Shortening fire return interval predisposes west‐central Canadian boreal peatlands to more rapid vegetation growth and transition to forest cover

Climate change in northern latitudes is increasing the vulnerability of peatlands and the riparian transition zones between peatlands and upland forests (referred to as ecotones) to greater frequency of wildland fires. We examined early post‐fire vegetation regeneration following the 2011 Utikuma complex fire (central Alberta, Canada). This study examined 779 peatlands and adjacent ecotones, covering an area of ~182 km2. Based on the known regional fire history, peatlands that burned in 2011 were stratified into either long return interval (LRI) fire regimes of >80 years (i.e., no recorded prior fire history) or short fire return interval (SRI) of 55 years (i.e., within the boundary of a documented severe fire in 1956). Data from six multitemporal airborne lidar surveys were used to quantify trajectories of vegetation change for 8 years prior to and 8 years following the 2011 fire. To date, no studies have quantified the impacts of post‐fire regeneration following short versus long return interval fires across this broad range of peatlands with variable environmental and post‐fire successional trajectories. We found that SRI peatlands demonstrated more rapid vascular and shrub growth rates, especially in peatland centers, than LRI peatlands. Bogs and fens burned in 1956, and with little vascular vegetation (classified as “open peatlands”) prior to the 2011 fire, experienced the greatest changes. These peatlands tended to transition to vascular/shrub forms following the SRI fire, while open LRI peatlands were not significantly different from pre‐fire conditions. The results of this study suggest the emergence of a positive feedback, where areas experiencing SRI fires in southern boreal peatlands are expected to transition to forested vegetation forms. Along fen edges and within bog centers, SRI fires are expected to reduce local peatland groundwater moisture‐holding capacity and promote favorable conditions for increased fire frequency and severity in the future.


| INTRODUC TI ON
The Canadian boreal biome spans an area of ~5.5 million km 2 , more than half the area of Europe.The biome is characterized by a heterogeneous mix of upland forests, lakes, and wetlands with significant global importance for modulating the impacts of climate change (Bonan et al., 1992;Helbig, Waddington, Alekseychik, Amiro, Aurela, Barr, Black, Blanken, et al., 2020;Helbig, Waddington, Alekseychik, Amiro, Aurela, Barr, Black, Carey, et al., 2020), driving weather (Davis et al., 1997;Pennypacker & Wood, 2023), and provision of natural resources (Brandt et al., 2013).Peatlands, defined as wetlands with thick layers of dead, organic soils and mosses, comprise ~22% of the boreal biome (Thompson et al., 2016) and are distributed between upland forests and in localized topographically low-lying areas.The effects of the successional history of peatlands, which typically develop from the accumulation of organic material within shallow ponds via terrestrialization or in upland areas via paludification (Bauer et al., 2003), combined with cool, moist climate (Vitt, 2006), and growth of mosses, shrubs, and trees over time (Vitt, 2006) contribute to their ecological and climatological importance.As such, peatlands are considered climate change (Stralberg et al., 2020) and fire (Thompson, Parisien, et al., 2017) refugia largely due to the moisture retention of mosses and typical near-surface water table (Waddington et al., 2015).These characteristics limit the spread of fire (Turetsky et al., 2004) and act as a long-term atmospheric carbon sink due to the reduced decomposition of peat as a result of cold and anoxic conditions (Wang et al., 2015).The amount of carbon contained within peatlands is extensive, representing the Earth's largest terrestrial C store (Immirzi et al., 1992;Roulet, 2000;Turetsky et al., 2015).
The potential for fire spread is further increased during prolonged warm, dry (severe fire weather) conditions, as determined by the Canadian Fire Weather Index (Thompson et al., 2019;Waddington et al., 2012).Upland forests are even more susceptible to rapid fire spread during these periods due to evapotranspiration, insect disturbances, and mortality (Frelich & Reich, 2010;Michaelian et al., 2011) with variations depending on species and local hydroecology (e.g., Pappas et al., 2018).This, combined with a lowering of the water table (Kettridge Turetsky et al., 2015;Wilkinson, Tekatch, et al., 2020;Wilkinson, Verkaik, et al., 2020), enhances fire spread and fire intensity (energy; Walker et al., 2020;Whitman et al., 2022) and may alter fire regimes (Whitman et al., 2019).
Enhanced potential for fire spread into peatlands from surrounding upland forests via surface moisture reduction and vegetation changes along the ecotonal boundaries may alter current and future fire regimes within the Canadian boreal biome, with unknown impacts to boreal peatlands in the years following fire.Such changes are essential to quantify because these may indicate resilience or a shift to a new ecosystem state (e.g., Gibson et al., 2018).Boreal peatland fire return intervals (years between fires) typically range from 100 to 130 years (Turetsky et al., 2004;Wieder et al., 2009); however, over the last two decades, peatland fire return intervals have shortened, with some fires occurring within 20-100 years of a previous fire.While few studies have examined boreal peatland fire return intervals, Sutheimer et al. (2021) found that historic return intervals over the last several centuries for hemi-boreal peatlands south of the Canadian/US Great Lakes could occur as frequently as 7 and 31 years between low-intensity fires.Proximal drivers may also influence the propensity for shortening return intervals and postfire effects (Turetsky et al., 2015).For example, Kettridge Turetsky et al. (2015), Whitman et al. (2019), andGrzesik et al. (2022) hypothesized a shift to more deciduous species, while Bourgeau-Chavez et al. (2020) show that bogs are more susceptible to greater burn severity.However, fens transition to greater susceptibility for burning when fire weather is severe.
Bogs receive moisture from precipitation, increasing bog hydrology coupling to fire weather conditions (Glaser et al., 1997).In peatland/upland ecotones, fluctuating hydraulic gradients (Bhatti et al., 2006;Goud et al., 2018) can enhance rapid temporal and spatial changes in vegetation species and structures associated with moisture retention at the peatland edge (Jones et al., 2022;Yarrow & Marín, 2007).Therefore, peatland shape and proximity to upland conifer versus deciduous forests are hypothesized to enhance the propagation of some species, including Salix spp.(Dimitrov et al., 2014;Mayner et al., 2018) into the ecotonal boundary during dry phases (e.g., Chasmer & Hopkinson, 2017).Fens, which are at an earlier successional phase to bogs within the southern and central Canadian boreal biome, are typically connected to groundwater discharge resulting in moderated water table fluctuations which may enhance resilience to fire in moderate fire weather scenarios (Bourgeau-Chavez et al., 2020).Therefore, the ratio of the peatland perimeter to area, where larger peatland areas relative to longer perimeter (which can enhance ecotonal edge effects) and adjacent species may alter the potential for a high-intensity fire, changes in fire regime, and post-fire effects (Mayner et al., 2018).
In this study, we consider the impacts of short versus typical fire return intervals (herein short return interval [SRI] fire and typical or long return interval [LRI]) on post-fire peatland and ecotone recovery within >700 peatlands using multi-temporal airborne lidar data in the Boreal Plains, Alberta Canada.We hypothesize that the recovery of peatlands with shortening fire return intervals will be directly related to hydrological functions as these vary with surficial geomorphological groups, peatland shape, and the potential for vegetation encroachment from proximal land covers (Ferone & Devito, 2004;Hokanson et al., 2018).To address this hypothesis, three questions This is the first study to consider the broad spatial and temporal range of factors influencing post-fire ecotonal and peatland regeneration following SRI versus LRI fires across hundreds of peatlands in western Canada.This provides a more complete assessment of potential future peatland and landscape evolution associated with changing feedback between climate, wildfire, and peatland eco-hydrology.
Wildland fires burned the area in 1956 and 2011 (Alberta Wildfire, 2018), with overlapping burn scars covering an area of about 12 km 2 .Peatlands within this overlapping area, therefore, had a firereturn interval of 55 years (SRI fire), where the two fires occurred in nearly half of the time frame typical for this area (Benscoter, 2006;Benscoter & Vitt, 2008;Tarnocai et al., 2011;Turetsky et al., 2002).
The area burned in 2011 is considered LRI and is more typical of the fire regime of this area, which has been monitored since the 1940s.

| Lidar data and processing
Six airborne lidar surveys were collected over 17 years (Table 1).
Lidar data were collected pre-fire in the summer of 2002 (Hopkinson et al., 2005) and 2008 (Chasmer et al., 2016;Montgomery et al., 2019), and post-fire in July to early August 2011, 2016, 2018, and 2019(methods described in Hopkinson et al., 2013;data available via Hopkinson et al., 2024).A postprocessed kinematic survey was collected using a Global Navigation Satellite System (GNSS) receiver along Highways 88 and 750 (Figure 2) to calibrate and align all lidar point clouds to a common elevation control.
Post-processing of lidar data included classification of ground and non-ground returns, quality control, and removal of isolated points using TerraScan (TerraSolid Inc., Finland).While lidar data provide an opportunity to characterize the 3D structure of vegetation and underlying ground surface elevation, we focused on canopy height models (CHM) to determine canopy height changes in the years following fire.This reduces the potential for uncertainty associated with the frequency and distribution of laser returns due to survey configuration (Hopkinson et al., 2006(Hopkinson et al., , 2016;;Naesset, 2009).
A digital elevation model (DEM) was created from ground-classified returns at 1 m pixel resolution and was used to normalize vegetation height derived from a digital surface model (DSM) relative to the DEM.The DSM includes canopy and ground elevation, calculated using the maximum height of lidar returns within 1 m 2 cells from the canopy.These were then resampled to 5 m 2 by taking the average maximum height (Naesset, 2004;Wulder et al., 2012).CHMs were produced for each year of lidar data collection to determine changes in vegetation height.To remove the influence of remaining burned/ standing tree stems from the CHM, these were segmented within each post-fire CHM dataset by removing cells taller than the regenerating vegetation.

| Peatland and ecotone classification and identification
Peatlands were identified using an enhanced wetland classification map modified from Chasmer et al. (2016)  Ecotone boundaries are challenging to identify and map using optical remotely sensed data due to vegetation species and structural similarities between forested uplands and peatlands.
To address this problem, ecotones were classified using a systematic "criteria-based" spatial analysis approach based on the topographic position of the elevation (using the DEM) and canopy height characteristics.Peatland maximum, mean elevation, and vegetation height were determined based on all cells for each peatland using the 2008 (pre-fire) lidar dataset.To determine the area of ecotones, 5 m incremental rings (or buffers) were created, extending from each peatland edge as elevation increased, transitioning from peatlands to uplands.The maximum and mean elevation and canopy height were calculated for each 5 m buffer ring to a maximum distance of 100 m (to reduce the potential to overlap with other peatlands).These were then differenced from those calculated for the peatland and that of the next closest buffer ring to the peatland (Figure 3).Based on this iterative process, a decision criterion was used to identify: (i) the maximum difference between elevation/canopy height from that found within the peatland, and (ii) the distance at which the differences between zones were minimized, indicating that elevation had become "flatter," as is expected when transitioning from ecotone to upland forest boundary (Figure 3; R Core Team, 2021).Based on this method, an adaptive transition distance was defined for each peatland on a per (5 m × 5 m) pixel basis (ecotonal classification available via Chasmer et al., 2024).This method assumes that the greatest changes occur along the peatland margins where fluctuating water tables and the combination of organic and mineral soil textures create a dynamic nutrient-rich environment (Hokanson et al., 2016;Jones et al., 2022;Lukenbach et al., 2015).

| Quantifying vegetation change using time-series lidar data
In this study, mapped vegetation height changes across the peatland and return interval strata were then determined between years by grouping and subtracting the vegetation CHM changes into the early post-fire period (from 2011 to 2016), the latter period (from 2016 to 2019), and the entire early post-fire period (from 2011 to 2019), each divided by the number of years to estimate a rate of growth per year (Hopkinson et al., 2006(Hopkinson et al., , 2016;;Naesset, 2009).To determine the area proportion of vegetation change per peatland, pixel difference >0.15 m was included (as 0.15 m is greater than the vertical error of the point cloud elevation data; Andersen et al., 2006;Falkowski et al., 2008;Hodgson & Bresnahan, 2004).Changes and areas of vascular vegetation, especially in areas of expansion and reduction of shrub and tree heights within strata, were also examined by considering entire peatland and ecotonal areas.Here, the areas of vegetation >0.5 m in height between 2008 (pre-fire) and 2019 (8 years post-fire) were compared within each SRI versus LRI burned areas, bogs, and fens, and within ecotones.To determine the distribution of vegetation height characteristics (R Core Team, 2021), ~6000 points were randomly distributed (with a minimum distance of 5 m) across the study area within each stratum.These quantify the range of variability of vegetation heights for each peatland strata: SRI and LRI bog, fen areas, and their adjacent ecotones.

| Proximity analysis of environmental gradients compared with post-fire vegetation change
Post-fire vegetation regeneration may be influenced by the ecohydrological characteristics of peatlands, broadly estimated based on peatland class, shape and morphology, and adjacent vegetation species, which can influence succession and life history.The results are introduced by describing variability in vegetation growth associated with underlying surficial geomorphology (fine-textured glaciolacustrine plains and hummocky moraine; Devito et al., 2016;Paulen et al., 2006).The influences of environmental gradients were then examined by regressing vegetation change within SRI and LRI burned areas to peatland shape and proximity to upland forest vegetation types.To determine peatland shape, the perimeter to area (P:A) ratio was determined from the area and perimeter of each pre-fire classified peatland, where lower P:A ratios indicate compact/round shapes and higher ratios indicate more complex shapes.Regression    with approximately 46% of peatlands in fine-hummocky moraine and 47% within the fine-clay plains, with the remaining area as coarse glacial-fluvial deposits.Peatlands within SRI burn areas were primarily found in fine-hummocky moraine (77%), with 23% in fine plains.Despite differences in geomorphological characteristics, there were no significant differences in vegetation heights in areas burned by SRI or LRI fires within these two geomorphological conditions (SRI p = .06;LRI p = .08;>.05 significance).This indicates that in the early stage of post-fire regeneration, the geomorphological landform did not greatly impact vegetation changes in this region.

| Statistical analysis
Vegetation heights in adjacent ecotonal areas demonstrated similar patterns, with even lower differences between fine hummocky moraines and fine plains (SRI p = .11;LRI p = .23).
Assuming similarity between geomorphological groups, annual rates of vascular vegetation growth were greater in ecotones than in peatlands.These changes were also greater following SRI fire compared with LRI fire (Figure 4). in peatlands, though peatland post-fire vegetation growth rates were not greatly different following SRI and LRI fire (p = .30).The average early post-fire growth rate was 0.74 m year −1 (SD = 0.2 m) in peatlands following SRI fire and 0.70 m year −1 (SD = 0.28 m).
Peatland edges also experienced the greatest growth of postfire vegetation, with peatland centers often experiencing lower rates of growth of shrubby vegetation (with the exception of R. groenlandicum and mosses).Dominant species regeneration along peatland margins included R. groenlandicum and some Salix (spp.)observed at the site in 2016 (Schneider et al., 2016).
Between 2016 and 2019, vegetation growth rates reduced substantially within peatlands, especially within ecotones following SRI and LRI fires (Figure 4).In some ecotones, height reductions occurred, which may have been associated with wet summers in 2018 and 2019 and possible mortality of some shrub species, which has been observed in peatlands and ecotones in other studies (e.g., Schneider et al., 2016).During this period, the average annual rate of growth reduced to 0.27 and 0.32 m year −1 within peatlands and ecotones, respectively, following SRI fire where there was a greater decline in growth rate in peatlands and ecotones compared with LRI fire (reduction of 8%, on average).Despite slowing of the growth rate in the latter early post-fire period, vascular vegetation, including P. tremuloides, Betula papyrifera, Picea (spp.), and Alnus (spp.), proliferated within ecotones with some expansion into peatlands, especially following SRI fire (e.g. Figure 1).

| Pre-fire trends and post-fire encroachment of vascular vegetation into peatlands
Prior to the 2011 Utikuma Complex fire, forested areas burned in 1956 (SRI), and those that were not burned in the recent past (LRI) were characterized as mid-to late-stage growth.Vascular vegetation heights within peatlands (Figure 5) burned in 1956, and those that had not been burned in the recent (monitored) past were similar in 2002 and 2008 prior to the 2011 fire.
Over the 8 years following fire, total annual vascular vegetation growth rates were significantly greater in bogs compared with fens and greater following SRI fire.Here, on average, bogs following SRI fire experienced 11% greater height growth than those following LRI fire (p = .012),including the reduced growth rates in 2018 and 2019 (Figure 5).Differences in post-fire vegetation growth following SRI versus.LRI fire in fens were not as large (4% on average greater following SRI fire, p = .023).Therefore, bogs were more susceptible to rapid post-fire growth of shrub and tree species than fens, especially with shortening fire return intervals.
Growth rates of vegetation within peatlands following SRI fire may have also been due to encroachment of vascular vegetation from ecotones.Average vascular vegetation heights (>0.5 m) were about 8.97 m adjacent to bogs and ~1.5 m taller (on average) than fen ecotones by 2018 following SRI fire (Figure 5).These were also taller (on average) than ecotonal vegetation growth following LRI fire.
However, while bogs experienced the greatest rate of vegetation growth, especially following SRI fire, the proportional area coverage of vegetation changes within peatlands at 8 years post-fire indicates changing vegetation succession (Figure 6).
Within the study area, bogs cover a much smaller area than fens and have larger ecotonal boundaries relative to fens.Here, we observed a shift to reduced shrub/tree cover within previously (prefire) treed/shrub bogs following SRI fire (from 44% of bog area, pre-fire to 24% shrub/tree cover, post-fire), while pre-fire open bogs with few shrubs/trees increased in shrub/tree cover from 4% (prefire) to 25% (post-fire; Figure 6a).These also had significantly greater vegetation heights (average = 0.30 m year −1 , p = .003)than open bogs burned in LRI fire.Bogs following LRI fire were largely resilient and did not experience significant shifts from pre-fire to post-fire shrub/tree vegetation cover.Similarly, there was a shift to post-fire encroachment of trees and shrubs into ecotones that were largely shrub/tree-free prior to fire.This was also found adjacent to fens, with greater encroachment of trees/shrubs into open post-fire fen ecotones where the average growth rate of trees and shrubs was 0.28 m year −1 and significantly greater (p = .0004)than shrub/treed fens following fire.Post-fire areas of regenerating shrubs and trees within fens were similar (within 3%) following LRI fire in previously shrub/treed and open fens (Figure 6b).However, a proportional increase in the area of trees/shrubs from 37% to 45% following SRI fire in fens illustrated differences in post-fire regeneration within fens following SRI fire.

| Influence of shape and proximal species on post-fire vegetation change
Peatland shape and area may influence the potential for postfire vegetation encroachment and rates of growth, especially in peatlands that are small or have greater edge complexity determined by longer perimeters.These are prone to greater vegetation changes associated with ecotonal and proximal forest boundaries (e.g., observed in temperate forests in Morreale et al., 2021) and variable wetting and drying conditions within the peatland margin (lag) (Hokanson et al., 2018).Here, the range of peatland perimeter to area ratio (shape complexity determined per peatland) was greater in LRI-burned peatlands (P:A > 0 to ~4) than in SRI (P:A > 0 to <2).Peatland shape complexity was weakly positively correlated with annual average percent vegetation growth following LRI fire (R 2 = .35,p < .0001)(Figure 7a).In LRI-burned peatlands, vegetation height increased to a threshold of about 60% proportional cover of taller (vascular, shrub, seedling) vegetation in peatlands with increasingly complex shapes (Figure 7a, R 2 = .45,p < .0001).Following SRI fire, there was no relationship between the complexity of peatland shapes (higher perimeter to area ratio) and post-fire percent annual vegetation growth (R 2 = .04,p = .73,

Figure 7b).
Adjacent deciduous and mixed wood forests (especially those producing aspen suckers) could also result in P. tremuloides expansion into peatlands.However, peatlands adjacent to other peatlands and not surrounded by high proportional areas of upland forests tended to have the greatest area of post-fire vascular vegetation growth (Figure 8).Counter to expectations, vascular vegetation growth rates were lower in peatlands burned following both LRI and SRI fires that were adjacent to larger areas of mixed and deciduous upland forests, r 2 = .48,.35,respectively (p < .0001, Figure 8a).
Peatlands adjacent to conifer forests experienced positive changes in vegetation heights.Post-SRI fire vascular vegetation growth increases in peatlands surrounded by more than 50% upland conifer forests (Figure 8b).

| Fire regime impacts early post-fire vegetation succession within peatlands
This study aimed to assess whether shortening fire return intervals impacts vegetation growth trajectories relative to LRI fires, which  sources of water for surrounding ecosystems (Bridgham et al., 2008;Waddington et al., 2015), and as fire-reducing land covers (Turetsky & Louis, 2006).
Accelerated vegetation growth and species shifts are predicted to enhance fuel connectivity with forest upland (Thompson, Parisien, et al., 2017).Here, SRI enhanced vascular vegetation growth rates, resulting in taller vegetation in bogs, fens, and adjacent ecotones compared with LRI fire (Figures 4 and 5).Accelerated growth occurred primarily in shrub and deciduous species within open peatlands, such as aspen (P.tremuloides), alders (Alnus spp.), willows (salix spp.), and Labrador tea (R. groenlandicum; Figures 4 and 6).While some of these species are commonly observed regrowing post-fire (Benscoter & Vitt, 2008), the increased occurrence of fire within open areas may have increased available nutrients such as nitrogen (Shenoy et al., 2013), resulting in faster regrowth and greater species diversity than that found in LRI-burned peatlands.Shifting fire regimes influence nutrient and moisture availability (Thompson, Parisien, et al., 2017), which may be associated with vegetation growth/mortality observed as a proxy indicator for environmental processes, including energy and mass exchanges.Deciduous species have greater demands on water resources, which could enhance transpiration losses in peatlands with greater regeneration of deciduous shrubs and trees instead of conifers (Johnston et al., 2015;Johnstone et al., 2004).While carbon losses from peatland burning contribute to atmospheric C emissions, Mack et al. (2021) demonstrate that rapid re-establishment of deciduous species may offset short-term C losses from fire.Since deciduous stands typically have lower flammability, resulting in lower severity fires, mixed conifer, and deciduous species could slow fire spread into peatlands (Parisien et al., 2011), though they also enhance the drying of peat fuel (Malhotra et al., 2020;Thompson et al., 2019).Within the first While early rapid regeneration of deciduous species following fire is not uncommon (Johnstone et al., 2004), growth rates of postfire vegetation within bogs and fens and between fire intervals were different, which could indicate vulnerability to long-term change.
Bogs demonstrated a greater response to SRI fire regimes than fens, with increased rate of vascular vegetation growth (Figure 5), especially in open bogs (Figure 6a).One potential explanation is that fens are hydrologically connected to their surroundings and have more continuous inputs of ground/surface water (Goodbrand et al., 2019;Kuhry et al., 1993).Sherwood et al. (2013) and Kettridge Turetsky et al. (2015) observed greater evapotranspiration rates from fens following fire, resulting in enhanced shrub/vascular vegetation productivity.Bourgeau-Chavez et al. ( 2020) also observed lower burn severity in fens, which, when combined with greater evaporative potential, produces conditions favorable for the encroachment of deciduous species (Grzesik et al., 2022;Jones et al., 2022;Langdon et al., 2020;Lukenbach et al., 2017;Moore et al., 2022).However, trait plasticity and adaptability of shrubs to varied environmental conditions, including hydrology, require further assessment (Dabros et al., 2022;Moore et al., 2022;Turetsky et al., 2008).
In comparison, domed bogs maintain drier peat soil conditions than fens, making them more susceptible to deep smoldering fire (Bourgeau-Chavez et al., 2020;Price & Maloney, 1994).Here, some bogs expanded immediately following fire due to vegetation mortality and loss of peat substrate, particularly within ecotones.
However, these areas also experienced significant post-fire vegetation encroachment (Figures 4 and 6a layers results in the development of a burn moat and water movement/flooding of the moat, particularly during wet periods (Ingram et al., 2019;Lukenbach et al., 2017), which can result in vegetation mortality, especially along peatland margins (Figures 4 and 8).
Furthermore, reduced elevation of the peat surface (and especially within hollows) maintains the high moisture characteristics of peatlands and reduces recolonization of upland species (Ott et al., 2006;Granath et al., 2010;Lukenbach et al., 2017).For example, bogs experienced high vegetation growth rates in the early years following both SRI and LRI fire.Rates of growth were significantly reduced between 2016-2018and 2019 (Figures 4c,d and 5a) (Figures 4c,d and 5a) due to high early summer precipitation and peatland flooding, potentially maintaining the bog areas over time (Ingram et al., 2019;Lukenbach et al., 2016;Waddington et al., 2015).
Geomorphology, peatland shape, and proximity to upland forest types have low to moderate influence on post-fire vegetation changes.Geomorphology plays an important role in peatland hydrology (Devito et al., 2016;Kuhry & Turunen, 2006), and post-fire regeneration is likely to be influenced by the potential for flooding (in hydrologically disconnected SRI-burned bogs) or evaporative losses (fens).Increasingly complex peatland shapes also correlate with greater rates of vegetation growth, especially following LRI fire (Figure 7a).Finally, mixed-wood forests had a greater influence on both bog and fen vegetation growth.However, this declined as the proportional area of forests surrounding peatlands increases (Figure 8b).This is somewhat counterintuitive, given the potential for the expansion of aspen suckers into peatlands; however, the first 8 years represent only the early phase of post-fire regeneration, and the influence of the remaining seedbed in peatlands following fire may yet manifest (Ellison & Bedford, 1995;Guêné-Nanchen et al., 2022;Langdon et al., 2020).

| Ecotones indicate that the greatest changes are yet to come
Ecotones are at greatest risk for fluctuations in water table (Dimitrov et al., 2014), flooding (Lukenbach et al., 2017;Ingram et al., 2019), and deeper burns (Hokanson et al., 2016;Morison et al., 2021;Lukenbach et al., 2017;Wilkinson, Tekatch, et al., 2020;Wilkinson, Verkaik, et al., 2020) resulting in greater environmental extremes and more rapid changes that could either enhance vascular vegetation growth or induce mortality especially following SRI fires.Here, peatland ecotones and adjacent peatlands demonstrated significantly greater changes in both rates of vegetation growth and decline (Figure 4) as well as the proportional area of vascular vegetation within peatlands and ecotones (Figure 5) following SRI as opposed to LRI fires.More dynamic responses to changing fire regimes may enhance variations in moisture conditions between fires (Hokanson et al., 2016).
If conditions increase in the periodicity and extent of drying, this could further exacerbate the potential for more intense and unpredictable fires (Kettridge Lukenbach et al., 2019;Turetsky et al., 2011).Additionally, burning of ecotones redistributes soil nutrients (Binkley & Fisher, 2019), further enhancing shrub and tree encroachment (Nelson et al., 2021;Shi et al., 2022).Given their proximity to peatlands, expanding ecotonal areas could introduce new seed beds into the peatlands and reduce overall peatland area while enhancing the shape complexity of peatlands (Figure 7; Langdon et al., 2020).Here, we find that ecotones are more responsive to changing conditions (Bhatti et al., 2006;Goud et al., 2018) (Figures 4 and 5) and are greater indicators of the impacts of climate change on ecosystems as these are often the first areas to change, indicated by changing vegetation (Neilson, 1993;Noble, 1993;Peteet, 2000).Therefore, rapid shifts in post-fire vegetation growth will likely affect peatland function for many years to come.

| CON CLUS I ON S AND FUTURE RE S E ARCH
Peatlands and adjacent ecotones burned by SRI fires tend to have the greatest rates of vegetation change (growth and mortality) compared with peatlands burned during an LRI fire.However, peatland/ upland ecotones demonstrated the greatest changes in post-fire vegetation change.While we expected ecotones to be more responsive to changing conditions (Bhatti et al., 2006;Goud et al., 2018), the rate at which vegetation is regenerating and, in many areas, expanding into peatlands is a potential concern as these processes increase forest fuel connectivity and propensity for burning (Senici et al., 2015;Thompson et al., 2019).
Additionally, shrubification of peatland centers increases transpiration and enhances drying (Moore et al., 2022), putting these peatlands at greater risk for fire due to the loss of peat moss through subsequent burns and increased ladder fuels (Shetler et al., 2008;Walker et al., 2020), thereby diminishing the traditional function of peatlands as fire refugia.We also observed the opposite response associated with vegetation mortality (height reduction) along ecotonal margins following a period of increased precipitation.The potential for enhanced flooding, especially within bog ecotones following SRI fire, could enhance peatland maintenance where hydraulic gradients are favorable.
This study is important because it presents early post-fire vascular vegetation recovery across hundreds of peatlands following SRI and LRI fires.With increasing frequency of fire in peatlands and the associated climate warming positive feedback (Thompson, Mendoza, et al., 2017), understanding the peatland responses to shortened fire return intervals is vital.Even within our study area, fire return intervals are shortening.In May 2023, another large fire burned across peatlands affected by the Utikuma Complex fire in 2011, shortening the fire return interval within the peatland complex to 12 years (Alberta, 2023).Relatively little has been examined regarding the impacts of increased fire frequency on peatlands, and as such, there are many opportunities for future research.We suggest future research focus on (1) determining the shrub vegetation can be a proxy indicator of shifting hydrology, what are the rates of post-fire regeneration of peatland vegetation following SRI versus LRI fires within the broader geomorphological groups found across the study area? 2. Are peatlands that undergo shortening fire return intervals shifting toward upland vegetative structural characteristics compared with peatlands that experience longer return interval fires? 3. How do rates of change vary within and between peatlands burned by SRI versus LRI fire associated with environmental gradients (peatland class, shape, and proximity to adjacent upland forest species)?F I G U R E 1 Photographs taken during the 2019 field campaign of (a) an unburned peatland, looking toward the center of the peatland from approximately one-third distance (of the total width) into the peatland; (b) a burned peatland following the Utikuma complex fire in 2011, looking into the center of the peatland from about one-third distance toward the unburned upland forest on the other side of the peatland; (c) an unburned upland forest to peatland ecotone, looking from approximately the upland forest edge through the ecotone and into the peatland beyond; (d) a narrow, burned ecotone (represented by sedges in the foreground) with the burned peatland beyond the ecotone.
mariana), bog birch (Betula glandulosa), and Labrador tea (R. groenlandicum), with Sphagnum moss species(Devito et al., 2016;Jones et al., 2022).Post-fire ecotones include alder species (Alnus viridis and Alnus crispa (Ait.)Pursh.), paper birch (Betula paperifera), and willow species (Salix spp.) (Figure1,Jones et al., 2022).F I G U R E 2 (a)The Utikuma Region Study Area (URSA) is located within the Boreal Plains Ecozone approximately 100 km north of Slave Lake, Alberta, Canada (inset).(b) Polygons indicate peatland class and form found within the areas burned by fire in 1956 only, 2011 only (LRI), and in the overlap of both fires (SRI).Ecotones exist along the perimeter between peatlands and upland forests (illustrated in dark green), where upland forests occur in gray areas.Wetland classification was modified from Chasmer et al. (2016) to include wetland class (bogs and fens, but also marshes, swamps and shallow open water; not included in this study) and form (predominantly open, with sedges and mosses, shrubby, or treed; Chasmer et al., 2024).
to include class (bog and fen) and form (open/sedge/moss, shrub, and treed) peatlands (Figure 3; classification available via Chasmer et al., 2024).Peatland class and form were identified in 2008 (pre-fire) remotely sensed data using a hillshade model from the 2008 lidar DEM, the CHM, and high-resolution optical SPOT imagery.Two hundred and twenty-one fens and 558 bogs were classified, and 30 were verified for classification accuracy from site visits, with 81% of peatlands correctly classified.About 18% of the 1956 burned area, 31% in the 2011 LRI fire, and 36% in SRI (burned both in 1956 and 2011) fire areas were classified as peatlands (Figure 2).
analysis was also used to determine if relationships exist between surrounding dominant upland forest types (deciduous, mixed wood, or coniferous) within the surrounding uplands, also based on prefire classifications fromChasmer et al. (2016) andBourgeau-Chavez et al. (2017).
Descriptive and statistical tests were performed to quantify differences between LRI and SRI fires in peatland classes/forms and ecotonal areas (R Core Team, 2021).All measured data variables were tested for normality using a Shapiro-Wilk test.Two-sample Kolmogorov-Smirnov tests and Mann-Whitney U-tests were applied to non-parametric variables to describe the entire data distribution, while normally distributed variables were tested using independent t-tests.To determine the impact of wetland form on vegetation growth, data were transformed logarithmically and then tested using a 2 × 3 factorial design with ANOVA.Differences in significance in variances for non-parametric variables were tested using the Fligner-Killeen test.Regression analysis was used to determine the relationship between peatland change in vegetation height, environmental drivers, and proximal landcover influences.3 | RE SULTS3.1 | Spatial rates of vegetation change in peatlands and ecotones following SRI versus LRI fireIn this study, peatlands largely intercepted two geomorphic features: fine-textured hummocky moraine and fine-textured glaciolacustrine plains.Peatlands in the LRI burn area were relatively evenly distributed among these geomorphological landform classes, TA B L E 1 Specifications of lidar data acquisitions (all systems by Teledyne Optech, Canada).

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Schematic of ecotone identification using lidar data.Iterative methods estimate the length (s) of the ecotones (s ecotone , point B) for each individual peatland, based on elevation differences (Δ z) and vegetation height difference (Δ h) derived from pre-fire (2008) airborne lidar data.
Figure 4 illustrates change in canopy height between lidar surveys collected in 2011 and 2016, 2016-2019, and across all years within peatlands and ecotones.In Figure 4, rates of vegetation growth proliferated in the early years, post-fire, especially in SRI ecotones, where average growth rates of trees and shrubs, including Salix (spp.) and Alnus (spp.), were 1.45 m year −1 (standard deviation [SD] = 0.41 m).Rates of growth were slightly reduced within ecotones following LRI fire (average = 1.35 m year −1 ; SD = 0.47 m) but were not significantly different from SRI fire (p = .19).Growth rates were much lower F I G U R E 4 Annual vegetation height changes (>0.15 m) within peatlands and ecotone areas following the 2011 fire in LRI-and SRI-burned areas, where (a-c) represent annual change in height in the early years following fire (up to 5 years), (d-f) represent later early stage growth from years 5-8 as vascular vegetation growth slows, and (g-i) represent annual change since the fire to 2019.Background light gray-scale image is a hillshade model from the lidar DEM, which indicates topographic variability indicated by shadows.

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Average height and split distributions of height percentiles in peatlands ("Peat") and ecotones in (a) bogs and (b) fens determined from ~6000 randomly distributed points within the strata of the broader study area.The vertical redline indicates the fire event in 2011, which occurred prior to lidar data collection in 2011.The split violin plots represent on the left, the distribution of the LRI strata, and the right the SRI strata.Horizontal lines indicate the 25th, median, and 75th percentiles.Significance between SRI and LRI per year (i.e., peat LRI vs. peat SRI per year) is indicated at the following confidence levels: 95% (*), 99% (**), and 99.9% (***).
are more typical of past fire regimes.While a few studies have quantified the impacts of SRI versus LRI fires on peatlands, this study used time-series airborne lidar data to quantify the spatial distribution and drivers of vegetation height change in the early post-fire period across a range of environmental conditions.This is important for the assessment of peatland vegetation succession following wildland fire, the maintenance of peatlands as critically important F I G U R E 6 Sankey diagram of proportional area coverage of tree/shrub vegetation by total area by land cover type within SRI and LRI in 2019 in (a) bogs and bog ecotones; and (b) fens and fen ecotones compared with 2008, 3 years pre-fire area vegetation conditions.Proportions are subdivided within each peatland class and adjacent ecotones.

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Nonlinear regressions, comparing perimeter:area ratios with annual percent growth between (a) LRI fire and (b) SRI fire within peatlands.
). Deep smoldering into peat F I G U R E 8 Linear regression models of the differences between percent growth compared with the percent area of surrounding upland type in (a) LRI-burned peatlands and (b) SRI-burned peatlands.Vegetation changes for the period between 2016 and 2019 were compared as a proportion of the rate of growth of shrubs and trees within peatlands between 2002 and 2008.Each data point represents the average of a single peatland.