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All content in this area was uploaded by Adam F. A. Pellegrini on Feb 26, 2021
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Articles
https://doi.org/10.1038/s41559-021-01401-7
1Department of Earth System Science, Stanford University, Stanford, CA, USA. 2Department of Plant Sciences, University of Cambridge, Cambridge, UK.
3Department of Natural Resources & Environmental Science, University of Nevada, Reno, NV, USA. 4Department of Environmental Systems Science,
Institute of Integrative Biology, ETH Zürich, Zürich, Switzerland. 5Physical and Life Sciences Directorate, Lawrence Livermore National Laboratory,
Livermore, CA, USA. 6Department of Ecology and Evolutionary Biology, Yale University, New Haven, CT, USA. 7Southern Research Station, USDA Forest
Service, Auburn, AL, USA. 8National Parks Service, Sequoia & Kings Canyon National Parks, Three Rivers, CA, USA. 9Department of Forestry, Wildlife, and
Fisheries, University of Tennessee, Knoxville, TN, USA. 10Scientific Services, South African National Parks, Kruger National Park, Skukuza, South Africa.
11School of Natural Resource Management, Nelson Mandela University, Port Elizabeth, South Africa. 12Southern Research Station, USDA Forest Service,
Pineville, LA, USA. 13Department of Ecology, Evolution, and Behavior, University of Minnesota, St. Paul, MN, USA. 14Department of Plant and Microbial
Biology, North Carolina State University, Raleigh, NC, USA. 15School of Forestry & Wildlife Sciences, Auburn University, Auburn, AL, USA. 16Department of
Agriculture and Fisheries, Queensland Government, Brisbane, Queensland, Australia. 17Rocky Mountain Research Station, USDA Forest Service, Flagstaff,
AZ, USA. 18Forestry Program, Holdsworth Natural Resources Center, University of Massachusetts, Amherst, MA, USA. 19Department of Biology, Stanford
University, Stanford, CA, USA. 20Department of Forest Resources, University of Minnesota, St. Paul, MN, USA. 21Hawkesbury Institute for the Environment,
Western Sydney University, Penrith, New South Wales, Australia. 22School of Geosciences, University of Edinburgh, Edinburgh, UK. 23College of Natural
Resources, University of Wisconsin–Stevens Point, Stevens Point, WI, USA. 24Department of Geography, University of Colorado-Boulder, Boulder, CO,
USA. 25Division of Forestry and Natural Resources, West Virginia University, Morgantown, WV, USA. 26Northern Research Station, USDA Forest Service,
Houghton, MI, USA. 27Department of Integrative Biology, University of Wisconsin, Madison, WI, USA. 28Tall Timbers Research Station, Tallahassee, FL,
USA. 29Woods Institute for the Environment, Stanford University, Stanford, CA, USA. 30Precourt Institute for Energy, Stanford University, Stanford, CA,
USA. ✉e-mail: afapellegrini@gmail.com
Ecosystem resilience to changing fire regimes1–3 will be a key
determinant of how terrestrial ecosystems respond to global
change3,4. Fire is a pervasive disturbance, burning ~5% of
global land area each year and releasing carbon stored in plant bio-
mass equivalent to ~20% of anthropogenic fossil fuel emissions5.
Historically, much of this carbon is re-sequestered through time as
plants recover and regrow, then lost again in the next fire. However,
in many systems, changes in climate and land use have shifted fire
frequencies, potentially changing the ability of plants, especially
trees, to regrow between fires1,6–8.
Although trees can be top-killed in fires, previous studies have
presented mixed results for the effect of burning frequency on tree
biomass9–13, highlighting the need to understand the factors lead-
ing to different fire effects. Climate extremes can moderate fire
effects on trees by influencing both tree growth and mortality as
well as fire intensity: sites with strong wet seasons have trees lack-
ing physiological adaptations to burning14, and sites with strong dry
seasons can have intense fires and droughts15. Furthermore, trees
in different ecosystems respond to changes in fire frequency dif-
ferently, partly owing to fuel load and composition but also their
physiology16. For example, traits conferring physiological protec-
tion from heating during fire and/or the capacity to colonize and
regrow rapidly could decrease tree biomass losses under frequent
burning13,17,18. Tests of these hypotheses using observational data are
Decadal changes in fire frequencies shift tree
communities and functional traits
Adam F. A. Pellegrini 1,2 ✉ , Tyler Refsland 3, Colin Averill4, César Terrer 1,5, A. Carla Staver 6,
Dale G. Brockway7, Anthony Caprio8, Wayne Clatterbuck9, Corli Coetsee10,11, James D. Haywood12,
Sarah E. Hobbie 13, William A. Hoffmann 14, John Kush15, Tom Lewis16, W. Keith Moser17,
Steven T. Overby17, William A. Patterson III18, Kabir G. Peay 19, Peter B. Reich 20,21, Casey Ryan 22,
Mary Anne S. Sayer12, Bryant C. Scharenbroch23, Tania Schoennagel24, Gabriel Reuben Smith 4,19,
Kirsten Stephan25, Chris Swanston 26, Monica G. Turner 27, J. Morgan Varner28 and
Robert B. Jackson 1,29,30
Global change has resulted in chronic shifts in fire regimes. Variability in the sensitivity of tree communities to multi-decadal
changes in fire regimes is critical to anticipating shifts in ecosystem structure and function, yet remains poorly understood.
Here, we address the overall effects of fire on tree communities and the factors controlling their sensitivity in 29 sites that expe-
rienced multi-decadal alterations in fire frequencies in savanna and forest ecosystems across tropical and temperate regions.
Fire had a strong overall effect on tree communities, with an average fire frequency (one fire every three years) reducing stem
density by 48% and basal area by 53% after 50 years, relative to unburned plots. The largest changes occurred in savanna eco-
systems and in sites with strong wet seasons or strong dry seasons, pointing to fire characteristics and species composition as
important. Analyses of functional traits highlighted the impact of fire-driven changes in soil nutrients because frequent burning
favoured trees with low biomass nitrogen and phosphorus content, and with more efficient nitrogen acquisition through ectomy-
corrhizal symbioses. Taken together, the response of trees to altered fire frequencies depends both on climatic and vegetation
determinants of fire behaviour and tree growth, and the coupling between fire-driven nutrient losses and plant traits.
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limited by the collinearity between fire regime and environmental
variables at individual sites19,20; instead, tests require experimental
manipulations across broad geographic ranges that are long-term,
ideally multi-decadal, because fire-driven mortality and top-kill can
take decades to impact the tree community10,13,21.
In addition to external factors, changes within the tree commu-
nity can modify the effect of repeated fires through the increasing
abundance of species that are more tolerant of the direct and indi-
rect effects of fire. Fire effects may be buffered by the colonization
and growth of species with structural traits that help trees to avoid
heat-induced mortality, but also of species with different nutrient
use and acquisition traits that help plants optimize for fire-driven
changes in nutrient availability22,23. Such shifts in nutrient use and
acquisition strategies have ecosystem-level implications, as changes
in these traits significantly influence soil nutrient cycling24,25. For
example, plants that form symbioses with ectomycorrhizal (ECM)
fungi, arbuscular mycorrhizal (AM) fungi, or nitrogen-fixing bacte-
ria may be better equipped to access nutrients26 that can be depleted
under frequent burning22. The distinction between strategies is
key because ECM plants tend to slow nutrient cycling and pro-
ductivity, while arbuscular and nitrogen-fixing species accelerate
nutrient cycling and productivity24–26. Tissue stoichiometry is also
relevant, with species containing lower nutrient concentrations bet-
ter equipped to tolerate low nutrient supply, but in turn contributing
to slower nutrient turnover and lower nutrient availability24,25. Thus,
how fire filters for species with different nutrient use and acquisi-
tion strategies can influence long-term site productivity.
Here, we evaluate how changes in fire frequency alter tree
communities, and how climate, vegetation composition and soils
influence the variability in the sensitivity of trees to changing fire
regimes. Furthermore, we quantify how fire changes the distribu-
tion of functional traits in the tree community to evaluate multiple
nutrient use traits that are indicative of tree responses to fire-driven
changes in nutrients. We combined published data with new sur-
veys on tree populations from 29 sites and 374 plots in four con-
tinents. At 27 of the sites (324 plots), surface fire frequency was
experimentally manipulated or varied naturally for 16–64 years
(mean of 30 years), and at two sites (50 plots), natural variation in
crown fire frequency presented a natural experiment. We focused
on tree responses because of their importance for carbon storage,
ecosystem productivity, and nutrient cycling23,27,28. The sites cover
locations in North America (coniferous, deciduous and mixed for-
ests, and broadleaf savanna), South America (tropical savanna and
temperate shrubland), Africa (broadleaf savanna) and Australia
(wet and dry eucalyptus forests), all of which are ecosystems that
experience frequent burning (Extended Data Fig. 1, Supplementary
Table 1 and Supplementary Information). For our main analysis, we
define ecosystems broadly based on dominant fuel type (grass ver-
sus tree litter) and tree functional composition (angiosperm broad-
leaf versus gymnosperm needleleaf trees), but test the robustness of
our conclusions to other classifications. Each surface fire site con-
tains replicate plots including an unburned treatment and differ-
ent prescribed burning frequencies (Extended Data Fig. 2), where
fire frequencies ranged from approximately one fire every decade
to one fire every year (Supplementary Table 1). We evaluated the
effects of fire alone and in combination with environmental covari-
ates using model selection (Supplementary Information) to test the
importance of climate, soil and species composition in modifying
the effect of fire.
Results and discussion
Effects of repeated burning compounded over multiple decades.
There were clear overall effects of fire treatments on tree population
size. Stem density (stems per hectare) and basal area were lower in
frequently burned plots relative to infrequently or unburned plots
(see Figs. 1a and 2a for effects in each site and Figs. 1b and 2b for
the response ratio across sites). A comparison between the most
extreme fire frequency treatments using response ratios illustrated
that density and total basal area were 44 ± 25% and 73 ± 45% lower,
respectively, in the most frequently burned plots compared with
unburned plots (Figs. 1b and 2b). The differences between fire treat-
ment effects were larger when fire frequencies contrasted more (for
example, the effect of intermediate versus high fire frequency was
52% and 28% lower than the effect of low versus high frequency
for both stem density and basal area, respectively; Figs. 1b and 2b,
and Supplementary Table 2, errors are 95% confidence intervals).
Duration of exposure to altered frequencies was also significant, as
sites with longer durations of altered fire frequencies had larger dif-
ferences between fire treatments, with the slope between duration
and community size growing more negative with more frequent
burning (density: F1,280 = 8.4, P = 0.004; basal area: F1,289 = 23.3,
P < 0.001; Figs. 1c and 2c). For example, relative to unburned plots,
plots with a three-year fire-return interval had 26% lower stem
density and 27% lower basal area after 30 years; the differences
increased to 48% lower stem density and 53% lower basal area after
50 years (Figs. 1c and 2c, and Supplementary Table 3; see Extended
Data Fig. 3 for non-transformed results). In annually burned plots,
the most extreme fire frequency, burned plots had 63% lower stem
density and 72% lower basal area than unburned plots after 50 years
(Figs. 1c and 2c, and Supplementary Table 3). Thus, both duration
of experiment and prescribed fire frequency help to reconcile the
variable effects of fire across studies. Accounting for these factors
illustrates that the effects of changing fire frequencies may take sev-
eral decades to become substantial, but the impact of the changes
will continue to increase for many decades.
Fire type was also important, with frequent crown fires affect-
ing tree populations to a greater degree than frequent surface fires.
Comparison of 50 plots in needleleaf forests that experienced natu-
ral variability in the frequency of stand-replacing crown fires (that
is, wildfires) illustrated that stands with shorter fire-return intervals
had significantly lower tree densities, especially when plots with
the shortest return intervals were considered (F1,26.5 = 5.2, P = 0.03
and F1,21 = 10.3, P = 0.004, respectively; Extended Data Fig. 4).
Experimental manipulation of surface fire frequency (that is, pre-
scribed fires) in needleleaf forests in North America showed that
stem densities were lower in more frequently burned plots, but less
so than differences caused by frequent crown fires (F1,47.1 = 17.2,
P = 0.001; Extended Data Fig. 4). The large effect of short-interval
crown fires on tree communities, supported by studies from other
regions29,30, highlights the importance of higher fire intensities hav-
ing more severe effects.
Fire sensitivities differed across biomes and climates. Although
fire frequency had a large overall effect on trees, there was substan-
tial variability in the sensitivity across sites. Part of the variability was
attributable to fire history prior to the establishment of the experi-
ment. In forest sites that burned regularly in the decades before the
onset of the experiment, fire exclusion resulted in stem density and
basal area being 47% (±40%) and 65% (±71%) higher, respectively,
than treatments that maintained historical burning frequencies (see
Supplementary Table 1 for site fire histories). In contrast, the rein-
troduction of prescribed fire into forests that had not burned for
several decades before the onset of the experiment had relatively
minimal effects. These results probably differ from those commonly
observed with wildfires, which can have larger effects in forests with
a history of fire exclusion due to high fuel accumulation31,32 because
prescribed surface fires are less intense. In savannas, where the fire
experiments were all initiated in landscapes that burned regularly
in the decades preceding the experiment, fire exclusion resulted
in basal area increasing by 41% (±18%), but increasing fire fre-
quency resulted in basal area declining by 48% (±22%), while stem
density remained unchanged, relative to an intermediate interval
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that maintained the pre-experiment frequency (statistics from log
response ratios ±95% confidence intervals, P < 0.001 for both basal
area comparisons; Supplementary Information). Taken together, the
largest effects of altered fire frequencies were owing to fire exclusion
in landscapes that had burned regularly for the past few decades.
Model selection illustrated that climate, vegetation type and con-
tinent played significant roles in explaining the variability in the sen-
sitivity of trees to fire. Fire effects were largest in areas that received
more rainfall in the wet season, less rainfall in the dry season and
had lower mean annual temperatures (F1,285.1 = 50.3, P < 0.001;
F1,285.6 = 7.5, P = 0.007; and F1,280.8 = 16.8, P < 0.001, respectively; Fig.
3a–c and Supplementary Table 4; see Supplementary Table 5 for
stem density). For example, plots that experienced more frequent
burning (two fires every three years, one standard deviation above
mean frequency) had 65% lower tree basal area than unburned plots
in sites with high wet season precipitation. In sites with average wet
season precipitation, the difference between the same treatments
was only 22% (Fig. 3a, Extended Data Fig. 5 and Supplementary
Table 4; see Supplementary Information for details on calculations).
Dry season precipitation had the opposite correlation with fire
effects: sites with lower precipitation in the dry season experi-
enced twice as large an effect of fire on basal area (frequent burning
resulted in 44% versus 22% lower tree basal area in sites with low
versus average dry season precipitation; Fig. 3c and Supplementary
Table 4). The contrasting response to precipitation in the wet versus
dry season is consistent with our understanding that fires are most
intense in areas with stronger wet seasons (leading to more fuel)
and more severe dry seasons (lower fuel moisture), thus contrib-
uting to potential losses with more frequent burning33–35. Rainfall
in the dry season probably also influences fire effects by determin-
ing the water available for tree growth. Neither soil texture nor soil
carbon explained the sensitivity to changing fire frequencies across
sites (Supplementary Table 4). Thus, climate was a key determinant
of fire effects, with more moderately seasonal sites being the least
sensitive to changing fire regimes.
The effect of fire on tree basal area also differed across eco-
systems, with frequent burning having a larger effect on tree
basal area in savannas relative to broadleaf and needleleaf forests
–5
–4
9
8
7
6
5
4
10 29 48 67 67 6710 29
Years of repeated burning
48 10 29 48
–3
–2
–1
0.5
0
–0.5
–1
0
1
2
a
c
b
Bauple
Dry Peachester
Highland Rim
Lombard
Morton
University Missouri
Alapaha
Chimney Spring
Escambia
Hitchiti
Kings Canyon
Kisatchie
Limestone
Sequoia
Tall Timbers
Wharton
YellStonL
YellStonH
Cedar Creek
IBGE
Kruger Mopani
Kruger Pretoriuskop
Kruger Satara
Kruger Skukuza
Lamto
La Pampa
Marondera
Broadleaf forest
Needleleaf forest
Savanna
Fire frequency 0–0.21 yr–1 Fire frequency 0.22–0.66 yr–1 Fire frequency 0.67–1 yr–1
Low versus
high
Low versus
mid
Mid versus
high
Treatment comparisons
Partial residual log(stem density) Stem density log(RR)
Stem density log(RR)
Fig. 1 | Tree stem density declines with both fire frequency and duration of fire regime. a,b, log response ratios (RR) of stem densities and the
surrounding 95% confidence intervals. Shown are comparisons between ln (burned/unburned) treatments within each individual site coloured by broad
ecosystem categorization with the burned treatment being the most extreme fire frequency treatments in a (a; n = 25; Supplementary Table 1, ecosystem
groups based on broad differences in fire fuel and tree composition; Supplementary Information). Also shown are averages across all sites comparing the
different levels of fire frequency in studies with ≥3 levels; less frequent treatment always in denominator (b; n = 23, n = 17, n = 18; Supplementary Table
2). c, Partial residuals plot from a mixed-effects model including fire frequency, the number of years of repeated burning and their interaction for ln stem
density (n = 303; Supplementary Table 3); site was used as a random intercept. Panels are centred on cross-section values of one fire every ten years, one
every three years and one every year, but encompass a range of fire frequencies within each panel. IBGE, Reserva Ecológica do IBGE; YellStonL, Yellowstone
low severity; YellStonH, Yellowstone high severity.
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(F2,279.7 = 14.0, P < 0.001, model incorporating climate effects; Fig. 3d
and Supplementary Table 4). Relative to the unburned plots, basal
area in frequently burned plots was 5% lower in needleleaf forests
and 22% lower in broadleaf forests (Fig. 3d, burn frequency of two
fires every three years; Supplementary Information; continent dif-
ferences in Extended Data Fig. 6). In savannas, frequently burned
plots had 70% lower basal area relative to the unburned plots (Fig. 3d
and Supplementary Table 4). Interestingly, stem density responses
to fire frequency were qualitatively different between savannas and
forests (Supplementary Table 5). Stem densities increased with
more frequent burning in forests while basal area decreased, poten-
tially due to higher light availability and tree recruitment in the for-
ests. To assess the sensitivity of these findings to our classification
of ecosystem type, we also tried using a subdivided classification
by partitioning broadleaf forests into oak and eucalypt types, and
needleleaf forests into those that transitioned between oak and pine
dominated (Supplementary Table 1). When included in the final
model, the subdivided vegetation classification still had a significant
main effect (F4,16.8 = 11.1, P = 0.0001) and a significant interaction
with fire frequency (F4,282.5 = 7.1, P < 0.001; Extended Data Fig. 7),
with basal area in savannas responding the most to changes in fire
frequency (Extended Data Figs. 7 and 8).
Fire changed the functional composition of tree communities.
Fire-driven changes in basal area and stem density are important for
ecosystem function, but fire can also impact ecosystems by chang-
ing the functional composition of trees. To address this, we anal-
ysed functional composition in only the experiments from North
America (77 tree species, 16 sites, 181 plots) because trait data were
available there to (1) categorize species by nutrient acquisition strat-
egies and (2) assign wood, leaf and root traits related to growth, sur-
vival and nutrient use strategies.
We found that structural traits were important for explaining
cross-site variability in the sensitivity of tree communities to fire.
Across sites, frequent burning impacted basal area more where
tree species had thinner bark and denser wood (bark thickness:
–1
0
1
2
3
4
5
Partial residual log(basal area)
Basal area log(RR)
–3.00
–2.00
–1.00
0
1.00
–5
–4
–3
–2
–1
0
1
2
a
c
b
Basal area log(RR)
Broadleaf forest
Needleleaf forest
Savanna
Bauple
Dry Peachester
Highland Rim
Lombard
Morton
Okmulgee
University Missouri
Wet Peachester
Alapaha
Chimney Spring
Escambia
Hitchiti
Kings Canyon
Kisatchie
Limestone
Sequoia
Tall Timbers
Wharton
Cedar Creek
IBGE
Kruger Mopani
Kruger Pretoriuskop
Kruger Satara
Kruger Skukuza
Marondera
Low versus
high
Low versus
mid
Mid versus
high
Treatment comparisons
Fire frequency 0–0.21 yr–1 Fire frequency 0.22–0.66 yr–1 Fire frequency 0.67–1 yr–1
10 29 48 67 67 6710 29
Years of repeated burning
48 10 29 48
Fig. 2 | Tree basal area declines with both fire frequency and duration of fire regime. a,b, log response ratios of basal area and the surrounding 95%
confidence intervals. Shown are comparisons within each individual site coloured by broad ecosystem categorization in the most extreme fire frequency
treatments ln (burned/unburned) (a; n = 23; Supplementary Table 1, ecosystem groups based on broad differences in fire behaviour and tree composition;
Supplementary Information). Also shown are averages across all sites comparing the different levels of fire frequency in studies with ≥3 levels; less
frequent treatment always in denominator (b; n = 23, n = 17, n = 17; Supplementary Table 2). c, Partial residuals plot from a mixed-effects model including
fire frequency, the number of years of repeated burning and their interaction for ln basal area (n = 309; Supplementary Table 3); site was used as a random
intercept. Panels are centred on cross-section values of one fire every ten years, one every three years and one every year, but encompass a range of fire
frequencies within each panel (ranges: unburned to one every five years; one every five years to two every three years; two every three years to every year).
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F1,164.3 = 9.2, P = 0.003; wood density: F1,165 = 10, P = 0.002; see
Supplementary Table 6 for both in the same model, and Fig. 4a,b).
However, within sites we found mixed evidence that fire filtered for
species with different structural trait values. Mean wood density
of the tree community tended to be lower in frequently burned
plots, potentially because of increasing dominance of gymnosperm
trees, which tend to have lower wood density. In contrast, we did
not observe any effect of fire on the mean bark investment of the
–1
0
1
2
3
4
a
b
c
d
Partial residual
log(basal area)
Partial residual
log(basal area)
–1
0
1
2
3
4
0.02 0.34
5–12
MAT (°C)
13–17 18–22
175–301
Precipitation wet quarter (mm):
302–450 451–737 12–79
Precipitation dry quarter (mm):
80–187 188–351
Broadleaf forest Needleleaf forest Savanna
0.66 0.34
Fire frequency (fires yr
–1
) Fire frequency (fires yr
–1
)
0.661 0.02 1 0.34 0.660.02 1 0.02 0.34 0.66 0.34 0.661 0.02 1 0.34 0.660.02 1
0.02 0.34 0.66 0.34
Fire frequency (fires yr
–1
) Fire frequency (fires yr
–1
)
0.661 0.02 1 0.34 0.660.02 1 0.02 0.34 0.66 0.34 0.661 0.02 1 0.34 0.660.02 1
–1
0
1
2
3
4
–1
0
1
2
3
4
Fig. 3 | Climate and ecosystem type modify effects of fire frequencies on tree basal area. a–d, Partial residual plots of the mixed-effects model illustrating
the interactive effects between covariates (site as a random intercept, n = 309 plots, n = 25 sites). Panels centred on cross-sectional values from one
standard deviation around the median (−1, 0, 1). MAT, mean annual temperature. Partial residual predictions account for the values of all other covariates
in the model. Comparisons of rainfall scenarios relative to the mean in the text used wet season precipitation of +1 standard deviation above the mean
(525 versus 375 mm yr−1) and dry season precipitation of −1 standard deviation below the mean (25 versus 133 mm yr−1). The duration of experiment held
at its mean of 28 years. All model fits are P < 0.05; statistics are in Supplementary Table 4.
Partial residual log(basal area)
0.25
10.5 mm 11.6 mm 16.0 mm 0.38 g cm
–3
0.54 g cm
–3
0.59 g cm
–3
0.50 0.75
Fire frequency (fires yr
–1
)
Arbuscular mycorrhizal Ectomycorrhizal
Fire frequency (fires yr
–1
)
Years of repeated burning
1.000
2.0
Relative BA (arc sine)
1.5
0
0.02 0.34 0.66 1 0.02
Fire frequency (fires yr
–1
)
0.34 0.66 1 10 29 48 67 10 29 48 67
0.5
1.0
14.8
Green leaf N (mg g
–1
)
Root N (mg g
–1
)
9.7
12.2
11
7
9
0.25 0.50 0.75 1.000 0.25 0.50 0.75 1.000 0.25 0.50 0.75 1.000 0.25 0.50 0.75 1.000 0.25 0.50 0.75 1.000
3
2
4
5
a
c d
b
Partial residual log(basal area)
2
1
3
4
5
Fig. 4 | Functional composition of tree communities both responds to and modifies the effects of frequent burning. a,b, Partial regression plots from
mixed-effects models with community-weighted means of bark investment (BI) scaled to a 10 cm stem size (a) and wood density (WD) as modifying
variables (b; n = 172 plots, n = 16 sites). Bark investment and wood density were included in the same model and were negatively correlated σ = −0.86.
Statistics are in Supplementary Table 6. Basal area is ln transformed. c, Relative basal area of nutrient acquisition strategies (arcsine transformed) of AM
versus ECM trees as a function of fire frequency. d, Community-weighted means of nutrient use strategies, showing green leaf and live root N as a function
of study duration. Statistics are in Supplementary Tables 7–9 (n = 172 plots and n = 16 sites), which also show data on litter N and resorption. Phosphorus
content data are shown in Extended Data Fig. 7.
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tree community (Supplementary Table 6), demonstrating that bark
investment at the community scale does not appear to change in
response to fire. Nevertheless, bark investment helped to predict
basal area loss patterns across broad biogeographic scales.
Frequent fire favoured conservative nutrient strategies. We found
evidence for compensatory responses in the nutrient use and acqui-
sition strategies of the tree communities under different fire fre-
quencies. On average, plots burned frequently for longer periods of
time were dominated by tree species with low nitrogen (N) concen-
trations in green and senesced leaves and roots, and that resorbed
a greater proportion of N before leaf senescence (P < 0.001 for all
variables; Fig. 4d and Supplementary Table 7). This is consistent
with the hypothesis that fire-driven soil N losses22 filter for species
with adaptations to low N conditions. Tissue phosphorus (P) con-
centrations also declined with frequent burning in leaves and litter
but not in roots (Extended Data Fig. 9 and Supplementary Table 7).
Given that P is less prone to being lost than N22,36,37, this result was
surprising and raises questions about how fire may change the P
economy of trees in an opposite way to changes in soil P availability.
Taken together, fire filters for species with more conservative nutri-
ent use strategies in both their leaves and fine roots across several
North American savannas and forests.
Fire also affected the relative abundance of nutrient acquisi-
tion strategies, evaluated by trees’ abilities to form symbioses38.
Trees that formed symbioses with ECM and AM fungi were the
most abundant nutrient acquisition strategies across our plots; eri-
coid and nitrogen-fixing trees were absent from most sites (Fig.
4c; categorization based on tree species’ taxonomy38). ECM trees,
which associate with fungal symbionts capable of acquiring N from
organic matter39, tended to be more successful in frequently burned
plots. The relative abundance of ECM trees increased from 85% in
unburned plots to nearly 100% in annually burned plots (Fig. 4c
and Supplementary Table 8). ECM trees were also more common
in warmer climates and on soils with low carbon concentrations
(Supplementary Table 8). ECM trees typically have lower con-
centrations of N and P in leaves, litter and roots than AM trees38
(Supplementary Table 9), suggesting the turnover in symbiont
composition may be driving the shift in stoichiometry of the tree
community. As repeated fire tends to decrease inorganic N and N
mineralization23, it makes sense that frequent fire causes the tree
community to shift towards species capable of obtaining N from
soil organic matter.
The tendency for frequently burned plots to have tree communi-
ties dominated by ECM species with low N and P content in leaves,
roots and litter indicates that frequent burning favours conserva-
tive nutrient use and acquisition strategies. This trend is probably a
result of fire-driven soil nutrient losses, which should favour species
better equipped to cope with low nutrient environments; however,
other factors such as phylogenetic conservatism of traits may be
at play38. These trait shifts may themselves reinforce an important
fire–nutrient feedback if N losses cause a decline in productivity
that limits the ability of trees to regrow between fire events, fur-
ther decreasing ecosystem N turnover. Our results unpack only one
part of this feedback loop (fire effects on traits), and further study
is needed to connect changes in strategies with the observed differ-
ences in nutrient availability and plant growth.
Our analysis also highlights several areas for future work. For
one, an improved representation of fire experiments in different
ecosystem types across continents in our dataset (for example,
tropical forests in Africa) would help further unpack the variabil-
ity across ecosystems and continents. Second, longitudinal data on
how fire effects emerge through time could assist with better under-
standing how the turnover in tree species composition influences
changes in total tree cover within experiments10,18,40. Third, consid-
ering other plant groups (for example, herbaceous plants) will help
obtain a more comprehensive picture of how shifting fire regimes
change ecosystem function. Finally, the extent of fire effects on plant
strategies across temperate ecosystems highlights the need for more
studies of plant strategies to consider fire. For example, the effect of
fire on fungal symbiosis strategies should be integrated into theories
seeking to explain their biogeographic distribution, which generally
rely on climatic factors alone41.
Our observation that changing fire frequencies shift both tree
basal area and density as well as trait composition is important
because it identifies two means by which fire can indirectly alter
carbon and nutrient storage in other ecosystem pools, such as
soil organic matter. For one, the observed multi-decadal decline
in tree populations mirrors multi-decadal shifts in soil carbon
and nitrogen22,23, both in timescale and across ecosystems. The
greater rates of change in tree population sizes in savannas and
broadleaf forests are consistent with the higher losses of mineral
soil carbon and nitrogen reported in those ecosystems relative to
needleleaf forests22,42. Furthermore, a shift towards conservative
nutrient use and acquisition traits is consistent with lower soil N
mineralization23,42. Consequently, our findings support hypotheses
that fire-driven changes in tree biomass inputs and the turnover
of plant traits may both contribute to changes in soil carbon and
nutrient pools42–45.
Conclusions
Widespread changes in fire regimes are likely to shift both the popu-
lation size and functional composition of tree communities, with
both factors affecting the storage and cycling of carbon and nutri-
ents. The effects are heterogeneous, however, with certain ecosys-
tems being especially sensitive, such as savannas experiencing rapid
encroachment of tree cover when fire is excluded. Climatic factors
were also key, as regions with extreme precipitation amounts in the
wet and dry seasons (high and low, respectively) changed more than
areas with more moderate precipitation. The effects of fire were
not limited to tree population sizes, but also extended to the func-
tional composition of the community. The convergent response of
frequent burning promoting conservative nutrient use strategies
indicates that fire impacts nutrient cycling not only over decadal
timescales, but also suggests that fire probably influences the evo-
lution of these plant strategies. Consequently, climatically sensitive
shifts in fire frequency, even when relatively low intensity, will alter
the structure and functioning of ecosystems through multiple direct
and indirect pathways.
Methods
Site descriptions and experimental designs. e majority of sites sampled are
in ecosystems that experience surface res (from re manipulation experiments,
n = 27). Our main analyses are based on the surface re experiments, but we
compare these data with a network of plots across n = 2 sites with natural
variability in the frequency of stand-replacing crown res to evaluate the eect
of re regime. We describe the sites briey in Supplementary Table 1 and present
detailed descriptions of site history in Supplementary Data 3. Most of the data were
obtained from existing studies, but we complemented these data with unpublished
surveys in the Sequoia and Kings Canyon sites. We identied these sites rst by
using a list of sites from a previous meta-analysis of multi-decadal changes in re
frequencies22. Not all of the sites in the other study contained vegetation surveys
and in some cases the authors of the other studies did not share data. We then
complemented these studies with other long-term re manipulation experiments
using a literature search and conference presentations. Finally, data from some of
the sites were collected specically for this study.
The surface fire experiments are mostly experimental prescribed burn plots.
The managers generally try to burn in a broad seasonal window (for example, a
spring fire in North America may occur anytime from March to May) to optimize
burn timing for the local fire conditions most suitable for their planned fire
intensities. Treatments were not applied in a uniform fashion across sites, which
is one motivation for using mixed-effects models (see description of our statistical
approach below).
The sites contained different land use histories before the establishment of
the experiment, which was not always documented in detail, but we describe key
factors in Supplementary Table 1 and Supplementary Data 3. We describe how we
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evaluated the potential role of land use history in ‘Testing role of fire and land use
history’ in the Supplementary Information.
Environmental data. Soil chemistry. We collected and analysed soil data
using several methods. First, we determined the dominant soil type using
either author descriptions or reported soil texture analysis. Second, we used
the highest resolution soil data possible (for example, soil samples from each
replicate plot within a re treatment), but some sites contained only site-level
soil properties. Consequently, we analysed overall eects of re on all sites
without any covariates, followed by a model that uses model selection to
account for collinearities among variables when testing for factors that modify
re eects. To extend data on soils across plots, we sampled soils (top 0–5 cm
of the mineral horizon) in 24 plots across four sites: Kings Canyon, Sequoia,
Limestone Flats and Chimney Springs. Each site contained three replicate plots
of an unburned treatment and a high re frequency treatment. We collected n = 5
pseudo-replicates within the true replicate plot, analysed the soils for carbon and
texture, and averaged within each plot. For chemical analyses, soils were sieved
to <2 mm, dried to constant weight and ground on a ball mill. Subsets were then
analysed using combustion on an elemental analyser at Stanford University;
duplicates were run every 10 samples (analytical accuracy was >90%). Texture
was analysed using the hydrometer method (adopted from the Kellogg Biological
Station, Long-Term Ecological Research, https://lter.kbs.msu.edu/protocols) on
sieved soils.
Climate. To obtain long-term climate averages at each site, we used WorldClim46.
WorldClim integrates data from 1970–2000 across 9,000–60,000 weather
stations and spatially extrapolates the values by integrating other covariates from
topography maps and satellite data (described in detail in ref. 46). The timespan of
the climate data overlaps the duration of the experiments reasonably well (mean
study initiation = 1983; mean survey year = 2006). Managers timed burning to
coincide with consistent weather conditions over the course of the experiment,
therefore we did not obtain high-resolution interannual variability in climate.
We focused on several climate variables based on ecologically relevant a priori
hypotheses: (1) precipitation partitioned into the driest and wettest quarters of the
year because precipitation influences fuel accumulation (primarily in the wettest
quarter) and fire conditions (primarily in the driest quarter); and (2) mean annual
temperature because of its large effect on a variety of biogeochemical processes.
Precipitation in wet and dry quarters are not as correlated with one another
but are highly correlated with mean annual precipitation and temperature
(Supplementary Table 10).
Ecosystem type. The vegetation composition at each site differs substantially,
ranging from diverse tropical savannas with dozens of tree species (for example,
Kruger sites) to monodominant coniferous forests (for example, Limestone
Flats and Chimney Springs). Classifying the sites into broad categories was
done methodologically, by balancing the need to maintain parsimony (and thus
statistical power) with accurately capturing how plant composition may modify
fire effects. Consequently, we performed two levels of classification: (1) a coarse
categorization separating savannas versus forests, and within forests treating
broadleaf and needleleaf forests separately, which we refer to as a vegetation type;
and (2) accounting for variability within forest types by partitioning broadleaf
forests into Myrtaceae (eucalypt) versus Fagaceae (oak) dominated, and needleleaf
forests into forests that are near completely dominated by needleleaf trees versus a
mixed forest containing both needleleaf and broadleaf trees, which we refer to as a
sub-vegetation type.
Quantifying environmental effects. Several methods exist to calculate variable
importance, with no clear optimal method47. We chose to use the regression
coefficients in the model to understand the sensitivity of basal area and stem
density to changes in relative values of each variable. Importantly, the models were
fitted to rescaled data by subtracting each value by the mean and dividing by the
standard deviation of the variable. Consequently, the product between the mean
value of a variable and its coefficient is always zero. Thus, we can compare the
relative impact of variables by comparing the magnitude of the fitted coefficients
because they reflect the potential change in basal area for a one standard deviation
change in a variable value.
Wet season precipitation. Wet season precipitation varied one standard deviation
above the mean versus at the mean (525 versus 375 mm yr−1, respectively). Fire
frequency varied from unburned to one standard deviation above the mean (two
fires every three years). Wet season precipiation is shown in Fig. 3a.
Dry season precipitation. Dry season precipitation was one standard deviation
below the mean versus at the mean (25 versus 133 mm yr−1, respectively). Fire
frequency varied from unburned to one standard deviation above the mean (two
fires every three years). Dry season precipiation is shown in Fig. 3c.
Vegetation type. Fire frequency effects were made using two levels of comparison.
Unburned plots versus burning at the mean frequency (one fire every three years)
and unburned plots versus burning at one standard deviation above the mean
frequency (two fires every three years). Vegetation type is shown in Fig. 3d.
Statistics. Fire eect calculations. We rst evaluated the overall eect of re
frequency and duration that frequency was altered on tree basal area and stem
density without considering any potential modifying role of covariates. To
accomplish this we analysed (1) a mixed-eects model containing re frequency,
re period and their interaction; and (2) log response ratios of stem density and
basal area relativized within each site. We excluded the 50 crown re plots for this
initial analysis. We tted the mixed-eects models with site as a random intercept.
e statistical design is nested because each site has several replicate plots receiving
dierent re treatments. In all cases of mixed-eects models, we tested for model
signicance using Satterthwaite’s approximation for degrees of freedom and a type
III analysis of variance48. In the event of an insignicant main eect but signicant
interaction, we tested whether the main eect could be dropped from the model
using a change in Akaike information criterion with a threshold of two.
Fire–environment interactions. For the plots with surface fires, we performed model
selection by incorporating covariates of climate, soil and plant composition into
mixed-effects models to test for pairwise interactions and possible collinearities
(see discussion below of collinearities). Finally, we constructed a full model
containing fire, climate, soil and composition variables based on our hypotheses
that these factors will interact with fire frequency as well as information gained
from the pairwise tests. There were several insignificant effects in the final model,
which we tested for removal using model selection with a threshold Akaike
information criterion of two. All variables were rescaled by subtracting the mean
and dividing by their standard deviation.
Traits. Bark thickness data were collected from a dataset in the Fire and Fuels
Extension to the Forest Vegetation Simulator (https://www.fs.fed.us/fmsc/ftp/
fvs/docs/gtr/FFEaddendum.pdf). Although broad syntheses of bark investment
exist for many tree species in North America, not all species contained data from
empirical measurements, thus we used the data from the Fire and Fuels Extension.
Bark thickness was assumed to scale linearly with stem diameter, which is generally
valid for smaller stems, but it is known that bark saturates with increasing stem
diameter49. The ability of bark investment to predict fire effects will probably
improve with better consideration of the non-linear relationship between bark
and stem diameter. We evaluate the relative bark investment, not absolute bark
thickness, which is based on bark investment as well as stem size.
Wood density was compiled from the literature using a global wood density
database50, supplemented with additional data51,52. We assigned a genus-level
average for 19 species lacking data. Plant tissue stoichiometry and mycorrhizal type
were determined using both trait data and phylogenetic trait estimates calibrated
to trait data used in a previous global analysis of plant mycorrhizal traits38.
Full data selection criteria are presented in ref. 38, but we describe them in the
Supplementary Information.
To test for fire effects on the relative abundance of symbiotic strategies, we
calculated the relative basal area of the different strategies (ECM, AM and the less
abundant ericoid mycorrhizal, non-mycorrhizal and nitrogen-fixing tree species).
Given the low occurrences of ericoid, non-mycorrhizal and nitrogen-fixing species,
we analysed the relative abundance of only AM and ECM species. We then fitted
mixed-effects models with relative basal area as the dependent variable and fire,
climate, broad vegetation type (broadleaf, needleleaf, savanna) and soil conditions
as the independent variables, each modified by a symbiont term. Relative basal area
was arcsine transformed. This analysis was conducted in the North American plots.
To test how fire influenced the trait composition of the community we
fit mixed-effects models to test the effect of fire and environmental factors in
explaining the community-weighted mean trait values. We do not include an
independent effect of either wood density or bark thickness because we are
primarily concerned with how they may modify fire effects. We also tested for
whether the symbiotic strategies differed in their traits. To do so, we assigned
symbiotic strategies and the dominant ecosystem in which they occurred to
different species. We then analysed linear models incorporating symbiotic strategy
and ecosystem type as additive effects.
Reporting Summary. Further information on research design is available in the
Nature Research Reporting Summary linked to this article.
Data availability
Wood density data are from refs. 50–52, plant tissue stoichiometry and mycorrhizal
type data are from ref. 38, and bark thickness data are from the Fire and Fuels
Extension to the Forest Vegetation Simulator (https://www.fs.fed.us/fmsc/ftp/fvs/
docs/gtr/FFEaddendum.pdf). Supplementary Data 1 and 2 contain the woody
population size data and Supplementary Data 3 contains land use data. Refs. 53–76
describe the fire experiments.
Received: 10 July 2020; Accepted: 25 January 2021;
Published: xx xx xxxx
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Acknowledgements
A.F.A.P. was supported by a NOAA Climate and Global Change postdoctoral
fellowship programme and the USDA National Institute of Food and Agriculture
grant 2018‐67012‐28077. R.B.J. received support from the Gordon and Betty Moore
Foundation. The experiments at the sites were organized and funded through the Cedar
Creek Long Term Ecological Research programme (DEB 1234162, 0620652, 1831944
and DBI 2021898), the National Park Service and Sequoia Parks Conservancy, and South
African National Parks. C.T. was supported by a Lawrence Fellow award supported by
the LLNL-LDRD Program under Project No. 20-ERD-055.
Author contributions
A.F.A.P. and R.B.J. conceived and designed the overall study. T.R., C.A. and C.T. helped with
data acquisition and provided feedback on statistical analyses. D.G.B., A.C.S., W.C., C.C.,
J.D.H., S.H.E., W.A.H., J.K., T.L., W.K.M., S.T.O., W.A.P., K.G.P., P.B.R., C.R., M.A.S.S., B.C.S.,
T.S., G.R.S., K.S., C.S., M.G.T. and J.M.V. provided data and/or assisted with interpreting the
field data from experiments. All authors contributed to the writing of the manuscript.
Competing interests
The authors declare no competing interests.
Additional information
Extended data is available for this paper at https://doi.org/10.1038/s41559-021-01401-7.
Supplementary information The online version contains supplementary material
available at https://doi.org/10.1038/s41559-021-01401-7.
Correspondence and requests for materials should be addressed to A.F.A.P.
Peer review information Nature Ecology & Evolution thanks Donald Falk and the other,
anonymous, reviewer(s) for their contribution to the peer review of this work.
Reprints and permissions information is available at www.nature.com/reprints.
Publisher’s note Springer Nature remains neutral with regard to jurisdictional claims in
published maps and institutional affiliations.
© The Author(s), under exclusive licence to Springer Nature Limited 2021
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Extended Data Fig. 1 | Distribution of sites. a, Map displaying the distribution of sites (dots) with the surface fire sites filled with black and the crown
fire sites filled with white. The coloration indicates the average fire frequency within a gridcell using 1. The sample size of plots is written adjacent to
the continent. b, distribution of sites in climate space overlying Whittaker’s biome distribution 77. (1 = tundra, 2=boreal forest, 3=woodland/shrubland,
4=temperate grassland/desert, 5=temperate forest, 6=temperate rainforest, 7=subtropical desert, 8=tropical forest and savanna, 9=tropical rainforest).
Dots colored according to broad vegetation type category. Plots span a mean annual temperate range from 5.2–27.3 °C and a mean annual precipitation
range from 408–2378 mm yr−1. c, aerial picture of two different fire treatment plots from Cedar Creek, a temperate oak savanna, where different fire
frequencies have created a stark biome boundary between forests in unburned plots and savannas in biennial burn plots.
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Extended Data Fig. 2 | Example of the experimental layout of a fire manipulation experiment. Example of the experimental layout of a fire manipulation
experiment taken from Cedar Creek (a temperate savanna in Minnesota, USA), where fires have been manipulated since 1964. Aerial imagery (taken in
2017) from the National Agriculture Imagery Program from the Farm Service Agency. Plots are outlined with a color corresponding to their fire frequencies
expressed in terms of number of fires per year (for example 0.33 is one fire every 3 years).
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Extended Data Fig. 3 | Untransformed data on stem density and basal area. Untransformed data on stem density (a-b) and basal area (c-d) as a function
of the duration that plots have been exposed to burning in the experiment (0 = unburned plots). Each dot represents a site and the dashed lines connect
treatments within sites. Columns represent two sets of fire frequency contrasts comparing unburned vs. the intermediate frequency in a and c, and
unburned vs. the high frequency in b and d (levels defined based on treatments within sites). Dots and bars based on mean and standard error calculated
across the replicate plots within a fire treatment in a site. Note y-axis is on a log10 scale.
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Extended Data Fig. 4 | Comparison between fire types. Comparison between fire types (surface in a, F1,94.3 = 50.6, p < 0.001, n = 9 sites and n = 104 plots;
and crown in b, F1,21 = 10.3, p = 0.004, n = 24 plots) in needleleaf forests with fire expressed in terms of return period (crown fire plots are all 12 years
postfire, data subset to include short-interval burn plots). c, illustrates the mean response ratios ± standard error for the fire types with crown fires split
into high (>2,400 m) and low (<2,400) elevation sites (Crown 1 and Crown 2, respectively; n = 25 plots for each elevation category). Analyses were
robust to considering surface fires in only Western US needleleaf forests: F1, 47.1 = 17.2, p = 0.001. Response ratios were split into long and short fire return
interval plots (Crown 1 and 2, respectively), with the justification for definition of interval in 17.
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Extended Data Fig. 5 | Fire frequency effects across precipitation in the wet quarter bins. Partial residual plot displaying the relationship between loge
basal area and precipitation in the wettest quarter cross-sectioned based on fire frequency. This plot is based on the same mixed-effects model presented
in Fig. 3 and Supplementary Table 4, just re-arranged to emphasize how precipitation-basal area relationship changes with more frequent burning. n = 25
sites and n = 309 plots.
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Extended Data Fig. 6 | Differences in fire effects across continents. Partial residual plot between the length of time plots were exposed to frequent
burning and the log basal area (a, n = 25 sites and n = 309 plots) and stem density (b, n = 25 sites and n = 303 plots) in the different continents (from the
main mixed-effects model with site as a random intercept in Supplementary Tables 4-5).
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Extended Data Fig. 7 | Fire effects in sub-vegetation classifications. Partial residual plot between the length of time plots were exposed to frequent
burning and the log basal area in the different sub-vegetation types (from the main mixed-effects model, presented in Supplementary Table 4 but
substituting the broad vegetation effect with the more detailed classification. We found no evidence that accounting for the finer-scale variability in
ecosystem classification increased the accuracy of the model or changed our conclusions. n = 25 sites and n = 309 plots.
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Extended Data Fig. 8 | Responses of stem densities to fire across environmental covariates. Partial residual plots of the mixed-effects model for
stem densities illustrating how fire frequency effects changed according to wet-season precipitation, mean annual temperature, and ecosystem type.
Panels structured by standard deviations around the median to visualize the spread (−1, 0,1), PWQ: precipitation in the wet quarter, MAT: mean annual
temperature. All model fits are p < 0.05 and specific results can be found in Supplementary Table 5. The predictor variables are mean-centered and
standard deviations are scaled to facilitate comparisons of variable influence. In needleleaf and broadleaf forests, stem densities actually increased with
more frequent burning initially, but declined with increasing experiment duration, potentially because of increased light availability initially stimulating
recruitment of small trees (Extended Data Fig. 6, Supplementary Table 5). Stem density in African sites changed little through time (Extended Data Fig. 6).
The trends in density may reflect the ability of many of the tree species to re-sprout in between fire events78. n = 25 sites and n = 303 plots.
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Extended Data Fig. 9 | Effect of fire on phosphorus stoichiometry. Partial residual plots of the phosphorus (P) stoichiometry of community weighted
means as a function of years of repeated burning. Taken from mixed-effects models presented in Supplementary Table 7. The models include a vegetation
type effect. Tissue P is rescaled by subtracting the mean and dividing by the standard deviation. n = 16 sites and n = 172 plots.
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nature research | reporting summary April 2020
Corresponding author(s): Adam F. A. Pellegrini
Last updated by author(s): 12/14/2020
Reporting Summary
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Ecological, evolutionary & environmental sciences study design
All studies must disclose on these points even when the disclosure is negative.
Study description We quantified the effects of fire frequency on tree cover across broad biogeographic and climatic scales, incorporating additional
factors that may explain variability in fire effects on tree communities. We used original survey data collected across 29 repeated
burn experiments within a large collaborative network. We also compiled data on plant functional traits within North American fire
manipulation experiments to test how traits could predict the effects of fire as well as how fire changed the functional composition
of the plant community.
Research sample We combined published data with new surveys on tree populations from 29 sites and 374 plots in four continents. At 27 of the sites
(324 plots), surface fire frequency was experimentally manipulated for 16-64 years (mean of 30 years), and at two sites (50 plots),
natural variation in crown fire frequency presented a natural experiment. We focused on tree responses because of their importance
for carbon storage, ecosystem productivity and nutrient cycling23,27,28. The sites cover locations in North America (coniferous,
deciduous, and mixed forests, broadleaf savanna), South America (tropical savanna and temperate shrubland), Africa (broadleaf
savanna), and Australia (wet and dry eucalyptus forests) all of which are ecosystems that experience frequent burning (Figure S1,
Table S1, Supplemental Information, SI). For our main analysis, we define ecosystems broadly based on dominant fuel type (grass vs.
litter) and tree functional composition (angiosperm broadleaf vs. gymnosperm needleleaf trees) but test the robustness of our
conclusions to other classifications. Each surface fire site contains replicate plots including an unburned treatment and different
prescribed burning frequencies (Figure S2), where fire frequencies ranged from approximately one fire every decade to one fire
every year (Table S1). We evaluated the effects of fire alone and in combination with environmental covariates using model selection
(SI) to test the importance of climate, soil, and species composition in modifying the effect of fire.
Sampling strategy Our study contains two levels of relevant sampling strategies. First, the level of choosing the sites included in our analysis. This paper
was a result of a large collaborative group that has communicated through a series of meetings with the 1st author Adam Pellegrini.
This study is not an exhaustive meta-analysis of the literature. Clearly, not every single fire manipulation experiment in the world is
included in this study, partly because in some cases, organizers of fire manipulation experiments declined to be a part of the study. In
other cases, organizers never responded or did not have raw data necessary for our analysis.
The second level of sampling included collecting data on the tree communities and soils. Transects to survey trees and collect soils
were randomly established in the plots. Survey area was generally designed to capture the ecological heterogeneity within a
landscape (e.g., larger surveys in areas with patchy tree cover such as savannas). All functional trait data were acquired from the
most extensive trait databases currently freely accessible.
Data collection We provide information on the experimental designs in Table S1 because each site varied (we also include the corresponding citation
that describes each experiment in greater detail). This table includes a list of sites with key meta-data such as the continent,
vegetation type present in broad categories and the families of the dominant tree species. Number of plots is the total within the
entire site. Duration is the number of years over which fire frequencies have differed across plots. We also describe the prior
conditions of the ecosystem at the beginning of the experiment (e.g. whether the site experienced regular burning prior to the
experiment).
Soil data were collected using previously published values, unpublished values, or our own sampling and analyses. The types of data
for each site are described in the section "Soil chemistry data".
Plant functional trait data were compiled from published datasets, with the appropriate citations given in the text. The section
describing these data is "Species classifications and functional traits".
Timing and spatial scale Sampling was conducted over the course of several decades because the timing that the fire manipulation experiments were
surveyed differed across sites. The plots were relatively large (>0.5 hectares, and in some cases up to 7 hectares in size). The tree
community surveys attempted to account for ecological heterogeneity within the plots.
Data exclusions In certain cases, experiments contained multiple surveys through time and treatments in different seasons. These are described in
the section "Choice of plots within 27 sites with surface fires" in our Supplemental Information section. Our decision to exclude plots
was conservative because we wanted to avoid pseudoreplication arising from repeated measures on the same fire treatment in
different years within a site. We avoided using different seasons of burns within the same site because the range of fire conditions in
the seasons varied considerably in some sites (e.g., burned anytime between March - June).
Reproducibility These experiments are large-scale fire manipulations at the ecosystem level for multiple decades. The reproducibility of the
experimental findings from a single study is partly captured by comparing the effects of fire from multiple independent experiments.
Randomization Replicate plots within each experiment were randomized at the onset of the experiment.
Blinding Blinding was not consistently used during data collection but we did randomize the location of the transects within a plot.
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nature research | reporting summary April 2020
Did the study involve field work? Yes No
Field work, collection and transport
Field conditions Fieldwork was generally conducted in the growing season, which allowed for the proper identification of tree species. Weather
conditions during the field season varied considerably across sites given their broad range in climates.
Location All experimental locations are provided in Table S1. We provide a broad description of the sites as well as the original references that
describe the experimental design of each site.
Access & import/export We did not conduct any importing or exporting of samples requiring permits.
Disturbance There was no disturbance of the sites. Soil cores are minimal and not considered destructive.
Reporting for specific materials, systems and methods
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Antibodies
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Palaeontology and archaeology
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Methods
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ChIP-seq
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