Effects of Pre-Fire Vegetation on the Post-Fire Plant Community Response to Wildfire along a Successional Gradient in Western Juniper Woodlands

Abstract

Western juniper was often historically restricted to fire refugia such as rocky outcrops but has since Euro-American settlement expanded into areas previously dominated by sagebrush steppe. Wildfires in developed woodlands have been rare. In 2007, the Tongue-Crutcher Wildland Fire burned 18,890 ha in southwestern Idaho along a woodland development gradient, providing unique research opportunities. To assess fire effects on vascular plants, field data were collected in 2012/2013 and 2019/2020. Species richness was uniform along the sere, while species diversity declined in late woodland stages attributed to juniper dominance. The greatest changes in species composition following fire occurred in later woodland development phases. Herbaceous vegetation increased following fire, but sagebrush cover was still lower in burned plots 12–13 years post-fire. Many stands dominated by juniper pre-fire became dominated by snowbrush ceanothus post-fire. Juniper seedlings were observed post-fire, indicating that juniper will reoccupy the area. Our research demonstrates resilience to fire and resistance to annual grasses particularly in early successional stages, which provides opportunities for fire use as a management tool on cool and moist ecological sites. Loss of old-growth juniper to wildfire underlines the importance of maintaining and provisioning for future development of some old growth on the landscape given century-long recovery times.

Full text

Citation: Strand, E.K.; Bunting, S.C.
Effects of Pre-Fire Vegetation on the
Post-Fire Plant Community Response
to Wildfire along a Successional
Gradient in Western Juniper
Woodlands. Fire 2023,6, 141.
https://doi.org/10.3390/fire6040141
Academic Editor: Grant Williamson
Received: 6 February 2023
Revised: 28 March 2023
Accepted: 28 March 2023
Published: 2 April 2023
Copyright: © 2023 by the authors.
Licensee MDPI, Basel, Switzerland.
This article is an open access article
distributed under the terms and
conditions of the Creative Commons
Attribution (CC BY) license (https://
creativecommons.org/licenses/by/
4.0/).
fire
Article
Effects of Pre-Fire Vegetation on the Post-Fire Plant Community
Response to Wildfire along a Successional Gradient in Western
Juniper Woodlands
Eva K. Strand * and Stephen C. Bunting
Department of Forest, Rangeland, and Fire Sciences, University of Idaho, Perimeter Drive 875 MS 1135,
Moscow, ID 83844-1135, USA
*Correspondence: evas@uidaho.edu; Tel.: +1-208-885-5779
Abstract:
Western juniper was often historically restricted to fire refugia such as rocky outcrops but
has since Euro-American settlement expanded into areas previously dominated by sagebrush steppe.
Wildfires in developed woodlands have been rare. In 2007, the Tongue-Crutcher Wildland Fire
burned 18,890 ha in southwestern Idaho along a woodland development gradient, providing unique
research opportunities. To assess fire effects on vascular plants, field data were collected in 2012/2013
and 2019/2020. Species richness was uniform along the sere, while species diversity declined in
late woodland stages attributed to juniper dominance. The greatest changes in species composition
following fire occurred in later woodland development phases. Herbaceous vegetation increased
following fire, but sagebrush cover was still lower in burned plots 12–13 years post-fire. Many stands
dominated by juniper pre-fire became dominated by snowbrush ceanothus post-fire. Juniper seedlings
were observed post-fire, indicating that juniper will reoccupy the area. Our research demonstrates
resilience to fire and resistance to annual grasses particularly in early successional stages, which
provides opportunities for fire use as a management tool on cool and moist ecological sites. Loss of
old-growth juniper to wildfire underlines the importance of maintaining and provisioning for future
development of some old growth on the landscape given century-long recovery times.
Keywords:
mountain big sagebrush; burn severity; diversity; species turnover; secondary succession
1. Introduction
1.1. Extent of Sagebrush Steppe and Juniper Woodlands in the Great Basin
Pinyon–juniper woodlands are a dominant vegetation type throughout the semi-arid
mountainous regions of the Great Basin, Columbia Basin and Colorado Plateau. Miller
and Tausch [
1
] estimated that about 30 million ha are occupied by these woodlands in the
western US. Before Euro-American settlement, it has been estimated that pinyon–juniper
woodlands occupied less than 3 million ha [
2
]. Rapid expansion of the woodlands and
infilling in established open woodlands [
3
] has occurred in the Euro-American period after
1800. The northwestern woodlands are dominated by western juniper (Juniperus occidentalis
Hook var. occidentalis) and occupy 3.6 million ha, primarily in central and eastern Oregon,
southwestern Idaho and northeastern California [
1
,
4
]. Causes of this expansion have been
attributed to a number of factors including active and passive fire suppression, historic
livestock grazing, fuel fragmentation (by roads, agricultural lands, etc.), climatic variation
and change, and elevated CO
2
levels [
4
,
5
]. Following severe drought in recent years,
increased mortality in pinyon pine (Pinus monophylla (Torr. and Frém.) and Utah juniper
(Juniperus osteosperma (Torr.) Little) woodlands has been documented on warm dry sites in
central Nevada [6]; however, diebacks have not been reported for western juniper.
Fire 2023,6, 141. https://doi.org/10.3390/fire6040141 https://www.mdpi.com/journal/fire

Fire 2023,6, 141 2 of 25
1.2. Woodland Development Phases, Mature Juniper and Juniper Expansion
The successional transition of sagebrush shrub steppe into mature western juniper
woodlands has been described in three phases [
4
]. In Phase I, the juniper trees are small,
widely spaced and generally have only minor influences on ecological processes. Herba-
ceous and shrub components remain relatively intact and are minimally affected by the few
and young juniper trees. In Phase II, the trees and understory shrub and herbaceous layers
are co-dominant. The juniper trees are beginning to influence the site’s composition and
ecological processes. In Phase III, juniper dominates the site’s composition and ecological
processes. Many shrub steppe species have declined on the site and will continue to decline
as the woodland develops. Mature woodland is dominated by juniper and sometimes
pine species. Understory shrub and herbaceous species coverage is diminished from the
shrub-steppe-dominated state due to competition with trees for light and other resources.
Bare ground in the tree interspaces is common. The tree component often has a mixed age
and size structure, but a majority of the site’s plant coverage consists of mature trees. Total
tree cover varies by region and soil type but usually exceeds 30%. The age of these mature
trees often exceeds 400 years, and some individuals may be over 1000 years old. Land-
scapes are frequently a mosaic of different phases, mature western juniper and sagebrush
steppe (Figure 1).
Fire 2023, 6, x FOR PEER REVIEW 2 of 25
has been documented on warm dry sites in central Nevada [6]; however, diebacks have
not been reported for western juniper.
1.2. Woodland Development Phases, Mature Juniper and Juniper Expansion
The successional transition of sagebrush shrub steppe into mature western juniper
woodlands has been described in three phases [4]. In Phase I, the juniper trees are small,
widely spaced and generally have only minor influences on ecological processes. Herba-
ceous and shrub components remain relatively intact and are minimally affected by the
few and young juniper trees. In Phase II, the trees and understory shrub and herbaceous
layers are co-dominant. The juniper trees are beginning to influence the site’s composi-
tion and ecological processes. In Phase III, juniper dominates the site’s composition and
ecological processes. Many shrub steppe species have declined on the site and will con-
tinue to decline as the woodland develops. Mature woodland is dominated by juniper
and sometimes pine species. Understory shrub and herbaceous species coverage is di-
minished from the shrub-steppe-dominated state due to competition with trees for light
and other resources. Bare ground in the tree interspaces is common. The tree component
often has a mixed age and size structure, but a majority of the site’s plant coverage con-
sists of mature trees. Total tree cover varies by region and soil type but usually exceeds
30%. The age of these mature trees often exceeds 400 years, and some individuals may
be over 1000 years old. Landscapes are frequently a mosaic of different phases, mature
western juniper and sagebrush steppe (Figure 1).
Figure 1. Overview of sagebrush steppe and western juniper woodland landscape in the Owyhee
Mountains of southwestern Idaho.
In the last couple of decades, juniper removal treatments have been implemented
with the goal of restoring sagebrush steppe vegetation and habitats [4] and reversing the
ecological effects of juniper expansion. Prescribed fire or a variety of mechanical treat-
ments, or a combination of the two, are common in early successional stages of juniper
encroachment, but sagebrush steppe restoration treatments have also been implemented
in Phase III woodlands [7]. Pre-settlement old woodlands are unique habitats and not a
target for juniper removal. Old-growth juniper is estimated to make up less than 10% of
the western juniper woodlands [8].
Figure 1.
Overview of sagebrush steppe and western juniper woodland landscape in the Owyhee
Mountains of southwestern Idaho.
In the last couple of decades, juniper removal treatments have been implemented
with the goal of restoring sagebrush steppe vegetation and habitats [
4
] and reversing
the ecological effects of juniper expansion. Prescribed fire or a variety of mechanical
treatments, or a combination of the two, are common in early successional stages of juniper
encroachment, but sagebrush steppe restoration treatments have also been implemented
in Phase III woodlands [
7
]. Pre-settlement old woodlands are unique habitats and not a
target for juniper removal. Old-growth juniper is estimated to make up less than 10% of
the western juniper woodlands [8].
1.3. Fire Regimes and Changes in Fuel Components
Fire regimes describe spatial and temporal characteristics of wildfire in an ecosystem,
for example fire frequency, burn severity, fire size and patchiness. Historical fire regimes
provide context for the role of fire in an ecosystem and can help explain changes that are
occurring within plant communities and landscapes as a result of deviations from historical
fire regimes. Fire frequency in mountain big sagebrush was historically higher than in

Fire 2023,6, 141 3 of 25
other sagebrush communities because of higher productivity and more continuous fine
surface fuels [
9
]. Fire frequency likely varied along moisture and temperature gradients and
depending on the fire ignition and spread probability in adjacent cover types [
10
]. Reported
historical fire intervals in mountain big sagebrush (Artemisia tridentata ssp. vaseyana [Rydb.]
Beetle) are variable depending on region and adjacency. Miller and Tausch [
1
] estimated
fire frequency in mesic mountain big sagebrush communities at 11 to 25 years based on
analysis of adjacent fire scarred trees. Fire return intervals in mountain big sagebrush were
suggested to be greater than 80 years in the mountain big sagebrush–bluebunch wheatgrass
(Pseudoroegneria spicata [Pursh] A. Löve)–Thurber needlegrass (
Acnatherum thurberianum
[Piper] Barkworth) plant association with lower fuel loads and associated with large
western juniper snags [
11
]. Burkhardt and Tisdale [
12
] suggested that mountain big
sagebrush currently encroached by western juniper must historically have had a fire return
interval < 50 years because western juniper trees less than 50 years of age are readily killed
by fire, which would inhibit woodland expansion. Longer fire return intervals are suggested
in the western juniper–mountain big sagebrush–western needlegrass plant associations
with limited fine fuel production; presence of old western juniper trees indicated long
fire-free periods [
10
]. Fires in mountain big sagebrush were likely small (<500 ha); large
fires were infrequent [
13
]. In a systematic review, Baker and Shinneman [
14
] evaluated
fire regime characteristics and the role of fire in pinyon–juniper woodlands. They found
that most fires in pinyon–juniper woodlands burned at a high-severity infrequent fire
regime and that low-severity surface fire was rare in developed woodlands because of
low herbaceous and shrub fuel loads. Rather than a spreading surface fire, high-severity
fires have been described as trees torching and ejecting large flaming embers composed of
ropelike juniper bark strips, igniting neighboring trees and vegetation [
15
]. We assume that
the pinyon–juniper woodlands described by Baker and Shinneman [
14
] are late successional
of the Phase III or Mature woodland classification proposed by Miller et al. [4].
The fuel complex dramatically changes along the successional gradient from sagebrush
steppe to mature juniper woodland. Fuel loading is primarily composed of herbaceous
and small-diameter woody fuels (<25.4 mm) in sagebrush steppe and Phase I woodlands.
As juniper becomes increasingly dominant on the site, the herbaceous fuel component
declines [
4
,
16
]. The shrub component of the vegetation decreases, and given time, the
woody shrub biomass transitions to surface fuels [
16
,
17
]. As the woodland matures, the fuel
complex becomes increasing dominated by those fuels produced by the juniper trees, i.e.,
large-diameter woody fuels [
16
] and juniper litter and duff [
18
]. Large fuels and compacted
litter and duff are less flammable but tend to smolder for longer periods of time, resulting in
injury to plant tissue and reproductive structures [
19
]. These changes, in turn, affect many
characteristics of the fires when they occur such as fire size and burn uniformity [
20
], flame
length and rates of spread [
16
]. Thus, fires along the maturing woodland development
gradient become less frequent and more severe when they occur [21,22].
Pre-fire plant composition and structure are important characteristics when evaluating
post-fire plant succession and resilience, and resistance to invasive plants [
10
]. Resilience is
defined as the capacity of ecosystems to reorganize and regain their structure, processes
and functioning following a disturbance [
23
]. Resistance refers to an ecosystem’s ability to
resist community structural change following disturbance [24,25].
Few studies have reported on plant community response to wildfire along the moun-
tain big sagebrush–western juniper woodland gradient. Increasing burn severity of wildfire
has been documented along the woodland development gradient supported by both remote
sensing and field reconnaissance [
22
]. A review of fire effects in mountain big sagebrush
documents general impacts. Deep-rooted perennial grasses decline after fire but generally
recover to pre-burn levels after 2–3 years, but recovery time depends on burn severity
and post-fire environmental conditions [
26
,
27
]. Sagebrush is sensitive to fire and can take
several years to decades to return to pre-fire cover levels [
28
]. Shrubs with the ability to
sprout after fire or shrubs with soil-stored seeds can more rapidly return post-fire lev-
els [
27
]. Western juniper trees under 1–2 m have thin bark and are readily killed by fire,

Fire 2023,6, 141 4 of 25
while older trees may survive low-intensity fire but are sensitive to crown scorch and killed
in crown fires [
29
]. Wildfire can have severe effects on soil properties. Soil erosion increased
by a factor of 20 following wildfire in western juniper woodlands and remained high in
areas previously occupied by trees but was reduced in the inter-canopy because of post-
fire herbaceous recruitment [
30
]. The 2007 Tongue-Crutcher Wildland Fire Complex that
burned 18,890 ha across the gradient of western juniper woodland development phases and
through mature juniper provided a unique opportunity to evaluate ecological relationships
and plant community response to fire in this ecosystem.
Several metrics are used by ecologists for evaluating ecological differences and simi-
larities between plant communities [
31
,
32
]. Species richness refers to the number of species
in a community, while diversity also accounts for the proportional abundance of those
species, for example, Shannon’s Diversity Index [
33
]. Whittaker [
34
] partitioned diversity
into alpha, beta and gamma components based on a spatial scale. Alpha diversity refers to
the local level diversity at stands within a community type. Beta diversity represents the
dissimilarity in species composition between sampling units and reflects the turnover or
replacement of species between community types [
32
,
34
,
35
]. Gamma diversity accounts
for the total number of species in the region under study. Biotic cover of live vegetation is
another measure commonly used to evaluate revegetation in post-fire environments [
36
,
37
].
The overarching research question we address in this paper is how does pre-fire
vegetation affect post-fire recovery in mountain big sagebrush steppe and western juniper
woodlands following wildfire? Specifically, we conduct analyses to address the following
questions: (1) What are the effects of wildfire on species diversity and species turnover
along the woodland successional gradient? (2) How does the plant community composition
change along the successional gradient and what are the effects of wildfire on the plant
community? (3) Are there differences in cover of species functional groups along the
successional gradient post-fire? (4) What are the dominant species in this plant community
and how is the cover of those species affected by wildfire? Finally, we synthesize our
results in a prediction of a long-term trajectory for stands that burn at different woodland
development phases and for mature woodlands.
2. Methods
2.1. Site Description
Western juniper woodlands are native to eastern Oregon, northern California, north-
western Nevada and southwestern Idaho (Figure 2a), occupying approximately 3.6 million
ha in the region. In 2007, the Tongue-Crutcher Wildland Fire Complex (TCWFC) burned
through 18,890 hectares of sagebrush steppe and western juniper woodlands on Juniper
Mountain, located in the southwestern part of the Owyhee Plateau in Idaho (42
◦
20
0
39
00
N;
116◦5003200 W).
Elevation within the study area ranged from 1300 m in the canyons to 2000 m towards
the top of Juniper Mountain. Average annual precipitation ranges from 270 mm at the
lowest elevation to 540 mm at the highest elevation within the fire perimeter [
38
]. The lowest
average monthly temperature of
−
3.2
◦
C occurs in December, and the highest average
monthly temperature of 21.5
◦
C occurs in July [
38
]. Parent material on Juniper Mountain
resulted from a basaltic eruption and is composed of ashflow tuff and ignimbrite [
39
], with
the dominant soil types being haplargids at lower elevations, haploxeralfs at mid-elevations
and argixerolls towards the very top of the mountain [
40
]. Sampling was conducted in
the northern part of the burn and adjacent unburned areas where a variety of woodland
development phases were present. The ecological site sampled is in the Fastjet soil series
classified as shrubby loam 330–406 mm precipitation, with dominant vegetation being big
sagebrush (Artemisia tridentata Nutt.), Idaho fescue (Festuca Idahoensis Elmer) and antelope
bitterbrush (Purshia tridentata [Pursh] DC.) [
40
]. The sampled area falls in the category
of ecological sites classified as high resilience to fire and resistance to invasion by annual
grasses [41].

Fire 2023,6, 141 5 of 25
Fire 2023, 6, x FOR PEER REVIEW 5 of 25
Figure 2. (a) Distribution of western juniper in the western United States. (b) Location of the 2007
Tongue-Crutcher Wildland Fire Complex on Juniper Mountain including burn severity classes
within the fire perimeter.
Elevation within the study area ranged from 1300 m in the canyons to 2000 m to-
wards the top of Juniper Mountain. Average annual precipitation ranges from 270 mm at
the lowest elevation to 540 mm at the highest elevation within the fire perimeter [38].
The lowest average monthly temperature of −3.2 °C occurs in December, and the highest
average monthly temperature of 21.5 °C occurs in July [38]. Parent material on Juniper
Mountain resulted from a basaltic eruption and is composed of ashflow tuff and ignim-
brite [39], with the dominant soil types being haplargids at lower elevations, haploxe-
ralfs at mid-elevations and argixerolls towards the very top of the mountain [40]. Sam-
pling was conducted in the northern part of the burn and adjacent unburned areas
where a variety of woodland development phases were present. The ecological site sam-
pled is in the Fastjet soil series classified as shrubby loam 330–406 mm precipitation,
with dominant vegetation being big sagebrush (Artemisia tridentata Nutt.), Idaho fescue
(Festuca Idahoensis Elmer) and antelope bitterbrush (Purshia tridentata [Pursh] DC.) [40].
The sampled area falls in the category of ecological sites classified as high resilience to
fire and resistance to invasion by annual grasses [41].
Pre-fire vegetation was characterized by a mosaic of sagebrush steppe and western
juniper (Juniperus occidentalis Hook. var. occidentalis) in various woodland development
stages ranging from stand initiation woodlands (Phase I) to mature woodlands with
trees upward of 500 years. The dominant sagebrush species within the sampled area was
mountain big sagebrush (Artemisia tridentata ssp. vaseyana [Rydb.] Beetle) on deeper soils
and little sagebrush (Artemisia arbuscula Nutt.) on shallow clay soils. Other xeric shrubs
included rabbitbrush (Chrysothamnus Nutt. spp.) and antelope bitterbrush (Purshia tri-
dentata [Pursh] DC). Curl-leaf mountain-mahogany (Cercocarpus ledifolius Nutt.) was
widespread within the area, ranging in size from small shrubs to tall tree-like shrubs
(>10 m tall). Mountain shrub species included mountain snowberry (Symphoricarpos ore-
ophilus A. Gray), bittercherry (Prunus emarginata [Douglas ex Hook.] D. Dietr.) and
chokecherry (Prunus virginiana L.). Snowbrush ceanothus (Ceanothus velutinus Douglas
ex Hook.) was not common prior to the burn but present in a few openings. Native per-
ennial grass species included bluebunch wheatgrass, Idaho fescue, needlegrasses
(Achnatherum P. Beauv. spp.), oniongrass (Melica bulbosa Geyer ex Porter and J.M. Coult.)
and Sandberg bluegrass (Poa secunda J. Presl). Native perennial forbs included arrowleaf
balsamroot (Balsamorhiza sagittata [Pursh] Nutt.), lupine (Lupinus L. spp.), tapertip
hawksbeard (Crepis accuminata Nutt.), desert parsley (Lomatium Raf. spp.) and buck-
wheat (Eriogonum Michx. spp.). Annual forbs were frequent but present at low cover
levels because of their delicate stature, for example autumn willow-herb (Epilobium
brachycarpum C. Presl), cryptantha (Cryptantha Lehm. ex G. Don spp.), blue-eyed Mary
Figure 2.
(
a
) Distribution of western juniper in the western United States. (
b
) Location of the 2007
Tongue-Crutcher Wildland Fire Complex on Juniper Mountain including burn severity classes within
the fire perimeter.
Pre-fire vegetation was characterized by a mosaic of sagebrush steppe and western ju-
niper (Juniperus occidentalis Hook. var. occidentalis) in various woodland development stages
ranging from stand initiation woodlands (Phase I) to mature woodlands with trees upward
of 500 years. The dominant sagebrush species within the sampled area was mountain
big sagebrush (Artemisia tridentata ssp. vaseyana [Rydb.] Beetle) on deeper soils and little
sagebrush (Artemisia arbuscula Nutt.) on shallow clay soils. Other xeric shrubs included
rabbitbrush (Chrysothamnus Nutt. spp.) and antelope bitterbrush (
Purshia tridentata
[Pursh]
DC). Curl-leaf mountain-mahogany (Cercocarpus ledifolius Nutt.) was widespread within the
area, ranging in size from small shrubs to tall tree-like shrubs (>10 m tall). Mountain shrub
species included mountain snowberry (Symphoricarpos oreophilus A. Gray), bittercherry
(Prunus emarginata [Douglas ex Hook.] D. Dietr.) and chokecherry (
Prunus virginiana L.
).
Snowbrush ceanothus (Ceanothus velutinus Douglas ex Hook.) was not common prior to
the burn but present in a few openings. Native perennial grass species included blue-
bunch wheatgrass, Idaho fescue, needlegrasses (Achnatherum P. Beauv. spp.), oniongrass
(Melica bulbosa Geyer ex Porter and J.M. Coult.) and Sandberg bluegrass (Poa secunda J.
Presl). Native perennial forbs included arrowleaf balsamroot (Balsamorhiza sagittata [Pursh]
Nutt.), lupine (Lupinus L. spp.), tapertip hawksbeard (Crepis accuminata Nutt.), desert
parsley (Lomatium Raf. spp.) and buckwheat (Eriogonum Michx. spp.). Annual forbs
were frequent but present at low cover levels because of their delicate stature, for example
autumn willow-herb (Epilobium brachycarpum C. Presl), cryptantha (Cryptantha Lehm. ex
G. Don spp.), blue-eyed Mary (Collinsia parviflora Lindl.), starwort (
Stellaria L spp.
), Dou-
glas’ knotweed (Polygonum douglasii Greene) and tiny trumpet (Collomia linearis Nutt.).
Non-native forbs were rare except yellow salsify (Tragopogon dubius Scop.), which was
common. Exotic annual grasses including cheatgrass (Bromus tectorum L.) and soft brome
(Bromus hordeaceus L.) were present at low cover throughout the area.
The TCWFC started by lightning in two locations along the Owyhee River on July 6,
following an unusually dry spring. Nearby weather stations in Rome, Oregon, and Murphy,
Idaho, reported wind gusts of up to 22 ms
−1
, temperatures over 38
◦
C and daytime relative
humidity below 10%. The high-intensity fire burned through sagebrush steppe vegetation
and juniper woodlands in a variety of woodland development stages (Phase I–III and
Mature), initially spreading northward on Juniper Mountain, then spreading towards the
east and west, exhibiting extreme fire behavior including spotting, torching and crowning.
The fire partially burned an area named the Big Tree Mesa, an area characterized by open-
canopy old pre-settlement woodlands surrounded by steep canyons both to the east and

Fire 2023,6, 141 6 of 25
west. These trees were likely some of the oldest western juniper in the region, implying the
rarity of the conditions under which this wildfire burned. Suppression action began around
11 July with dozer lines implemented on the northern edge; the fire was mostly contained
by 21 July. The burn severity index (delta Normalized Burn Ratio [
21
]) varied within
the fire perimeter, with the highest burn severity index associated with late successional
woodlands in the northern portion of the burned area (Figure 2b; [22]).
The remote land area is federally managed by the Bureau of Land Management (BLM)
with a few parcels of private and Idaho State land. The study area is primarily located in
two BLM grazing allotments. The Trout Springs allotment was largely unburned, while
the Castlehead Lambert allotment’s Between the Canyons pasture was mostly burned but
contained unburned sections within the fire perimeter. Prior to the fire, target utilization
was less than 50%. Within burned areas, the goal was 80% ground cover before continuing
grazing. None of the burned pastures were grazed by cattle in 2008 or 2009, but grazing
resumed in 2010 in the Trout Springs allotment. In the Between the Canyons pasture, where
the bulk of the burned plots occurred, no livestock grazing occurred in 2010. Aside from
cattle grazing in the late summer or fall, land use activities include hunting elk and deer,
and recreation such as camping, hiking and all-terrain vehicle (ATV) use.
2.2. Field Sampling
Post-fire vegetation sampling to assess vascular plant community composition was
conducted in 2012/2013 and again in 2019/2020. In 2012 and 2013 (5–6 years post-fire),
56 vegetation sampling plots were established, stratified by four woodland phases (Phase
I–III and Mature) in areas that burned in the TCWFC (36 plots referred to as “burned”) and
in adjacent areas that did not burn in the fire (20 plots referred to as “unburned”). The same
plots were resampled in 2019 and 2020 (12 and 13 years post-fire). In the following, we will
refer to these sampling efforts as sampling period 1 (2012/2013) and sampling period 2
(2019/2020). At each plot location, two 25 m transects were established at random locations
along a 20 m baseline. Percent canopy cover by species of all herbaceous vascular plants
was recorded at each meter-mark within a 0.5
×
0.5 m quadrat, resulting in 50 quadrats
total per plot. Percent rock was recorded within the quadrat considering the minimum
size of a rock to be 5
×
5 cm. Plant species with less than 1% cover were referred to as
“trace”, and in later calculations, we assigned the value 0.1% to these species with low cover.
Many of the species with low cover were, however, very frequent in our plots (present in
many quadrats), and thus the sum cover for a plot may be exaggerated. The number of
post-fire juniper seedlings was documented within the quadrats. Shrub cover by species
was estimated along the same transects using the line intercept method. Ocular cover
estimates were made for the remaining post-fire juniper (if any) for the plot as a whole. The
location of each plot was recorded using a Garmin 76CSx Global Positioning System unit.
Slope, aspect and elevation at the center of the baseline were recorded.
2.3. Statistical Analysis Methods
2.3.1. Species Richness and Diversity
Plant species data were summarized to the plot level at each location by averaging the
plant cover in each of the 50 quadrats along the transects. Species richness was computed
by adding all species of vascular plants present within the 50 quadrats within the plot plus
the shrub species documented along the line intercept and juniper if present. We used
Shannon’s Diversity Index (H’) as a measure of alpha diversity, the diversity at the plot level.
Shannon’s Diversity Index was calculated at the plot level using the following equation:
H0=−∑R
ipiln(pi)(1)
where p
i
is the proportion of the ith species, and R is the species richness, the maximum
number of species within a plot.

Fire 2023,6, 141 7 of 25
We tested for differences in species richness and H’ between sampling period (sam-
pling period 1 vs. 2), burn status (burned vs. unburned) and pre-fire woodland develop-
ment phase (Phase I, II, III and Mature) and their interactions with a repeated measures
Analysis of Variance (ANOVA). Prior to analysis, the variables were evaluated for normality
using Q–Q plots and histograms. Pairwise comparisons were performed using Tukey’s
Honestly-Significant-Difference Test. All comparisons were conducted between plots of the
same pre-burn woodland development phase.
To better understand the contribution of each woodland development stage to species
diversity within the landscape, we also compared species richness and diversity between
woodland development phases and the total landscape. For this analysis, we randomly
selected five plots from each woodland development phase (unburned and burned) to
avoid an uneven sample size since species richness is dependent on area sampled.
2.3.2. Plant Community Turnover and Composition
To assess differences in species turnover (a measure of beta diversity) between wood-
land phases along the successional gradient and to quantify the species turnover that
occurred as a result of the fire, we computed the Sorensen (Bray–Curtis) distance measure
of dissimilarity based on group averages [
31
]. Statistical differences in plant community
composition were evaluated between woodland development phases for burned and
unburned plots for the two sampling periods using the Multi-Response Permutation Proce-
dure (MRPP) with pairwise comparisons [
31
]. The MRPP analysis generates two statistics,
the p- and the A-value. The p-value represents the probability of a type I error under the
null hypothesis that there is no difference between samples. The A-value is a measure of
agreement between groups where A = 1 for complete within-group homogeneity,
A=0
when the heterogeneity within groups is equal to the expectation and A < 0 if there is less
agreement within groups than expected by chance. Ecological communities are commonly
A < 0.1, and within-group agreements with A > 0.3 are uncommon [
31
]. In the plant
community analysis, we included those species with more than three non-zero values to
reduce noise in the dataset introduced by infrequent species [31].
2.3.3. Functional Groups and Dominant Species
The vascular plant species documented in the plots were grouped in plant functional
groups, and canopy cover by functional groups was calculated by adding the cover of the
plant species within each group: annual grass, annual forbs, perennial grass, perennial
forbs, shrubs and trees. Differences in functional group canopy cover between burned and
unburned plots by woodland development phase and sampling period were tested with
Student’s t-test. We excluded annual forbs from this analysis because of the difficulty in
estimating cover of species with <1% cover. Instead, we calculated and report frequency
of frequent annual forbs. Frequency was calculated as the percentage of the quadrats that
contained the plant.
The most common plant species in the dataset (>3% canopy cover across plots) were
further evaluated. We tested for differences by woodland development phase by burn status
(burned vs. unburned) between years and for the difference between burned and unburned
plots by phase within the same year using Analysis of Variance. Pairwise differences
between woodland development phases were evaluated with a post-hoc Tukey’s Honestly-
Significant-Difference test.
Juniper seedlings established post-fire were counted within the 50 quadrats along the
transects and converted to seedling density (seedlings/m
2
). Difference in mean seedling
density between burned and unburned plots was tested with Student’s t-test for each
woodland development phase and for mature woodlands. We used SYSTAT [
42
] for statis-
tical analysis and PCORD [
43
] for plant community analysis. Differences were considered
significant at p< 0.05.

Fire 2023,6, 141 8 of 25
3. Results
3.1. Richness and Diversity
Differences in species richness and Shannon’s Diversity Index (H’) for sampling
periods (TIME), the pre-fire woodland development phase (PHASE) and burn status
(unburned vs. burned; BURN) were evaluated in a repeated-measures three-way ANOVA.
Differences in species richness were observed by sampling period and pre-fire woodland
development phase, while there was no difference in richness between unburned and
burned stands (Table 1). None of the interaction terms were significant. For H’, all variables
(TIME, PHASE and BURN) were significant, while none of the interactions were (Table 1).
Table 1.
Evaluation of changes in species richness per plot and Shannon’s Diversity Index by sampling
period (TIME; repeated measure), pre-fire woodland development phase (PHASE) and fire (BURN)
and their interactions were conducted with a three-way repeated-measures ANOVA (df—degrees of
freedom; MS—mean squares; F-ratio and p-value). Significant relationships (p< 0.05) are in bold font.
Variable Richness Shannon’s Diversity Index
df MS F-Ratio pdf MS F-Ratio p
TIME 1 611.02 24.34 <0.001 1 4.73 31.06 <0.001
TIME*BURN 1 26.27 1.05 0.311 1 0.23 1.50 0.226
TIME*PHASE 3 20.07 0.80 0.500 3 0.10 0.64 0.593
TIME*BURN*PHASE
3 48.18 1.92 0.139 3 0.06 0.38 0.766
BURN 1 30.46 0.80 0.375 1 3.66 14.22 <0.001
PHASE 3 217.36 5.72 0.002 3 8.82 34.21 <0.001
BURN*PHASE 3 6.61 0.17 0.914 3 0.15 0.57 0.635
Pairwise comparison analysis confirmed differences in species richness between early
and late woodland development phases for sampling period 1, with decreasing richness
along the woodland development gradient. The trend was similar for unburned and burned
stands (Figure 3a). In the second sampling period, there was no difference in richness
between the woodland development phases for unburned or burned stands (Figure 3b).
No difference was observed in H’ between unburned and burned woodland development
phases in either sampling period, but H’ was lower for the late-development stands (Phase
III and Mature) in both sampling periods (Figure 3c,d).
Differences observed between sampling periods included increases in species richness
for burned Phase I (p= 0.008), burned Phase III (p= 0.001) and burned Mature stands
(
p< 0.001
). Increases in H’ were observed for unburned Phase I (p= 0.014) and Phase III
(p= 0.024) and burned Phase III (p= 0.016).
Total species richness within woodland development phases, the sum of all species
within the woodland development phase rather than the average as displayed in Figure 3,
was relatively constant along the sere for both sampling periods (Figure 4a,b). Little
difference in richness was observed because of the burn. However, the total number of
species sampled across the five randomly selected unburned and burned plots by phase
was 124 during the first and 131 in the second sampling period. Diversity (H’), when
assessed for all species within the woodland development phases, similar to the plot
averages presented in Figure 3, was highest in the early woodland development phase
and decreased along the sere. H’ was higher in burned plots when compared to unburned
plots of the same phase. H’ was higher in the early woodland development phases. At the
landscape scale, H’ was higher in the burned plots due to the reduction in the dominance
of juniper, which increased equitability between species.

Fire 2023,6, 141 9 of 25
Fire 2023, 6, x FOR PEER REVIEW 9 of 25
and burned stands (Figure 3a). In the second sampling period, there was no difference in
richness between the woodland development phases for unburned or burned stands
(Figure 3b). No difference was observed in H’ between unburned and burned woodland
development phases in either sampling period, but H’ was lower for the late-
development stands (Phase III and Mature) in both sampling periods (Figure 3c,d).
Figure 3. Species richness by woodland development phase in unburned and burned stands 5–6
years post-fire (a) and 12–13 years post-fire (b). Shannon’s Diversity Index (H’) by woodland de-
velopment phase in unburned and burned plots 5–6 years post-fire (c) and 12–13 years post-fire
(d). Error bars represent standard deviation between plots within the same woodland develop-
ment phase. Capital letters A and B indicates statistical difference (p < 0.05) between phases for
unburned plots and lower-case a and b indicates difference between phases in burned plots. Bars
with the same letter combination indicates there is no difference.
Differences observed between sampling periods included increases in species rich-
ness for burned Phase I (p = 0.008), burned Phase III (p = 0.001) and burned Mature
stands (p < 0.001). Increases in H’ were observed for unburned Phase I (p = 0.014) and
Phase III (p = 0.024) and burned Phase III (p = 0.016).
Total species richness within woodland development phases, the sum of all species
within the woodland development phase rather than the average as displayed in Figure
3, was relatively constant along the sere for both sampling periods (Figure 4a,b). Little
Figure 3.
Species richness by woodland development phase in unburned and burned stands 5–6 years
post-fire (
a
) and 12–13 years post-fire (
b
). Shannon’s Diversity Index (H’) by woodland development
phase in unburned and burned plots 5–6 years post-fire (
c
) and 12–13 years post-fire (
d
). Error bars
represent standard deviation between plots within the same woodland development phase. Capital
letters A and B indicates statistical difference (p< 0.05) between phases for unburned plots and
lower-case a and b indicates difference between phases in burned plots. Bars with the same letter
combination indicates there is no difference.

Fire 2023,6, 141 10 of 25
Fire 2023, 6, x FOR PEER REVIEW 10 of 25
difference in richness was observed because of the burn. However, the total number of
species sampled across the five randomly selected unburned and burned plots by phase
was 124 during the first and 131 in the second sampling period. Diversity (H’), when as-
sessed for all species within the woodland development phases, similar to the plot aver-
ages presented in Figure 3, was highest in the early woodland development phase and
decreased along the sere. H’ was higher in burned plots when compared to unburned
plots of the same phase. H’ was higher in the early woodland development phases. At
the landscape scale, H’ was higher in the burned plots due to the reduction in the domi-
nance of juniper, which increased equitability between species.
Figure 4. Total species richness (a,b) and diversity (c,d) for each woodland development stage
(unburned and burned) and for the total landscape for the two sampling periods.
3.2. Plant Community Turnover and Composition
Differences in plant community composition were assessed using the Sorensen
(Bray–Curtis) Dissimilarity Index (SDI) based on group averages [31]. The SDI ranges
from zero if there is no difference in plant community composition (richness and abun-
dance) to one if complete species turnover has occurred. Species turnover measured by
the SDI was 0.414 between unburned Phase I and II, increased to 0.586 between Phase II
and III, and was lower (0.075) between Phase II and Mature woodlands (Figure 5, top
row). Turnover in species composition occurring as a result of fire increased along the
successional gradient. The difference in the SDI between unburned and burned plots of
Phase I was 0.419, while the SDI for Phase II, Phase III and Mature was 0.643, 0.745 and
0.945, respectively, when comparing unburned and burned plots 5–6 years after the fire
(Figure 5). The species turnover between sampling periods (Time 1 and 2) was similar
along the successional gradient, ranging from 0.407 to 0.490 (Figure 5).
Figure 4.
Total species richness (
a
,
b
) and diversity (
c
,
d
) for each woodland development stage
(unburned and burned) and for the total landscape for the two sampling periods.
3.2. Plant Community Turnover and Composition
Differences in plant community composition were assessed using the Sorensen (
Bray–Curtis
)
Dissimilarity Index (SDI) based on group averages [
31
]. The SDI ranges from zero if
there is no difference in plant community composition (richness and abundance) to one if
complete species turnover has occurred. Species turnover measured by the SDI was 0.414
between unburned Phase I and II, increased to 0.586 between Phase II and III, and was
lower (0.075) between Phase II and Mature woodlands (Figure 5, top row). Turnover in
species composition occurring as a result of fire increased along the successional gradient.
The difference in the SDI between unburned and burned plots of Phase I was 0.419, while
the SDI for Phase II, Phase III and Mature was 0.643, 0.745 and 0.945, respectively, when
comparing unburned and burned plots 5–6 years after the fire (Figure 5). The species
turnover between sampling periods (Time 1 and 2) was similar along the successional
gradient, ranging from 0.407 to 0.490 (Figure 5).

Fire 2023,6, 141 11 of 25
Fire 2023, 6, x FOR PEER REVIEW 11 of 25
Figure 5. Average of Shannon’s Diversity Index (H’) for plots by woodland successional phase pre-
fire (blue circles), 5–6 years after the wildfire (Time 1, orange circles middle row) and 12–13 years
after the wildfire (Time 2, yellow circles bottom row). Sorensen’s Dissimilarity Index (SDI) is a
measure of dissimilarity or species turnover along the successional gradient and following the
wildfire. High values represent high turnover of species. The largest observed change along the
woodland development gradient is seen between Phase II and III. Species turnover because of the
fire increased along the successional gradient with almost a complete replacement of the plant
community in the Mature woodlands (SDI = 0.945).
Differences in plant community composition were evaluated with the Multi-
Response Permutation Procedure [31]. Overall differences between woodland develop-
ment phases were observed for both unburned and burned plots during both sampling
periods (Table 2). No significant difference was observed between woodland develop-
ment Phase I and II for the burned plots in either sampling period or for the unburned
plots in sampling period 1. However, a statistical difference was observed between un-
burned Phase I and II plots in sampling period 2, but given the relatively low A value (A
= 0.144), the ecological difference is small. Differences were observed between Phase I
and III, Phase II and III, and Phase II and Mature for both unburned and burned plots in
both time periods. No difference was observed between Phase III and Mature within
unburned or burned plots for either sampling period.
Table 2. Differences in plant community composition between woodland development phases (PI,
PII, PIII and Mature) for unburned and burned plots for the two time periods sampled. MRPP
outputs are reported: T = test statistic; A = chance-corrected within-group agreement; p = probabil-
ity. Probability values < 0.05 are in bold font. See methods section for more details about the MRPP
statistics.
Sample Year 1
Unburned Plots
Burned Plots
T
A
p
T
A
p
Overall
−6.186
0.305
<0.001
−8.063
0.162
<0.001
PI vs. PII
−1.522
0.060
0.079
−1.778
0.039
0.054
PI vs. PIII
−5.532
0.383
0.001
−6.360
0.124
<0.001
Figure 5.
Average of Shannon’s Diversity Index (H’) for plots by woodland successional phase
pre-fire (blue circles), 5–6 years after the wildfire (Time 1, orange circles middle row) and
12–13 years
after the wildfire (Time 2, yellow circles bottom row). Sorensen’s Dissimilarity Index (SDI) is a
measure of dissimilarity or species turnover along the successional gradient and following the
wildfire. High values represent high turnover of species. The largest observed change along the
woodland development gradient is seen between Phase II and III. Species turnover because of the fire
increased along the successional gradient with almost a complete replacement of the plant community
in the Mature woodlands (SDI = 0.945).
Differences in plant community composition were evaluated with the Multi-Response
Permutation Procedure [
31
]. Overall differences between woodland development phases
were observed for both unburned and burned plots during both sampling periods (Table 2).
No significant difference was observed between woodland development Phase I and II for
the burned plots in either sampling period or for the unburned plots in sampling period 1.
However, a statistical difference was observed between unburned Phase I and II plots in
sampling period 2, but given the relatively low A value (A = 0.144), the ecological difference
is small. Differences were observed between Phase I and III, Phase II and III, and Phase
II and Mature for both unburned and burned plots in both time periods. No difference
was observed between Phase III and Mature within unburned or burned plots for either
sampling period.

Fire 2023,6, 141 12 of 25
Table 2.
Differences in plant community composition between woodland development phases (PI, PII,
PIII and Mature) for unburned and burned plots for the two time periods sampled. MRPP outputs are
reported: T = test statistic; A = chance-corrected within-group agreement; p= probability. Probability
values < 0.05 are in bold font. See methods section for more details about the MRPP statistics.
Sample Year 1 Unburned Plots Burned Plots
T A pT A p
Overall −6.186 0.305 <0.001 −8.063 0.162 <0.001
PI vs. PII −1.522 0.060 0.079 −1.778 0.039 0.054
PI vs. PIII −5.532 0.383 0.001 −6.360 0.124 <0.001
PI vs. Mature −5.958 0.360 0.001 −7.696 0.240 <0.001
PII vs. PIII −4.439 0.246 0.003 −5.048 0.107 0.001
PII vs. Mature −4.207 0.215 0.004 −6.186 0.207 0.001
PIII vs. Mature 0.399 −0.020 0.550 −1.222 0.024 0.112
Sample Year 2 T A pT A p
Overall −7.448 0.294 <0.001 −7.570 0.144 <0.001
PI vs. PII −3.900 0.144 0.005 −0.591 0.011 0.240
PI vs. PIII −5.591 0.387 0.002 −8.418 0.158 <0.001
PI vs. Mature −6.187 0.358 0.001 −7.837 0.233 <0.001
PII vs. PIII −4.551 0.167 0.003 −4.297 0.077 0.004
PII vs. Mature −4.872 0.162 0.002 −5.267 0.137 0.001
PIII vs. Mature −0.468 0.012 0.306 0.311 −0.005 0.515
According to expectation, differences in plant community composition were observed
between unburned and burned plots within each woodland development phase for both
sampling periods. The differences were more pronounced in the advanced woodland
development stages (A = 0.409, p< 0.001 for sampling period 1 and A = 0.415, p< 0.001 for
sampling period 2 in Mature plots) compared to Phase I (A = 0.083, p= 0.015 for sampling
period 1 and A = 0.160, p= 0.040 for sampling period 2), indicating longer-lasting effects of
wildfire in the more advanced woodland phases.
Differences in plant community composition were observed in all woodland develop-
ment phases between sampling periods in the burned plots, indicating continued succes-
sional development in the plots. However, the A-values were relatively low (0.046–0.098),
indicating moderate ecological differences, which could be expected given that the samples
were taken only seven years apart in a semi-arid ecosystem. As expected, no difference in
plant community composition was observed between time periods for unburned plots in
any phase.
3.3. Functional Groups
Changes in plant cover by functional group were summarized by time period and
woodland development phase, and differences between unburned and burned plots were
evaluated statistically.
Annual grass cover was low across unburned control plots and burned plots for
all woodland development phases for both sampling periods. No significant difference
was detected when unburned and burned plots of the same phase and time period were
compared (Figure 6). Frequent annual grasses were cheatgrass and soft brome. Generally,
cheatgrass cover was less than 0.1% with occasional plots up to 2% (Table S1). Annual
grasses increased slightly after the burn, particularly in developed woodlands; 5–6 years
after fire, cheatgrass cover was 1.8
±
4.0% in woodlands that were in Phase III pre-fire and
0.8
±
1.4% in woodlands that were in the Mature class pre-fire (Table S1). By the second

Fire 2023,6, 141 13 of 25
sampling period, cheatgrass had decreased to 0.3
±
0.5 in the burned Phase III woodlands
and to 0.5
±
0.4 in the burned Mature woodlands. Soft brome was present on many plots
at low cover levels (0–1%).
Fire 2023, 6, x FOR PEER REVIEW 13 of 25
Figure 6. Comparison of canopy cover of functional plant groups between burned and unburned
plots by woodland development phase and sampling period. Significant differences (p < 0.05) in
cover between burned and unburned plots are marked with a star (*).
Figure 6.
Comparison of canopy cover of functional plant groups between burned and unburned
plots by woodland development phase and sampling period. Significant differences (p< 0.05) in
cover between burned and unburned plots are marked with a star (*).
Perennial forb cover was not significantly affected by the wildfire when compared
within woodland development phase for either sampling period. Perennial forbs were
around 10% cover in Phase I in the first sampling period and around 15% in the second
sampling period. Decreasing cover levels of perennial forbs were observed along the
successional gradient. In Phase II woodlands, the perennial forb cover was around 8%
during the first sampling period and around 10% in the second sampling period. In
Phase III woodlands, perennial forb cover was <5% and varied in Mature woodlands
(2–9% cover).

Fire 2023,6, 141 14 of 25
Perennial grass cover was not affected by the wildfire in any woodland phase in the
first sampling period. However, during the second sampling period, perennial grass cover
was significantly higher on burned plots of Phase II (Figure 6d) and Phase III (Figure 6f).
Overall, perennial grass cover was about 10% in both unburned and burned Phase I
woodlands in sampling period 1 and around 20% in sampling period 2. In burned Phase
II woodlands, the perennial grass cover was around 7% in sampling period 1 and over
20% by sampling period 2, a significant increase compared to unburned plots. In Phase
III woodlands, perennial grasses were lower, around 5%, but increased to over 10% by the
second sampling period. Trends in Mature woodlands were similar to those in Phase III,
but no significant differences were documented.
Shrub cover was significantly lower in burned Phase I woodlands compared to un-
burned, indicating that shrubs had not returned to pre-fire cover levels after 5–6 years
(Figure 6a) while 12–13 years post-fire, shrub cover was approaching pre-fire cover levels,
and the difference in shrub cover was no longer significant between burned and unburned
plots (Figure 6b). Shrub cover in unburned Phase I plots was dominated by mountain
big sagebrush (13.4
±
9.0%; Table S1), while burned plots had a higher diversity of shrub
species. Mountain big sagebrush cover was still low in burned Phase I plots (3.9
±
9.0%;
Table S1), but sprouting shrubs (e.g., rabbitbrush and bitterbrush) and seedbank shrubs
(e.g., snowbrush ceanothus) had increased. In Phase II woodlands, significant reductions
of shrub cover was documented because of the fire in the first sampling period (Figure 6c)
with no discernable differences in the second sampling period. Similar to unburned Phase
I woodlands, mountain big sagebrush was the dominant shrub in unburned Phase II wood-
lands (5.9
±
4.3%; Table S1), while the burned Phase II woodlands had low shrub cover
in sampling period 1, but by sampling period 2, shrub cover had increased to almost 14%
on average and was composed of a variety of shrubs including mountain big sagebrush
(
2.4 ±3.1%
; Table S1), snowbrush ceanothus (8.3
±
10.7%; Table S1) and small amounts
of other shrubs including rabbitbrush, bitterbrush, mountain snowberry and wax currant
(Ribes cereum Douglas). Shrub cover in unburned Phase III and Mature woodlands was
low (<4%) and composed of small amounts of mountain big sagebrush (<1%), curl-leaf
mountain mahogany, rabbitbrush, bitterbrush and mountain snowberry; notably, snow-
brush ceanothus was not detected along the transects in unburned plots but occasionally
observed in the general area.
The only tree species present on the plots was western juniper. Juniper cover decreased
within all woodland phases as a result of the wildfire as expected, with a more pronounced
change in the later woodland development phases (Figure 6). Juniper cover pre-fire was
<5% in Phase I woodlands, around 15% in Phase II woodlands, and in Phase III and Mature
woodlands, juniper cover was over 40%. Juniper trees were still absent after 12–13 years in
burned areas (Figure 6b,d,f,h). Juniper seedlings will be discussed later.
Annual forbs were not analyzed for change in cover because of small cover values
(<1%) across plots, which are difficult to accurately assess in the field. Common annual
forb results are therefore reported in terms of frequency later in the manuscript.
3.4. Changes in Cover of Dominant Species
Cover of the 14 most dominant species in the dataset are reported by woodland
development phase, burn status and sampling period in Table S1. Dominant perennial
forbs were arrowleaf balsamroot, tapertip hawksbeard and lupine. They were present in
all woodland development phases during both sampling periods and in unburned and
burned plots. Generally, higher cover levels (up to 6.3% on average) were present in earlier
woodland phases. By the second sampling period arrowleaf balsamroot was higher in
burned than in unburned Phase I woodlands. Tapertip hawksbeard was higher in all
woodland development phases in the second sampling period in burned plots and was
significantly higher in burned compared to unburned plots, with cover levels around 2% on
average. Lupine did not show a significant change in either sampling period or woodland

Fire 2023,6, 141 15 of 25
phase but tended to be higher (2–5%) in early woodland phases and lower in late woodland
phases (<1%).
Dominant perennial grasses included Columbia (A. nelsonii [Scribn.] Barkworth)
and western needlegrass, bottlebrush squirreltail (Elymus elymoides [Raf.] Swezey), Idaho
fescue, Sandberg bluegrass and bluebunch wheatgrass. These grasses were present in all
unburned or burned woodland development phases in both sampling periods. Overall,
these perennial grasses tended to increase over time and respond positively to the burn,
except Idaho fescue, which was lower in burned compared to unburned Mature woodlands.
Columbia needlegrass cover was on average 0.2–1.9% across unburned and burned plots.
Columbia needlegrass was significantly higher in burned Phase II, Phase III and Mature
plots in the second sampling period and was also higher in burned compared to unburned
Phase II plots. No significant changes were observed for western needlegrass, which was
present at 0.1–1.7% on average across unburned and burned woodland phases. Bottlebrush
squirreltail was significantly higher in burned Phase II, Phase III and Mature plot in the
second sampling period and was also higher in burned Phase III and Mature plots compared
to unburned plots. Idaho fescue was higher in the second sampling period in unburned
Phase I and Mature plots and in burned Phase I and II plots. Sandberg bluegrass was higher
in the second sampling period in unburned Phase I plots and in burned Phase I, II and III
plots. Sandberg bluegrass was higher in burned Phase III and Mature plots compared to
unburned plots of the same phases. Bluebunch wheatgrass was higher in burned Phase I
and II plots compared to unburned plots and increased over time in these phases.
Dominant shrub species included mountain big sagebrush, snowbrush ceanothus
and mountain snowberry. Mountain big sagebrush was unsurprisingly lower in burned
Phase I woodlands compared to unburned, and mountain snowberry was lower in burned
Phase III woodlands compared to unburned. Snowbrush ceanothus was higher in burned
Phase III and Mature woodlands compared to unburned and increased significantly over
time in those woodland phases. Snowbrush ceanothus was not present in unburned plots
of any phase but by the second sampling period was very dominant in burned Phase III
(35.0 ±27.3%) and Mature woodlands 38.0 ±20.3%).
The only tree species present in the area was western juniper. The juniper cover in
Phase I was low prior to the fire (3–4% on average) and increased along the successional
gradient to 10–20% in Phase II and over 40% in Phase III and in the Mature woodlands.
Juniper decreased to close to 0% in Phase II, Phase III and Mature woodlands because of the
high fire severity. A couple of Phase III plots at the edge of the fire burned at low severity,
and in these plots a few large juniper trees survived.
Cheatgrass was absent in unburned plots in the first sampling period but present at
low levels (~1%) during the second sampling period. No significant changes in cheatgrass
were detected along the successional gradient or because of the wildfire (Table S1).
3.5. Species with High Frequency but Low Cover
Species with low cover were compared using frequency rather than cover because it is
very difficult to assign accurate cover values to species with low cover (<1%). Frequent
species are listed in Table 3. Frequent annual forbs with low cover included Collomia linearis,
Collinsia parviflora,Cryptantha spp., Epilobium brachycarpum,Lactuca serriola and Stellaria spp.
These annual forbs occurred frequently across woodland development phases and in both
burned and unburned areas. A few trends can be observed. Collomia linearis occurred across
unburned and burned plots of all woodland phases, with higher frequency the second
sampling period. Collinsia parviflora frequency tended to be lower in burned plots and also
decreased between time periods. Cryptantha spp. frequency was higher in burned plots
but decreased over time; however, even in the second sampling period, the frequency was
higher in burned plots. Epilobium brachycarpum and Lactuca serriola frequency was higher in
burned plots compared to unburned, and L. seriola tended to decrease over time. Stellaria
was higher in burned plots the first sampling period but had decreased by the second
sampling period.

Fire 2023,6, 141 16 of 25
Table 3. Mean frequency (%) and standard deviation in parenthesis for species that occurred frequently in sampled quadrats but at low percent cover.
Period Phase Burn N dNBR
1-yr
Bromus
tectorum
Collomia
linearis
Collinsia
parviflora
Crepis
acuminata
Cryptantha
spp.
Epilobium
brachty-
carpum
Lactuca
serriola
Phacelia
heterophylla
Phlox
longifolia
Stellaria
spp.
Tragopogon
dubius
Viola
purpurea
1 1 0 5 −36 (11) 0.0 76.0 (15.5) 49.2 (9.7) 22.8 (17.5) 8.4 (18.8) 5.6 (11.4) 0.0 0.0 46.4 25.5) 47.6 (24.6) 1.6 (2.6) 10.8 (11.4)
1 1 1 7 47 (93) 11.4 (18.8) 60.9 (10.0) 35.1 (22.4) 51.1 (20.3) 24.3 (16.1) 33.1 (36.0) 12.6 (14.6) 0.0 53.7 (14.9) 65.1 (21.9) 4.6 (4.6) 7.1 (8.8)
1 2 0 5 −23 (35) 0.0 53.6 (49.9) 48 (48.4) 36.8 (33.4) 3.6 (3.1) 8.0 (9.6) 0.0 0.0 44.8 (42.6) 60.0 (62.0) 0.0 15.2 (12.6)
1 2 1 5 139 (123) 4.7 (8.2) 51.0 (25.3) 37.7 (19.4) 46.0 (10.9) 53.3 (16.1) 29.3 (22.1) 19.3 (18.3) 8.7 (17.4) 32.7 (20.7) 73.7 (14.6) 8.3 (6.3) 3.0 (2.1)
1 3 0 5 −20 (12) 0.0 27.2 (22.3) 52.4 (32.1) 5.2 (6.4) 5.6 (6.7) 0.4 (0.9) 0.0 0.8 (1.8) 10.4 (10.9) 49.6 (13.0) 0.0 32.8 (10.6)
1 3 1 15 284 (16) 21.7 (25.3) 19.9 (16.0) 3.0 (23.3) 10.0 (13.7) 50.0 (28.5) 27.6 (25.7) 36.9 (29.8) 18.7 (18.3) 6.3 (8.0) 61.5 (15.9) 7.2 (9.9) 8.5 (7.5)
1 4 0 5 −12 (17) 11.7 (15.1) 38.7 (14.8) 76.0 (11.8) 8.7 (12.2) 5.3 (4.3) 4.3 (5.0) 0.3 (0.8) 1.3 (1.6) 10.7 (9.4) 50.3 (32.3) 0.0 40.0 (15.4)
1 4 1 8 396 (74) 23.5 (20.5) 19.8 (16.5) 23.5 (19.9) 25.5 (17.3) 77.3 (14.8) 67.5 (20.2) 47.8 (29.7) 30.8 (22.9) 14.5 (11.1) 49.8 (26.8) 5.0 (4.4) 5.8 (5.7)
2 1 0 7 −36 (11) 2.8 (5.2) 57.2 (19.6) 9.2 (3.0) 24.8 (15.4) 12.8 (13.2) 12.0 (10.8) 0.4 (0.9) 0.0 57.2 (23.8) 23.2 (14.5) 0.4 (0.9) 12.8 (10.0)
2 1 1 7 47 (93) 10.3 (14.3) 63.7 (15.3) 14.0 (13.0) 47.1 (14.3) 23.1 (17.2) 54.3 (29.4) 2.6 (4.3) 0.3 (0.8) 54.0 (15.6) 10.3 (10.5) 17.7 (19.2) 12.0 (13.1)
2 2 0 5 −23 (35) 1.2 (1.1) 37.2 (27.5) 23.6 (15.1) 18.0 (14.8) 17.6 22.7) 9.6 (10.4) 0.0 1.2 (1.1) 33.2 (23.9) 47.2 (20.5) 0.4 (0.9) 28.4 (16.9)
2 2 1 5 139 (123) 11.3 (18.4) 59.3 (16.5) 9.0 (10.5) 58.0 (11.0) 31.0 (20.1) 43.3 (14.1) 7.7 (9.4) 3.7 (9.0) 40.7 (12.6) 11.3 (9.3) 22.7 (10.4) 7.3 (6.2)
2 3 0 5 −20 (12) 4.8 (7.8) 34.8 (31.3) 56.0 (15.0) 2.0 (2.8) 11.6 (9.0) 17.2 (15.0) 1.2 (1.8) 6.8 (8.3) 11.2 (11.2) 47.6 (28.8) 0.0 53.2 (6.6)
2 3 1 14 284 (165) 22.3 (18.8) 52.6 (17.3) 9.3 (10.2) 24.9 (17.8) 28.1 (15.0) 36.0 (26.0) 10.7 (8.4) 21.6 (19.8) 11.9 (13.3) 17.9 (12.7) 32.9 (19.5) 8.6 (6.3)
2 4 0 5 −12 (17) 19.2 (20.5) 35.6 (19.1) 31.2 (17.9) 11.2 (9.2) 12.0 (9.9) 11.6 (6.2) 1.2 (1.1) 4.4 (5.4) 19.6 (15.6) 57.2 (22.5) 0.4 (0.9) 43.6 (16.8)
2 4 1 8 396 (74) 39.0 (21.3) 49.5 (21.7) 6.0 (4.4) 27.5 (17.1) 32.8 (12.7) 37.8 (17.0) 13.8 (9.0) 38.5 (27.9) 13.5 (12.1) 18.3 (15.6) 42.0 (27.6) 7.5 (10.1)

Fire 2023,6, 141 17 of 25
Frequent perennial forbs included Crepis acuminata, Phacelia heterophylla, Phlox
longifolia and Tragopogon dubius. C. acuminata and P. heterophylla increased with fire,
particularly in the second time period. P. longifolia decreased with fire. The frequency of
Bromus tectorum increased with fire.
3.6. Juniper Seedlings
Juniper seedlings were counted within the quadrats along the transects on burned
plots and unburned control plots. The number of seedlings was zero or near zero on both
burned and unburned plots when the pre-fire juniper cover was less than 10%, but as
pre-fire juniper cover increased, more seedlings were observed, particularly on unburned
plots (Figure 7a). Significantly higher density of juniper seedlings was found on unburned
plots for Phase II, while significant differences between burned and unburned plots were
not detected in Phase I, III or Mature woodlands (Figure 7b). Lack of difference in Phase III
and Mature woodlands is likely due to high variability in the seedling density in unburned
woodlands.
Fire 2023, 6, x. https://doi.org/10.3390/xxxxx www.mdpi.com/journal/fire
and in both burned and unburned areas. A few trends can be observed. Collomia linearis
occurred across unburned and burned plots of all woodland phases, with higher fre-
quency the second sampling period. Collinsia parviflora frequency tended to be lower in
burned plots and also decreased between time periods. Cryptantha spp. frequency was
higher in burned plots but decreased over time; however, even in the second sampling
period, the frequency was higher in burned plots. Epilobium brachycarpum and Lactuca
serriola frequency was higher in burned plots compared to unburned, and L. seriola tend-
ed to decrease over time. Stellaria was higher in burned plots the first sampling period
but had decreased by the second sampling period.
Frequent perennial forbs included Crepis acuminata, Phacelia heterophylla, Phlox
longifolia and Tragopogon dubius. C. acuminata and P. heterophylla increased with fire,
particularly in the second time period. P. longifolia decreased with fire. The frequency of
Bromus tectorum increased with fire.
3.6. Juniper Seedlings
Juniper seedlings were counted within the quadrats along the transects on burned
plots and unburned control plots. The number of seedlings was zero or near zero on
both burned and unburned plots when the pre-fire juniper cover was less than 10%, but
as pre-fire juniper cover increased, more seedlings were observed, particularly on un-
burned plots (Figure 7a). Significantly higher density of juniper seedlings was found on
unburned plots for Phase II, while significant differences between burned and unburned
plots were not detected in Phase I, III or Mature woodlands (Figure 7b). Lack of differ-
ence in Phase III and Mature woodlands is likely due to high variability in the seedling
density in unburned woodlands.
Figure 7. Juniper seedlings per m2 versus pre-fire percentage juniper cover for unburned and
burned plots (a) and by woodland development phase (b). Error bars represent standard devia-
tion, and significance (p < 0.05) is denoted with a star (*). Both graphs reflect data from sampling
period 2 (12–13 years post-fire); no juniper seedlings were observed along the transects on burned
plots in sampling period 1.
3.7. Successional Pathways
Our research suggests that late successional plots that burn at high severity may not
necessarily return to sagebrush steppe vegetation but rather may follow a different sere
such as one dominated by snowbrush ceanothus during the pathway back to mature ju-
niper woodland (Figure 8). The snowbrush ceanothus pathway may occur in areas
where climate conditions are suitable for the species, and snowbrush seeds are present
in the seedbank. In Idaho, snowbrush ceanothus has been reported to occur across a
wide elevation range (1150–3000 m) and annual precipitation range (690–1140 mm) [44],
Figure 7.
Juniper seedlings per m
2
versus pre-fire percentage juniper cover for unburned and burned
plots (
a
) and by woodland development phase (
b
). Error bars represent standard deviation, and
significance (p< 0.05) is denoted with a star (*). Both graphs reflect data from sampling period 2
(12–13 years post-fire); no juniper seedlings were observed along the transects on burned plots in
sampling period 1.
3.7. Successional Pathways
Our research suggests that late successional plots that burn at high severity may not
necessarily return to sagebrush steppe vegetation but rather may follow a different sere
such as one dominated by snowbrush ceanothus during the pathway back to mature juniper
woodland (Figure 8). The snowbrush ceanothus pathway may occur in areas where climate
conditions are suitable for the species, and snowbrush seeds are present in the seedbank.
In Idaho, snowbrush ceanothus has been reported to occur across a wide elevation range
(1150–3000 m) and annual precipitation range (690–1140 mm) [
44
], and in Utah, snowbrush
ceanothus has been observed at lower precipitation ranges (410–510 mm; [45]).

Fire 2023,6, 141 18 of 25
Fire 2023, 6, x FOR PEER REVIEW 2 of 25
and in Utah, snowbrush ceanothus has been observed at lower precipitation ranges
(410–510 mm; [45]).
Figure 8. Proposed post-fire successional pathways as influenced by pre-fire plant community
composition and structure. Early successional stages (Phase I and II) result in a herb-dominated
community 12–13 following fire. Sorensen’s Dissimilarity Index (SDI) indicates dissimilarity be-
tween phases and ranges from 0 to 1, where 0 means the plots are identical, and 1 means no simi-
larity (complete dissimilarity). Pre-fire, plant communities of Phase I and II plots are somewhat
similar (SDI = 0.419), while Phase III and Mature plots are very similar (SDI = 0.075). Following
fire, burned Phase I and II plots become more similar (SDI = 0.344 after 5–6 years, and SDI = 0.328
after 12–13 years). Burned Phase III and Mature plots also stay similar as succession progresses
(SDI = 0.263 after 5–6 years, and SDI = 0.162 after 12–13 years). The dashed-line box indicates that
the communities within the box are not significantly different from each other in spite of being ini-
tially different pre-fire.
4. Discussion
4.1. Richness and Diversity
Average richness of vascular plant species decreased slightly along the successional
gradient in unburned plots, which was also reflected in burned plots during the first
sampling period, but this decrease was not observed in the second sampling period. The
average richness per sampled plot remained fairly constant along the sere and in both
unburned and burned plots, at 28–35 species 5–6 years post-fire and 35–40 species 12–13
years post-fire. Species richness increased between the two sampling periods across
woodland development phases, except in Phase II. We attribute this change to succes-
sional development within the plots. Diversity (H’) was lower in late successional phases
(Phase III and Mature) compared to the early successional phases (Phase I and II) in both
unburned and burned plots. Diversity accounts for both changes in richness and relative
abundance, and since richness did not change much along the sere, we attribute the de-
crease in diversity to the increased dominance of juniper trees in unburned late-
successional plots and to the dominance of snowbrush ceanothus in burned plots that
were late-successional pre-fire. Diversity increased over time in burned Phase III plots
but not in earlier successional phases. The increase in diversity can be attributed to the
documented increased richness as the plots began to recover from the burn.
Figure 8.
Proposed post-fire successional pathways as influenced by pre-fire plant community compo-
sition and structure. Early successional stages (Phase I and II) result in a herb-dominated community
12–13 following fire. Sorensen’s Dissimilarity Index (SDI) indicates dissimilarity between phases
and ranges from 0 to 1, where 0 means the plots are identical, and 1 means no similarity (complete
dissimilarity). Pre-fire, plant communities of Phase I and II plots are somewhat similar (
SDI = 0.419
),
while Phase III and Mature plots are very similar (SDI = 0.075). Following fire, burned Phase I and II
plots become more similar (SDI = 0.344 after 5–6 years, and SDI = 0.328 after
12–13 years
). Burned
Phase III and Mature plots also stay similar as succession progresses (SDI = 0.263 after 5–6 years, and
SDI = 0.162 after
12–13 years
). The dashed-line box indicates that the communities within the box are
not significantly different from each other in spite of being initially different pre-fire.
4. Discussion
4.1. Richness and Diversity
Average richness of vascular plant species decreased slightly along the successional
gradient in unburned plots, which was also reflected in burned plots during the first
sampling period, but this decrease was not observed in the second sampling period.
The average richness per sampled plot remained fairly constant along the sere and in
both unburned and burned plots, at 28–35 species 5–6 years post-fire and 35–40 species
12–13 years
post-fire. Species richness increased between the two sampling periods across
woodland development phases, except in Phase II. We attribute this change to successional
development within the plots. Diversity (H’) was lower in late successional phases (Phase
III and Mature) compared to the early successional phases (Phase I and II) in both unburned
and burned plots. Diversity accounts for both changes in richness and relative abundance,
and since richness did not change much along the sere, we attribute the decrease in diversity
to the increased dominance of juniper trees in unburned late-successional plots and to the
dominance of snowbrush ceanothus in burned plots that were late-successional pre-fire.
Diversity increased over time in burned Phase III plots but not in earlier successional
phases. The increase in diversity can be attributed to the documented increased richness as
the plots began to recover from the burn.
We also evaluated the contribution of richness and diversity of each woodland de-
velopment phase to the richness and diversity of the entire landscape. For this analysis,
we computed the sum of all species in plots of each phase rather than averages. Species
richness was in the 46–74 range across phases for burned and unburned plots; however,

Fire 2023,6, 141 19 of 25
total species richness for the landscape (all plots) was double that, 124 in sampling period 1
and 131 in sampling period 2 (Figure 4a,b). The higher values for the total landscape can
be explained by species turnover resulting from successional development and from the
burn. Earlier successional phases, however, contribute more to the vascular plant diver-
sity of the landscape compared to later successional stages (Figure 4c,d). In fact, species
diversity of the early successional phases is equal to or higher than the species diversity
of the landscape. We can also conclude that fire contributed to increased vascular plant
diversity given that diversity is higher in burned plots when compared to unburned plots
of the same phase. These data emphasize the importance of maintaining early successional
habitats for maximizing vascular plant diversity, but also confirm that all woodland devel-
opment phases contribute to species richness in the landscape. It has been demonstrated
that the changes in plant community composition and vegetation structure cascade into
changes in other taxa, birds for example. Pavlacky and Anderson [
46
] observed richness
and diversity in the bird community along successional and elevational gradients in Utah
juniper woodlands, concluding that all woodland successional stages contribute to bird
gamma diversity in these woodlands. Bird surveys following juniper removal projects have
demonstrated that juniper removal increases abundance of nesting pairs of several ground-
and shrub-nesting species, while species associated with woodland vegetation decline [
47
].
These data emphasize the need of having multiple plots of a variety of successional stages
present in the total landscape to provide habitat for all species.
4.2. Plant Community Turnover and Composition
Even though no differences in species richness or diversity were detected between
unburned and burned plots of the same phase in any of the time periods sampled, the burn
resulted in changes in species composition demonstrated by the Sorensen Dissimilarity
Index (SDI). The difference in species composition (tested with MRPP) between unburned
Phase I and II plots was significant in sampling period 2 but not in sampling period 1.
The species turnover indicated by an SDI of 0.419 was also relatively low, suggesting
strong similarity between Phase I and II woodlands. The largest dissimilarity in species
composition (largest turnover of species) in unburned plots was seen between woodland
development Phase II and III. The transition from Phase II to Phase III woodlands is the
stage in succession when woodland species become dominant over sagebrush steppe
species, as previously described by Miller et al. [
4
]. Strand et al. [
22
] also documented
that the wildfire burn severity was significantly higher in Phase III plots compared to
Phase I or II plots in the TCWFC. Weiner et al. [
18
] showed that the increase in litter and
duff ground fuels that have the potential to smolder for long periods of time after the fire
front has passed could be one mechanism that explains the higher burn severity in the
late successional woodlands. Obviously, the heat pulse from torching juniper crowns also
contributes heat to the soil and seed bank that may sustain the burning ground fuels and
contribute heat that consumes the organic soil layer, seed banks and other regenerative plant
structures in the soil. No difference in species composition between Phase III and Mature
woodlands was detected (MRPP analysis) and the dissimilarity between the plots was very
low (
SDI = 0.075
). We interpret these similarities between Phase III and Mature woodlands
as a support for the statement that Phase III woodlands have passed a threshold for when
ecological processes are dominated by the juniper rather than the steppe vegetation as
suggested [
4
]. Additional support for the suggestion that an ecological threshold has
been passed in Phase III woodlands is the dissimilarity observed between unburned and
burned plots along the successional gradient. The dissimilarity between unburned and
burned plots of Phase I is 0.419, then 0.643, 0.745 and 0.945 for Phase II, III and Mature
plots, respectively. The very high dissimilarity between unburned and burned plots (0.945)
indicates near complete species turnover since max SDI = 1 for complete turnover in species
between plots.

Fire 2023,6, 141 20 of 25
4.3. Functional Groups and Common Species
Increases in flammable exotic annual grasses following fire in sagebrush steppe and
juniper woodlands has been documented and is a concern. However, we did not detect
a significant difference in annual grass cover comparing unburned and burned plots.
Cheatgrass cover was absent or low (<1%) in unburned plots, but higher values were
detected after fire, with the largest increase in Phase III woodlands (1.8
±
4.0%; Table S1),
although the increase was not significant at p= 0.05. Frequency of cheatgrass was also
higher in burned plots (10–40%) compared to unburned (0–10%), see Table 3. We do not
anticipate that cheatgrass will become dominant in the burned area since the low cover
levels have remained low for more than 10 years after the fire. Variability in annual grass
response after fire in western juniper woodlands has been reported in the literature. A
study by Weiner et al. [
18
] documented high cover of cheatgrass (9.7
±
22%) under burned
tree canopies six years post-fire in the TCWFC. The highest cover of cheatgrass occurred
in areas that burned at low severity with very little cheatgrass in the high-severity areas,
possibly because the organic layer of the soil was consumed by the fire, including the
cheatgrass seedbank. The sampled plots within the TCWFC were situated on an ecological
site classified as high resistance to annual grass invasion because of its higher precipitation
zone and soil moisture and temperature regime. On plots with lower resistance and warmer
and dryer soil moisture regimes, effects of wildfire on annual grasses could have been very
different, with risk of annual grass dominance and initiation of an annual grass–fire cycle.
Several annual forbs occurred frequently across unburned and burned plots through-
out both sampling periods including Collomia linearis,Collinsia parviflora,Cryptantha spp.,
Epilobium brachycarpum,Lactuca seriola,Agoseris Raf. spp. and Stellaria spp. These species
establish quickly after fire, but little else is known about the ecology of them. We found it
interesting that these forbs were frequent across burned areas and throughout the sere in
unburned areas. Forbis [
48
] suggested that native annual forbs are phenologically similar to
cheatgrass and therefore may use similar resource pools. Leger et al. [
49
] documented such
competition in greenhouse experiments for selected forbs. For example,
Amsinckia tessellata
A. Gray reduced B. tectorum biomass by 97%, and Amsinckia intermedia Fisch. and C. A. Mey.,
Amsinckia tessellate and Descurainia pinnata (Walter) Britton reduced seed output between
79 and 87%. Further research is needed to better understand the ecology of these frequent
and persistent annual forbs.
Perennial grasses and forbs were more abundant in early successional stages when
compared to late successional stages in unburned plots. Several other studies have doc-
umented this decline in understory herbaceous perennial vegetation along the juniper
woodland development gradient [
1
,
4
,
7
,
50
]. No difference in perennial grass cover was
documented between unburned and burned areas 5–6 years post-fire, but by the second
sampling period, 12–13 years post-fire, perennial grass cover was higher in burned plots
compared to unburned plots for Phase II and III woodlands. Bottlebrush squirreltail,
Columbia needlegrass and bluebunch wheatgrass all increased post-fire. However, Idaho
fescue was lower in burned Mature plots compared to unburned even after 12–13 years,
which can be expected given the high severity burn in Mature plots. Idaho fescue is rela-
tively sensitive to fire because the budding areas are located above or at the soil surface and
are therefore more sensitive to fire compared to bluebunch wheatgrass for example [
51
].
No significant differences were detected in perennial forb cover when unburned plots were
compared to burned plots of the same woodland phase. Bates et al. [
7
] similarly found
no difference in perennial forb yield when comparing prescribed burn plots to unburned
control plots; however, they observed an increase in tall perennial forbs and a decrease
in mat-forming forbs after the burn. It is important to recognize that response to fire is
different for each individual forb species as described by Pechanec et al. [
52
] who classified
forbs common to sagebrush steppe by susceptibility to fire damage.
The shrub community was significantly different between unburned and burned plots
in all woodland development phases after 5–6 years. In the early unburned woodland
phases, big sagebrush was the dominant shrub. However, fire is lethal to mountain big

Fire 2023,6, 141 21 of 25
sagebrush, and the shrub tends to be slow to return to the site after fire because it does not
sprout from root or crown, the seed is short-lived in the seed bank [
53
] and seeds do not
spread more than 3 m from shrubs in nearby unburned areas [
54
,
55
]. In the years following
the fire, mountain big sagebrush cover increased slowly and was 3.9
±
9.0% in burned
Phase I woodlands and 2.4
±
3.1% in Phase II woodlands 12–13 years post-fire. Innes and
Zouhar [
56
] synthesized literature reporting that it may take 20–26 years for sagebrush
cover to return to pre-burn levels on mountain big sagebrush sites. Shrub cover of sprouting
shrubs such as rabbitbrush and bitterbrush increased more rapidly, overall increasing the
shrub diversity in burned early successional plots. Increased shrub diversity with a similar
species composition has previously been reported following prescribed fire in mountain
big sagebrush steppe [
27
]. While shrub cover in early woodland development phases
decreased by fire, shrub cover in late woodland successional stages increased. Average
shrub cover in late woodland development phases was low in unburned plots (0–4%)
composed of various mixes of mountain big sagebrush, rabbitbrush, antelope bitterbrush,
curl-leaf mountain mahogany, mountain snowberry and wax currant. In burned late
woodland development phases, snowbrush ceanothus was by far the most dominant
shrub, with average cover levels approaching 40% canopy cover 12–13 years post-fire. The
dramatic increase in snowbrush ceanothus in burned late woodland development phases
was a surprise because the shrub was rarely present, and if so, only in small amounts,
in unburned plots. Snowbrush ceanothus is native to western North America and has
commonly been observed to increase following wildfire or prescribed fire even in areas
where it was uncommon or absent prior to fire [
57
,
58
]. The seed of snowbrush ceanothus
can remain viable for more than two centuries in the soil [
59
], and it has long been known
that the seed requires a heat treatment to germinate [
60
]. Apparently, snowbrush ceanothus
seeds must have been present deep in the seedbank of the late successional woodlands,
which is surprising given that most of the organic matter had been consumed by the high-
severity fire. Ceanothus was particularly prevalent in areas under the burned tree crowns
where duff and litter from the old juniper trees had accumulated for in some cases several
centuries [
18
]. Even though some sagebrush was present in the Phase III and Mature
woodlands pre-fire, it is clear that those sites are unlikely to return to sagebrush steppe
vegetation but will more likely remain a snowbrush ceanothus shrubland for a period of
time, then eventually returning to western juniper dominance.
Juniper seedling counts were low (0.01 seedlings per square meter) in unburned
Phase I woodlands but almost 1 seedling per square meter on average in unburned Phase
II woodlands. By the time the woodland had transitioned into Phase II, many juniper
trees had reached maturity, and we anticipate that seeds were abundant in the seed bank.
Phase III and Mature woodlands had 0.75 and 0.5 seedlings per square meter on average,
respectively, with high variability around the mean, indicating that seeds were present,
and germination was occurring. Wozniak and Strand [
61
] documented similar seedling
counts in Phase I western juniper woodlands with higher counts in later successional stages.
Juniper trees were generally killed by the high-intensity fire, including centuries-old trees
in Mature plots. Only a few juniper trees survived in plots located on the edge of the fire
that burned in low or moderately low severity. The loss of juniper in Phase I plots did
not result in a large change in composition since juniper only accounted for up to 10% of
the canopy cover in these plots. The loss of juniper in Phase III and Mature plots where
pre-fire juniper cover was upward of 40% resulted in major changes in plant community
composition. In addition, these plots burned at high severity, which resulted in loss of
shrubs and herbaceous vegetation as well. No juniper seedlings were detected on burned
plots in the first sampling period, but by the second sampling period, 12–13 years post-fire,
juniper seedlings were present on burned plots, but the abundance varied by woodland
development phase. No juniper seedlings were counted in burned Phase I woodlands, and
only 0.01 seedlings per square meter were counted in burned Phase II woodlands. Seedling
counts were higher in burned Phase III and Mature woodlands but still less than half the
count observed in unburned woodlands of the same phase. Wozniak and Strand [61] also

Fire 2023,6, 141 22 of 25
documented very low seedling counts 10 years after prescribed burns across woodland
development phases, ranging from 0.0001 seedlings per square meter in Phase I woodlands
to 0.002 seedlings per square meter in Phase III woodlands. The lower seedling counts
reported by Wozniak and Strand [
61
] can be attributed to the fact that their study design
excluded seedlings shorter than 5 cm. Given the fact that juniper seedlings were detected
on the site 12–13 years post-fire, this suggests that juniper is likely to return to the area, but
the juniper establishment will be slow.
5. Conclusions
Little information is available about effects of pre-fire vegetation on post-fire plant com-
munities in western juniper woodlands that burned in high intensity wildfire. Given that
the area burned in wildfire is increasing across ecosystems in the western US, this research
provides new and timely information. Our research demonstrates that early successional
stages of woodland development (Phase I and II) can be resilient to wildfire within ecologi-
cal sites classified as cool and moist soil temperature and moisture regimes [
41
,
62
], given
that perennial grasses and forb cover were above that of unburned plots, and sagebrush
was increasing in cover. Although annual grasses were present, they were not dominant,
but we noted that they increased in frequency. These results are encouraging and suggest
that prescribed fire or wildland fire use likely will have positive effects on plant community
richness, diversity and resilience to future fires on cool and moist mountain big sagebrush
ecological sites.
Plots with greater amounts of pre-fire western juniper cover had lower coverage of
most shrub and perennial grass species, with lower species diversity. Greater juniper
coverage also resulted in greater fire intensity and severity [
22
]. Thus, late-successional
plots had a diminished pre-fire understory but contained species and habitats such as
cavities that were uncommon in other stages. The understory was further reduced by the
greater fire intensity and severity resulting in lower species diversity. Much of the available
area, 35 and 38%, was occupied by Ceanothus 12 years post-fire in the Phase III and Mature
plots, respectively.
Perennial grasses were an important functional group in the Phase I and II plots
and increased to mean coverages of greater than 20% by the end of the study. Perennial
grasses also increased on burned Phase III plots but to a lesser degree than on the earlier
successional stages.
The species richness data show the importance of maintaining multiple examples
of all successional stages on the landscape in order to provide habitat for all potential
vascular plant species. The species richness at the landscape scale was approximately twice
that of any single successional stage. Similar results have been shown for birds and small
mammals [
46
]. A key issue in this consideration is the management and protection of
mature woodlands because they require very long time periods to develop (>500 years).
Although they do not commonly burn, this study demonstrates that they can burn under
extreme fire conditions. Consequently, some Phase III stands should be allowed to develop
into mature woodlands. This requires very long-term management strategies.
Supplementary Materials:
The following supporting information can be downloaded at:
https://www.mdpi.com/article/10.3390/fire6040141/s1, Table S1: Table of mean % cover and
standard deviation (in parenthesis) for the 14 most common species in the dataset..
Author Contributions:
Authors E.K.S. and S.C.B. have equally contributed to the manuscript includ-
ing original research idea, experimental design, data collection, analysis, writing and editing. All
authors have read and agreed to the published version of the manuscript.
Funding:
This research was partially funded by the U.S. Forest Service Rocky Mountain Research
Station grant number 12-JV-11221637-136.
Institutional Review Board Statement:
Not applicable because this study did not involve humans
or animals.

Fire 2023,6, 141 23 of 25
Data Availability Statement: Data are available by contacting the corresponding author.
Acknowledgments:
We would like to thank Penny Morgan, Emerita, Fire Ecology, University of
Idaho, and undergraduate and graduate students who helped with field sampling. We thank Steve
Jirik, Post Fire Recovery & Noxious Weed Program Lead, Bureau of Land Management, Idaho State
Office for review of an early draft of the manuscript.
Conflicts of Interest: The authors declare no conflict of interest.
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