer forests under dry fuels and high wind speeds (Stephens and Moghaddas 2005, Ager et al. ²⁰⁰⁷, Vaillant et al. ²⁰⁰⁹, Johnson et al. ²⁰¹¹).

In forests characterized by moderate‐ and high‐severity fire regimes, a limited number of studies suggest that fuel reduction treatments are ineffective at reducing fire behavior and effects, particularly under extreme weather conditions (e.g., Graham ²⁰⁰³, Martinson et al. ²⁰⁰³, Schoennagel et al. ²⁰⁰⁴). The rationale is that fires burning within moist and cold forest patches are generally controlled by climate (i.e., a warmer and drier than average year) and not controlled by fuel within patches (Turner and Romme 1994, Bessie and Johnson ¹⁹⁹⁵). However, at larger spatial scales, there is strong evidence that patchwork burn mosaics resulting from reburns reduce landscape contagion, and consequently, spread and severity of wildfires, even under extreme fire weather (Stine et al. ²⁰¹⁴, Parks et al. ^2015b , Hessburg et al. ²⁰¹⁶, Spies et al. ²⁰¹⁸).

Dependent on the forest type and environmental setting, some fuel treatments are more effective at reducing adverse fire effects than others, and this can also contribute to confusion as to whether or not treatments are effective under extreme fire weather. Several studies highlight that the most effective fuel treatments include coupled thinning and burning (Kalies and Yocom Kent ²⁰¹⁶), and emphasize the importance of retaining large, fire‐resistant trees in dry mixed conifer forests (DellaSala et al. ²⁰⁰⁴, Agee and Skinner ²⁰⁰⁵, Stephens et al. ²⁰⁰⁹). Furthermore, other studies showed that fire severity decreased as wildfires progress further into areas with more treated area (Arkle et al. ²⁰¹², Kennedy and Johnson ²⁰¹⁴), strongly suggesting that small fuel treatments or those with large perimeter‐to‐edge ratios are less effective than larger treatments under extreme fire weather conditions (Kennedy et al. ²⁰¹⁹).

Finally, fuel treatments generally are designed to mitigate wildfire intensity and effects but they are not necessarily intended to impede fire spread or reduce fire size (Reinhardt et al. ²⁰⁰⁸). Consequently, when fires burn large areas under extreme fire weather some may conclude that burned‐over fuel treatments were ineffective (e.g., Schoennagel et al. ²⁰¹⁷). However, the occurrence of large fires does not necessarily suggest that existing fuel treatments were unsuccessful. Large fires have always been a part of fire‐prone forests, and within large fire events fuel treatments can allow fires to continue burning but mitigate fire severity and enhance the heterogeneity of fire effects.

Is the scale of the problem too great? Can we ever catch up?

Recent meta‐analyses of fuel treatment effectiveness demonstrate that at landscape and regional scales, fuel treatments account for only a small fraction (˜1%) of the area burned by wildfires (e.g., Barnett et al. ^2016a , Kolden ^{2019, 2019}). Therefore, there is some concern that treatments are ineffective because under current prescription levels, wildfires may not actually encounter treated areas during the duration of their potential effectiveness (Odion and Hanson 2006, Rhodes and Baker ²⁰⁰⁸). While this is factually accurate at the current pace and scale of treatment in wNA, the question is not whether every wildfire can be impacted by fuels treatments, but whether treatments can be strategically used to multiply their benefits and promote greater opportunities for applying wildland fire across landscapes. The scientific evidence that fuel reduction treatments can mitigate fire behavior and effects strongly supports a conclusion that expanding treated areas, including the use of forest thinning, prescribed burning, cultural burning, and managed wildfires, will lead to greater landscape resilience to future wildfires.

Ongoing warming and drying are linked to increasing large fire occurrence, contributing to large increases in area burned (Abatzoglou and Williams 2016) and area burned as high severity (Parks and Abatzoglou 2020) in wNA in recent decades. Given projected increases in warming due to climate change, burn probability is increasing in many wNA forests (Littell et al. ²⁰¹⁸, Hurteau et al. ²⁰¹⁹) along with increasing likelihood that future wildfires will impact a larger proportion of landscapes. In this light, the current pace and scale of fuels treatments is insufficient to address the scale of fire exclusion. Furthermore, treated areas require ongoing maintenance to retain efficacy (Krofcheck et al. ²⁰¹⁷, Vaillant and Reinhardt ²⁰¹⁷), making it difficult to expand treated areas across a landscape without significant additional financial and personnel investments (North et al. ^2015a ). Thus, the scope, scale, and urgency of adapting wNA forests to climate change and future wildfires is immense.

Given the complexity of forest ecosystems, the economic and personnel investment required, and the policy and management constraints, there is no single management tool that is adequate to increase the resilience of wNA landscapes to future wildfires. Coupled thinning and burning treatments will be especially helpful in dry pine, oak woodlands, and dry mixed conifer forests, while restoration of more characteristic forest successional and nonforest patchworks using managed moderate and high severity wildfires will be key in cold forests. Forest managers in western Australia have reduced the frequency of large and severe wildfires, but only after building extensive landscape networks of strategic treatments (i.e., spatially linked naturally occurring and treated areas of reduced fuels prior to the outbreak of wildfires) and by conducting frequent prescribed burning under moderate fire weather and including Indigenous fire use over large areas (Boer et al. ²⁰⁰⁹, Sneeuwjagt et al. ²⁰¹³). Similar approaches are being used in U.S. national forest, wilderness, and park areas to allow for more area of managed wildfires (Table 2). Given limitations on where mechanical thinning, prescribed and cultural burning, and managed wildfire are practical or allowed, combining these tools over broad areas can markedly expand treatment extent and reduce impact of large wildfires.

Fire hazard, burn probability, and fire ecology vary widely across wNA forest landscapes. Prior knowledge of cultural burning practices, ignition and weather patterns, vegetation and fuel distributions, and topography all provide critical information for prioritizing fuel treatments in areas with the highest risk of burning (Ager et al. ^{2010, 2016}). Near population centers, humans are often responsible for the majority of wildfire ignitions, and they provide ignition sources in highly predictable areas and seasons of the year, when natural ignitions are rare (Balch et al. ²⁰¹⁷, Keeley and Syphard ²⁰¹⁸). Ignition pattern and frequency interact with fuels, weather, and topography to influence fire occurrence, leading to heterogeneous burn probabilities across a landscape (Ager et al. ²⁰¹², Povak et al. ²⁰¹⁸). Using prior knowledge of human and lighting‐caused fire starts coupled with knowledge of the probability of fire spread and likely severity, managers can identify the areas of any landscape where uncharacteristic or impactful fires will likely occur (Parisien and Moritz 2009, Parisien et al. ²⁰¹²), and decrease the proportion of the landscape that requires treatment.

There are a number of available tools and approaches to identify areas that would benefit from strategically placed fuel treatments. In general, fuel treatments are not implemented at random, and for good reason (Finney et al. ²⁰⁰⁷). A comparison of random vs. strategically placed treatments showed that a significant reduction in area could be achieved with strategic placement (Ager et al. ^{2013, 2016}), where that opportunity exists. Quantifying the probability of high‐severity wildfire across a given landscape and focusing thinning treatments on high‐probability areas can decrease the required treatment area by >50% (Krofcheck et al. ²⁰¹⁹). However, the success of these strategies depends on maintaining the treatments and reintroducing fire to a larger portion of the landscape (Agee and Skinner 2005, Barros et al. ²⁰¹⁸). Where reserved areas are abundant or widely distributed, opportunities for spatially optimizing fuel treatments are limited, and considerably more treated area may be required outside of reserves (Finney et al. ²⁰⁰7).

In summary, justifying inaction based on the scale of the problem is too large is highly circular. Evidence supports increasing the pace of treatments to significantly reduce the area impacted by uncharacteristic wildfire, even under a changing climate (Liang et al. ²⁰¹⁸). For example, managers can expand areas where burn prescriptions are applied to reduce fuels and increase forest heterogeneity (Safford et al. ^{2012a, 2012b}, Striplin et al. ²⁰²⁰). The efficacy of these was historically demonstrated by Indigenous burning practices that amplified natural lightning ignitions in many seasonally dry forests, thereby modifying active fire regimes and fire effects, and diversifying the seasonality and frequency of fires (Crawford et al. ²⁰¹⁵, Trauernicht et al. ²⁰¹⁵, Taylor et al. ²⁰¹⁶). Managed wildfires can also increase forest and fuel heterogeneity, constraining subsequent fire size and severity (Collins et al. ²⁰⁰⁹, Parks et al. ^2015b , Barros et al. ²⁰¹⁸). When used in conjunction with mechanical treatments and prescribed or cultural burning, managed wildfire presents an opportunity to increase the effectiveness of treatments across large landscapes (North et al. ²⁰¹²).

Will planting more trees in wNA forests help to mitigate climate change?

Tree plantations have long been a debated aspect of forest management, and more recently, climate change mitigation (Alig 1997, Chmura et al. ²⁰¹¹). Planting after harvest to increase forest productivity were the central justifications for past clearcut logging, even as a growing body of science demonstrated that plantations (1) did not provide the needed ecological structures or functional diversity of old‐growth forests, (2) were not necessarily more productive than mature forests (Franklin et al. ²⁰⁰²), and (3) without surface fuel treatment, could be conducive to high‐severity wildfires (Thompson et al. ²⁰⁰⁷). Similarly, planting seedlings after post‐fire salvage logging is sometimes used to expedite tree regeneration following high‐severity fire. Without strategic management, post‐fire plantations may be overstocked, dominated by a single species (North et al. ²⁰¹⁹), lack tree clumping and canopy gaps, and pose significant wildfire hazard (Kobziar et al. ²⁰⁰⁹), particularly without post‐harvest slash reduction (Donato et al. ²⁰⁰⁹).

A recent proposal to combat climate change includes planting a trillion trees globally, including substantial reforestation in the western United States (Bastin et al. ²⁰¹⁹). The study suggested that these additional trees would sequester sufficient atmospheric carbon to curb climate change. Baseline assumptions and findings from this study have been contested by scientists (Veldman et al. ²⁰¹⁹, Holl and Brancalion ²⁰²⁰) as the study failed to account for forest interactions with climate, drought, and wildfire dynamics. In addition to future disturbance resilience, numerous other barriers currently impede large‐scale reforestation efforts (Fargione et al. ²⁰²¹).

Across wNA, most of the forest carbon is captured in moist temperate forests with high precipitation levels and net primary productivity, including the coastal ranges along the Pacific Coast, western Cascade and western Sierra Nevada Mountain Ranges (Hudiburg et al. ²⁰⁰⁹). These forests possess complex, heterogeneous structures, some of which developed with infrequent wildfires. Others, including those in southwestern Oregon and northern California, were also influenced by a long legacy of Indigenous burning (Anderson 2013, Merschel et al. ²⁰¹⁴). Because most of the standing biomass in high productivity wNA forests occurs in live trees, when these forests burn, relatively low levels of carbon are initially emitted, with most of the biomass retained either in standing trees and snags or to newly downed heavy fuels that slowly release carbon to the atmosphere through decomposition, unless they subsequently burn in a reburn fire event (Stenzel et al. ²⁰¹⁹, Lutz et al. ²⁰²⁰). By contrast, even‐aged stands, both naturally occurring (e.g., lodgepole pine forests) and in young plantations, are relatively homogeneous in structure, and with elevated surface fuels, can facilitate high‐intensity, severe fire (Bowman et al. ²⁰¹⁹). Climate change‐induced shortening of fire return intervals may ultimately convert some of these live carbon pools from sinks to sources (Turner et al. ²⁰¹⁹, Foster et al. ²⁰²⁰).

In fire‐adapted dry mixed conifer forests, dense tree plantations are highly susceptible to future wildfires and drought. However, a promising approach to retaining and sequestering carbon in dry, fire‐prone forests is to retain existing large‐diameter trees and restore characteristic low‐severity fire to maintain low‐severity fire to maintain resilient forest structure and composition (Hurteau and North 2009). It is still debatable whether prescribed burning and removal of small diameter trees and ladder fuels will actually increase or decrease aboveground carbon stores (Campbell et al. ²⁰¹², Restaino and Peterson ²⁰¹³) and is likely site dependent, but there is broad scientific agreement that these management actions are key to increasing forest ecological resilience, which ultimately stabilizes forest carbon stocks (Hurteau et al. ²⁰¹⁹, Krofcheck et al. ²⁰¹⁹, Westlind and Kerns ²⁰²¹). Managed landscape mosaics will be particularly critical to maintaining legacy old‐growth forests and minimizing sink‐to‐source conversions due to fire and other disturbances (Barbero et al. ²⁰¹⁵, Liang et al. ²⁰¹⁷). Finally, governmental cap‐and‐trade and carbon taxation programs must accurately account for the complex role fire plays in carbon cycle feedbacks and carbon maintenance, rather than simply characterizing fire as a net carbon loss (Hurteau et al. ²⁰⁰⁸, North et al. ²⁰⁰⁹).

Across wNA forests, tree planting can serve as an important tool to nudge the trajectory of post‐fire landscapes towards more climate‐adapted tree species or genotypes, particularly in areas where seed source is limited (North et al. ²⁰¹⁹). However, traditional high density plantations will often predispose forests to high‐severity fire where pre‐commercial thinning and associated fuel treatments are not implemented, which is increasingly the case (McCarley et al. ²⁰¹⁷). Alternatives to traditional plantations are emerging that are designed to promote resilience to future fire and drought from the beginning of the planting process. These include planting drought‐conditioned seedlings reared from lower‐elevation seed stock, planting discontinuous “founder stands” or “nucleation islands” of trees into portions of stand‐replacing patches far from tree refugia, and planning for the reintroduction of fire into younger planted stands as they develop (Peterson et al. ²⁰⁰⁷, Landis et al. ²⁰¹¹).

Is post‐fire management needed or even ecologically justified?

Many contemporary wildfires exhibit a range of post‐fire effects (Thode et al. ²⁰¹¹); variable sized patches of stand‐replacing or partial stand replacing fire are embedded within a matrix of live forest (Stevens et al. ²⁰¹⁷). Among large fires, these patches of stand‐replacing fire may themselves contain isolated and variably sized patches of live trees often referred to as fire refugia (Meddens et al. ²⁰¹⁸, Krawchuk et al. ²⁰²⁰). Thus, the post‐fire landscape can be viewed as a complex patchwork of interconnected surviving forest, the product of low and moderate severity fires, high‐severity patches, and isolated refugia (Coop et al. ²⁰¹⁹). However, these post‐fire landscapes are not necessarily on resilient trajectories. Fire refugia may be in uncharacteristic locations, and active forest and fuels management are often required after the fire to promote future forest resilience to disturbance and climate change and to protect valued cultural resources.

Patches of low‐ and moderate‐severity fire generally have short‐term resistance to future fire due to the reduction of surface fuels from the first burn (Prichard et al. ²⁰¹⁷). Compared to low‐severity fire, moderate‐severity fire events can create a residual stand structure that more closely approximates historical conditions (Collins et al. ²⁰¹¹, Huffman et al. ²⁰¹⁷). However, moderate‐severity fires that burn through previously dense forest also leave considerable standing and down wood, which can lead to elevated fuel loads and high‐severity fire in subsequent reburns (Collins et al. ²⁰¹⁸). Thus, post‐fire fuel reduction of the trees that encroached during the period of fire exclusion can be warranted to improve the fire resilience of residual forests, including fire refugia.

Smaller refugial patches within larger burned patches are increasingly recognized as having significant cultural and ecological value by preserving biological and cultural legacies that can contribute to forest succession via seed dispersal (Johnstone et al. ²⁰¹⁶, Meddens et al. ²⁰¹⁸). Small refugia in particular make disproportionate contributions to reforestation potential within larger patches of stand‐replacing fire (Shive et al. ²⁰¹⁸, Coop et al. ²⁰¹⁹). However, isolated tree refugia can have a significant standing and down fuel component around their edges due to adjacent high‐severity burn effects (Lydersen et al. ^2019a ). Given their outsized importance as biological legacies, surface fuel reduction to “harden” the edges of refugia may be critical to their future resilience and prioritize refugia retention during wildland firefighting operations (Meddens et al. ²⁰¹⁸).

Large patches of stand‐replacing fire are an increasing focus of research (Coop et al. ²⁰²⁰). Independent of subsequent fire dynamics, regeneration is challenged by seed dispersal limitations and a warming climate (Stevens‐Rumann and Morgan ²⁰¹⁹). Fuel conditions in large patches of stand‐replacing fire are usually dominated by coarse wood, regenerating shrubs, and hardwoods, increasing the risk of subsequent high‐severity, and occurrence of long‐duration re‐burns (Coppoletta et al. ²⁰¹⁶, Prichard et al. ²⁰¹⁷). Collectively, these conditions pose a substantial management challenge if the objective is to restore at least a portion of large burn patches to conifer forest, as this is unlikely over decades to centuries without management intervention (Coop et al. ²⁰²⁰).

Fuels management and regeneration dynamics in stand‐replacing patches are closely related. In high‐severity patches, management to reduce coarse wood accumulations and flammable shrubs may promote post‐fire tree regeneration and mitigate future fire severity (Peterson et al. ²⁰¹⁵, Lydersen et al. ^2019b ). In planted forests, coarse wood presents a different challenge, as downed logs facilitate seedling survival through shading and moisture retention (Castro et al. ²⁰¹¹) but pose a risk to seedlings if they burn (Peterson et al. ²⁰¹⁵). Understanding the range and variability of historical reburning would provide essential guidance of restoration targets to improve the post‐fire resilience of regenerating landscapes.

Strategic tree planting can be used to encourage the re‐establishment of some post‐fire landscapes and for climate change adaptation, particularly where conditions are not favorable to natural regeneration (see previous question). Post‐fire mechanical thinning (e.g., salvage logging) is often driven by economic and safety considerations but may have some ecological benefits in terms of reduced future surface fuel loads and fire hazard 10–20 yr post‐fire (Peterson et al. ²⁰¹⁵). Future research in this area is warranted to investigate the impacts of variable density harvests and how potential ecological tradeoffs vary over time (e.g., Ritchie et al. ²⁰¹³).