Journal of Applied Ecology 2014, 52, 475–485
doi: 10.1111/1365-2664.12375
Consequences of habitat change and resource
selection specialization for population limitation in
cavity-nesting birds
Thomas E. Martin*
Montana Cooperative Wildlife Research Unit, U.S. Geological Survey, University of Montana, 205 Natural Science,
Missoula, MT 59812, USA
Summary
1. Resource selection specialization may increase vulnerability of populations to environmental
change. One environmental change that may negatively impact some populations is the
broad decline of quaking aspen Populus tremuloides, a preferred nest tree of cavity-nesting
organisms who are commonly limited by nest-site availability. However, the long-term consequences of this habitat change for cavity-nesting bird populations are poorly studied.
2. I counted densities of woody plants and eight cavity-nesting bird species over 29 years in
15 high-elevation riparian drainages in Arizona, USA. I also studied nest-tree use and specialization over time based on 4946 nests across species.
3. Aspen suffered a severe decline in availability over time, while understorey woody plants
and canopy deciduous trees also declined. The decline of plants resulted from increased elk
Cervus canadensis browsing linked to declining snowfall.
4. Woodpeckers exhibited very high specialization (>95% of nests) on aspen for nesting,
and densities of all six species declined with aspen over time. Mountain chickadees Poecile
gambeli and house wrens Troglodytes aedon exhibited increasingly less specialization on aspen.
Chickadees strongly increased in density over time, despite a relatively high specialization on
aspen. House wren densities declined moderately over time, but nest-box addition experiments
demonstrated that nest-site availability was not limiting their population. House wren densities increased with understorey vegetation recovery in elk exclosures via increased generality
of nest-site use, demonstrating that the decline in understorey vegetation on the broader landscape was the cause of their population decline.
5. Synthesis and applications. Management should target species that specialize in resource
selection on a declining resource. Species with greater resource selection generalization
can reduce population impacts of environmental change. Resource generalization can allow a
species like the wren to take advantage of habitat refuges, such as those provided by the elk
exclosures. Yet, resource generalization cannot offset the negative impacts of broad-scale
declines in habitat quality on the landscape, as demonstrated by the general decline of wrens.
Ultimately, aspen is an important habitat for biodiversity, and land management programmes
that protect and aid recovery of aspen habitats may be critical.
Key-words: cavity-nesting, climate change, habitat change, habitat selection, nest-site limitation, Paridae, population limitation, Troglodytes, ungulate browsing, woodpeckers
Introduction
Identifying specific elements of habitats that directly influence populations is important to allow targeted management
*Correspondence author. E-mail: tom.martin@umontana.edu
for conservation of populations and biodiversity (Martin
1992). The population impact of habitat change should
vary among species depending on differential habitat
requirements and habitat specialization; species with narrow, or specialized, habitat requirements are thought to
be most vulnerable to population problems from habitat
Published 2014. This article is a U.S. Government work and is in the public domain in the USA.
476 T. E. Martin
change (Warren et al. 2001; Harcourt, Copeton & Parks
2002; Korkeam€
aki & Suhonen 2002; Julliard, Jiguet &
Couvet 2004; Munday 2004).
One habitat that appears to be broadly changing is the
availability of quaking aspen Populus tremuloides. Area of
aspen habitat has declined by as much as 75% in many
places of the world, with climate and elk browsing being
important contributors (Romme et al. 1995; White, Olmsted & Kay 1998; Anderegg et al. 2012; Brodie et al.
2012; Martin & Maron 2012), as well as forest management (Latva-Karjanmaa, Penttil€a & Siitonen 2007). This
decline is predicted to increase into the future (Rehfeldt,
Ferguson & Crookston 2009) and is potentially critical
because aspen forests are considered biodiversity hotspots
(Tewksbury, Hejl & Martin 1998; Hansen et al. 2000;
Oaten & Larsen 2008).
One component of biodiversity that may be impacted
by the loss of aspen is cavity-nesting organisms because
aspen is often a preferred cavity-nesting tree (Harestad &
Keisker 1989; Li & Martin 1991; Kalcounis & Brigham
1998). Cavity-nesting organisms are commonly, although
not always, limited by availability of nesting cavities or
substrates for creating cavities (e.g. von Haartman 1957;
reviewed in Newton 1994). Given that (i) cavities or substrates commonly limit populations of cavity-nesting
organisms, (ii) aspen often provides preferred cavity
resources, and (iii) aspen is broadly declining, we might
expect a strong negative consequence for populations of
cavity-nesting species that prefer aspen for breeding sites.
Moreover, the negative impact of declining aspen abundance might be expected to be greater for species that
show greater specialization on aspen in their nest-site
selection.
Cavity-nesting bird species in high-elevation snowmelt
drainages in northern Arizona provide an interesting
example and test for this theory. Aspen, along with understorey deciduous tree recruits, have been strongly
declining in abundance in this system over the past quarter century due to increased browsing by elk as a result of
reduced snowfall, which allows elk to remain overwinter on these high-elevation sites (Martin 2007; Martin &
Maron 2012). Changes in the understorey may affect
foraging habitat for understorey foragers (e.g. house wrens
Troglodytes aedon) or habitat productivity, while loss of
aspen might affect nest-site limits. Cavity-nesting birds in
this system show a strong historical preference for nesting
in aspen (Li & Martin 1991). Here, I test whether: (i)
declining abundance of aspen was linked to population
declines in cavity-nesting birds, (ii) declines in abundances
of bird species were greater in species that showed stronger preferences for aspen, and (iii) environmental change
other than nest trees limited populations. I report changes
in aspen availability, nest-tree use, vegetation density and
associated population sizes of eight cavity-nesting bird
species across 29 study years. I also report experimental
tests of the environmental limits of population sizes
through two experiments: addition of nest boxes to test
nest-site (cavity) limitation for one species (house wrens)
and elk exclusion to allow understorey habitat recovery
and its possible effects on populations of all eight species.
Materials and methods
STUDY SITES AND SPECIES
This study was conducted in 15 high-elevation (c. 2350 m above
sea level) snow-melt drainages in Coconino National Forest of
northern Arizona, and each drainage represented one study plot.
The vegetation in these shallow drainages is characterized by a
mix of deciduous and coniferous tree species (Martin 2001),
including canyon maple Acer grandidentatum, quaking aspen
P. tremuloides, New Mexican locust Robinia neomexicana, Gambel’s oak Quercus gambelii, white fir Abies concolor and Douglas
fir Pseudotsuga menziesii. These drainages are narrow (100–150 m
wide) and contrast with the surrounding upland and drier forest
dominated by ponderosa pine Pinus ponderosa.
Eight cavity-nesting bird species were relatively common on
these sites. Six species were woodpeckers which usually excavate
their own holes, although reuse of old holes varies among species
(Martin 1993): red-shafted flicker Colaptes auratus cafer, rednaped sapsucker Sphyrapicus nuchalis, Williamson’s sapsucker
Sphyrapicus thyroideus, hairy woodpecker Picoides villosus,
downy woodpecker Picoides pubescens and acorn woodpecker
Melanerpes formicivorus. House wrens and mountain chickadees
Poecile gambeli depend on existing holes created by woodpeckers
or natural holes.
BIRD POPULATION SIZES, NEST SEARCHING AND
MONITORING
I counted numbers of singing, calling or drumming birds
throughout the length of these narrow habitats using the plotmapping technique (Christman 1984; Martin 2007). Songbirds
were censused from 1985, and woodpeckers were added to the
survey in 1987. The same nine plots were censused from midMay to mid-June each year from 1985 to 2013 for long-term bird
population trends, and four more plots were added for an elk exclosure experiment from 2004 to 2013 (see Fig. S1, Supporting
information). Censuses were not conducted in 1990 due to a lapse
in funding. Censuses were checked against territory maps made
by each nest searcher who visited the plots every other day
(Martin 2001), and if I missed a breeding pair detected for the
territory maps, I added that pair to my count, but this was rare.
Most censused plots were about 10 ha in size, but two were 6 ha,
and one was 4 ha. Bird abundance was standardized to numbers
of pairs per 10 ha across all plots.
Nests were found using parental behaviour, vocal cues such as
nest-leaving calls, and/or wood chips on the ground from excavation (Li & Martin 1991; Martin & Geupel 1993). Nests were
found and monitored on 15 plots (Fig. S1, Supporting information) every 3–4 days until offspring fledged or were depredated.
The tree species was identified for each nest.
VEGETATION SAMPLING
Densities of dominant woody plant species were measured in the
understorey at stratified random sites in a 5-m radius on each
Published 2014. This article is a U.S. Government work and is in the public domain in the USA., Journal of Applied Ecology, 52, 475–485
Population limitation in cavity-nesting birds
plot from 1987 to 2013, and larger trees were measured in 113m-radius plots from 1995 to 2013 (Martin 2001, 2007). A transect
of five points spaced equidistant up the side of the drainage was
established every 50 m on alternating sides for the length of the
drainage, yielding 30–60 sampling points per plot (Martin 2001,
2007). Vegetation sampling was not performed in 1997 because
fire danger closed the forest. Sampling points were not permanently marked (except on exclosures and their control pairs – see
later) causing differing points to be sampled every year. Understorey woody stems >20 cm tall were counted within the 5-mradius plots by woody species (Martin 2001, 2007). Canopy and
subcanopy trees (>15 cm d.b.h.) were counted in the 113-mradius plots by tree species. Snags with d.b.h. ≥ 20 cm were also
counted in the 113-m-radius plots by species.
Canopy aspen have been dying and falling at a steady rate
throughout this study with no recruitment for canopy replacement (Martin 2007). I implemented sampling in 1995 to document the decline in aspen. We permanently tagged all aspen
(dead or alive) and snags with d.b.h. ≥ 20 cm encountered within
25 m of the centre in each of the 15 drainages (878 total aspens;
897 other snags). Each year until 2011, field assistants resampled
all marked trees to determine whether they remained standing or
had fallen.
EXPERIMENTS
Tests of changing habitat structure due to elk browsing were conducted using exclosures. In autumn 2004, elk exclosures were
erected around one drainage (10 ha) in each of three canyons
(roughly 2–5 km apart, Fig. S1, Supporting information) and
paired with adjacent control drainages (Martin & Maron 2012).
The exclosures yielded extensive recovery of the understorey vegetation (Martin & Maron 2012) to allow a test of understorey
habitat quality on cavity-nesting populations.
Nest boxes were erected on four unfenced plots in 2012 prior
to nesting to examine whether nest sites limited house wren populations. Nest boxes were erected on a typical nest tree at 50-m
intervals along the length of plot and near the bottom of each of
the four drainages. House wren territories typically occupy about
100 m of the length of drainages (T. E. Martin, personal observation) such that the box density yielded two or three boxes per
potential territory.
STATISTICAL ANALYSES
Habitat change
I first tested differences in relative rates of loss of aspen among
three life classes (alive, partly alive, dead) characterized at the
time that trees were tagged in 1995. I tested for differences in the
relative rates of loss based on testing the interaction between life
class and year using a general linear model. I used SPSS (ver. 22;
IBM, Armonk, NY, USA) for all analyses, except when specified
differently.
Habitat change was relatively consistent over time (see Martin
2007; Martin & Maron 2012). As a result, I tested differences
between the two time endpoints (1995 and 2013) sampled with
large radius plots (113 m radius) to characterize general habitat
changes. I first tested changes in aspen and conifer (fir and pine
species summed) snags between these two time periods using a
general linear model. I next used multivariate general linear
477
models to examine the change in density of tree species in the
canopy (113-m-radius plots) and understorey (5-m-radius plots)
between the two time periods. I pooled the two pine species as
total pine and pooled Douglas fir and white fir as total fir. Pillai’s
Trace was used to test for significance of the multivariate test
because it is the most robust to model assumptions (Quinn &
Keough 2002).
Bird densities
Changes in densities (breeding pairs 10 ha 1) of the eight species
were examined across years using generalized linear models of
actual counts and plot size as a covariate.
I used a linear mixed model based on the Lme4 package in R
v3.0.3 for Windows (R Development Core Team, Vienna,
Austria) for a Before-After-Control-Impact (BACI) design to test
for differences in bird densities between elk exclosures (n = 3)
and control (n = 10) plots while accounting for repeated measures
for woodpeckers and mountain chickadees. For house wrens,
four of the 10 control plots served as box addition plots in the
later years, so I used the other six control plots for the elk exclosure test across all years for wrens. Treatment (exclosure vs. control) and a before–after factor (2004 = before; 2005–2013 = after)
served as fixed effects, year and plots as random factors (repeated
measures). I used the same linear mixed model BACI approach
to analyse the nest-box experiment, comparing house wren densities in nest-box addition (n = 4) to control (n = 6) plots in the
2 years before and the 2 years after nest boxes were added.
I examined potential habitat and competitor correlates of bird
densities using generalized linear modes (GLMs). House wrens
and mountain chickadees used natural cavities provided by cracks
in maple as the primary alternative to aspen. I used GLMs to
examine the trend in use of maple across years and relative to
conspecific density in control and exclosure plots.
Results
TEMPORAL CHANGE IN ASPEN AND SNAG
AVAILABILITY
The direct count of snags (standing dead trees) and aspen
in 1995 yielded 1775 standing individuals, with nearly half
(878 trees) being aspen, and fir snags represented the
other most abundant category of potential nest trees
(Fig. 1a). Of the aspen, about half (511%) were alive in
1995 and the rest divided between partly alive (225%,
some dead branches) and dead trees (264%) (Fig. 1a).
Aspen trees that were alive in 1995 fell at the slowest rate
over time (b = 226 043), with partly alive trees falling at a slightly faster rate (b SE = 293 043), and
trees that were dead in 1995 falling at the fastest rate
(b = 439 043) and leaving only 105% still standing
by 2011 (Fig. 1b; GLM with no intercept; life
class 9 year: F2,45 = 127, P < 0001; life class: F3,45 =
1200, P < 0001; year: F1,45 = 3278, P < 0001). Out of
all 878 aspen that were tagged in 1995, only 476%
remained standing by 2011 (Fig. 1b). Of these remaining
standing trees, 83% were still alive, 466% were partly
alive, and 451% were dead. In other words, only 36% of
Published 2014. This article is a U.S. Government work and is in the public domain in the USA., Journal of Applied Ecology, 52, 475–485
478 T. E. Martin
(a)
50
n = 1775 trees
Alive
30
20
Partly
Alive
10
Dead
0
Aspen
% of aspens
that remained standing
(b)
Oak
Maple
Fir
100
Pine
n = 878 aspens
60
1995
2013
P = 0·039
8
6
4
2
P < 0·001
ns
ns
P = 0·001
0
aspen maple locust
(b)
35
oak
fir
pine
P < 0·001
Understory
40
Alive in 1995
Partly alive in 1995
Dead in 1995
20
1995
No. of snags
(11 m radius circle)
10
80
0
(c)
No. of trees > 15 cm dbh
(11 m radius circle)
40
P < 0·001
12
Canopy trees
0·8
2000
2005
2010
1995, n = 338
2013, n = 386
0·6
No. of stems > 20 cm tall
(5 m radius circle)
Percentage of
snags and aspen in 1995
(a)
30
P = 0·001
15
P < 0·001
10
P = 0·003
5
P = 0·008
0
0·4
aspen maple locust
0·2
0·0
P = 0·013
Aspen
Conifer
Fig. 1. Availability of potential nest trees (aspen and snags) differed by tree type and changed over time. (a) A direct count of
all snags and aspen in 1995 in 50-m-wide transects down 15 plots
yielded 1775 trees distributed among five tree types. About half
were aspen that were alive, partly alive (some branches alive) or
dead (snag) trees. Fir snags were the next most abundant. (b)
The aspen trees from these transects were permanently tagged
(n = 878) in 1995, and the proportion that remained standing
declined over time at different rates depending on their life class
at the time of tagging in 1995. (c) The number of snags sampled
in 11-m-radius plots sampled systemically declined from 1995 to
2013 for both aspen and conifers.
the aspen trees tagged in 1995 remained alive in 2011,
with the remaining dying or falling. Thus, aspen trees and
snags have been disappearing from plots because of mortality, falling and no recruitment.
The rate that aspen trees fell exceeded the rate at which
alive and partly alive trees became snags. Density of both
aspen and conifer snags was lower in 2013 than in 1995,
and aspen snag density was less than conifer snag density
in both years (Fig. 1c; year: F1,1444 = 161, P < 0001;
snag type: F1,1444 = 274, P < 0001; interaction: F1,1444 =
0005, P = 09).
TEMPORAL CHANGE IN VEGETATION
The densities of canopy and subcanopy tree species changed from 1995 to 2013 (Pillai’s Trace = 022,
oak
fir
pine
Fig. 2. Densities of trees in the canopy (>15 cm d.b.h.) and understorey (>20 cm tall). (a) Densities of canopy trees sampled in
11-m-radius systematic plots, and (b) densities of understorey
woody stems sampled in 5-m-radius systematic plots. Statistical
significance is based on univariate analyses from multivariate
GLMs.
F6,717 = 336, P < 0001), but changes were restricted to
decreases in densities of aspen and maple, and increases
in canopy locust and fir from 1995 to 2013 (Fig. 2a). The
decrease in aspen was faster than maple, such that aspen
made up 191% of the density of these two species in
1995, but made up only 28% of the density of these two
canopy trees in 2013 (Fig. 2a).
The densities of woody recruits of canopy tree species
declined dramatically from 1995 to 2013 (Pillai’s
Trace = 038, F6,717 = 577, P < 0001), with all species
except pine decreasing in the understorey from 1995 to
2013 (Fig. 2b). Overall, the density of deciduous woody
stems declined dramatically in the understorey due to elk
browsing (Fig. 2b; also Martin & Maron 2012).
NEST-TREE USE
All woodpeckers used aspen for the majority of their nests
across all years. Woodpeckers together used aspen for
963% of nests (n = 2246 nests) over all years, while
mountain chickadees used them for 795% of nests overall
(n = 1136 nests), and house wrens used them for 664%
of nests overall (n = 1564 nests). Yet, proportional use of
aspen declined over time for all bird species (Fig. 3) as
Published 2014. This article is a U.S. Government work and is in the public domain in the USA., Journal of Applied Ecology, 52, 475–485
Population limitation in cavity-nesting birds
(a)
Percentage of nests
80
60
20
1985
Total woodpeckers; n = 2246 nests
Mountain Chickadee; n = 1136 nests
House Wren; n = 1564 nests
1990
1995
2000
2005
40 House Wren
Alive
Partly alive
Dead
30
20
10
0
aspen
2010
Fig. 3. Percentage of nests placed in aspen across years for woodpeckers, mountain chickadees and house wrens. Woodpeckers are
pooled because all woodpecker species place >90% of their nests
in aspen.
aspen declined (i.e. Fig. 1b). The three types of bird species differed in the rate of decline in use of aspen (GLM:
bird species 9 year: F2,72 = 238, P < 0001; bird species:
F2,72 = 233, P < 0001; year: F1,72 = 912, P < 0001).
House wrens showed the steepest decreases in use of
aspen
with time
(b = 163 0163; r = 090,
P < 0001). Mountain chickadees showed intermediate
decreases over time (b = 079 0205; r = 062,
P = 0001), and woodpeckers showed a very small decline
(b = 013 0045; r = 051, P = 0008) in the use of
aspen (Fig. 3).
Cavity nesters used alive and partly alive aspen as well
as aspen snags for nests (Fig. 4). House wrens and mountain chickadees placed nearly half (46% and 44%, respectively) of the aspen nests in alive or partly alive trees
(Fig. 4a,b). Nest holes in these trees were made by woodpeckers, and woodpecker species differed in their use of
alive vs. dead aspen (Fig. 4c). Red-naped sapsucker and
hairy woodpecker are strong excavators (Martin 1993),
and they excavated holes in alive and partly alive aspen
most frequently among the woodpecker species (Fig. 4c).
Their congeners (Williamson’s sapsucker and downy
woodpecker, respectively) are weaker excavators (Martin
1993) and depended extensively on dead aspen for nests
(Fig. 4c). Although northern flickers show some use of
alive and partly alive aspen (Fig. 4c), they do not excavate new holes in these life classes. Flickers are weak
excavators that commonly reuse existing holes (nearly
60%, Martin 1993) and by enlarging holes originally created by red-naped sapsuckers or hairy woodpeckers in
these life classes of aspen. They excavated new holes only
in dead aspen. The same was true of acorn woodpeckers.
House wrens and mountain chickadees used holes created by woodpeckers in aspen, but also used maple as
their secondary preference (Fig. 4a,b). Maple was the
dominant canopy tree in this system, although it has been
superseded by the combined abundance of Douglas fir
and white fir in recent years (Fig. 2a). Nonetheless, maple
(b)
Percentage of nests
40
maple
oak
maple
oak
50 Mountain Chickadee
40
30
20
10
0
aspen
(c)
Percentage of nests in aspen
% of nests in aspens
100
479
100
Woodpeckers
80
60
40
20
0
r
r
r
r
r
r
ke
ke
ke
ke
cke ucke
ec
ec
ec
uc
s
s
p
p
p
d
d
ap
ap
oo
er
oo
oo
dS
's S y W
rth
y W rn W
e
n
o
n
p
r
o
i
N
w
a
o
ms
Ha
Do
Ac
d-n
llia
Re
Wi
li
nF
Fig. 4. Percentage of nests in substrates classified by tree life class
[alive, partly alive (some dead branches), dead]. Percentage of
nests in the three main tree species used by life classes for (a)
house wrens and (b) mountain chickadees. (c) Percentage of nests
in aspen life classes for the six woodpecker species. Woodpeckers
tree use is represented only for aspen because aspen is used for
>90% of nests overall.
remains abundant in the canopy (Fig. 2a) and it develops
cracks and holes naturally while alive that are used by
both house wrens and mountain chickadees during all life
stages (Fig. 4a,b).
In all years, all cavity-nesting species chose aspen out
of proportion to its availability. Woodpeckers do not
excavate live trees of any tree species other than aspen in
Published 2014. This article is a U.S. Government work and is in the public domain in the USA., Journal of Applied Ecology, 52, 475–485
480 T. E. Martin
3
BIRD POPULATION TRENDS
Each of the six woodpecker species showed significant
population declines over the 27-year period of censuses
(Fig. 5a). Total woodpecker density across all species
exhibited a decline over time that was quite steep
(Fig. 5b). Indeed, the low density in 2010 was only 37%
of the peak (1992) recorded over 27 years (Fig. 5b). The
decline in density of each woodpecker species was significantly correlated with the (log-transformed) decline in
aspen density over the 18 years in which we sampled large
radius vegetation plots (r = 057–087, P = 0013 to
<0001 for the six species). The decline in total density of
woodpeckers was strongly correlated with the log-transformed decline in aspen, indicating a faster rate of decline
in woodpeckers as aspen became increasingly uncommon
(Fig. 5c). In contrast, mountain chickadees increased
strongly over the nearly three decades of censuses
(Fig. 6a). House wren density declined moderately over
the 29-year period of censuses, although the decline was
not constant (Fig. 6b).
Possible correlates of annual changes in densities of
mountain chickadees and house wrens were examined.
Mountain chickadees foraged primarily in the canopy on
the two fir species (T. E. Martin, personal observation),
2
1
0
(b)
Total woodpeckers
(pairs 10 ha—1)
8
6
4
2
b = —0·142 ± 0·0133, P < 0·001
1990
(c)
8
1995
2000
2005
2010
(a)
1995
2013
Mountain Chickadee
(pairs 10 ha–1)
Total woodpeckers
(pairs 10 ha—1)
7
6
5
4
3
0·0
5
4
3
2
log10 aspen: b ± se = 3·75 ± 0·298, r = 0·95, P
0·5
1·0
1·5
2·0
b ± se = 0·074 ± 0·0050, P < 0·001
2·5
Aspen density (trees per 11 m radius circle)
Fig. 5. Mean (SE) woodpecker density (pairs 10 ha 1) based on
censuses of the same nine plots over a 27-year period. Change in
density across time for (a) the six individual woodpecker species
and (b) all woodpeckers summed together. (c) The correlation
between total woodpecker density and aspen density measured in
11-m-radius plots starting in 1995. Each point represents 1 year.
this system and instead depend on dead trees of other species. Conifer snags were more abundant than aspen snags
(Fig. 1c), but conifer snags were used for a trivial percentage (<1%) of woodpecker nests. House wrens and mountain chickadees used aspen for more than half of their
nests (Fig. 3) even though canopy maple, their secondary
nest site (Fig. 4a,b), was substantially more abundant
(b)
6
House Wren
(pairs 10 ha–1)
Woodpecker species
(pairs 10 ha—1)
(a)
than aspen (Fig. 2a). Thus, aspen was clearly the preferred nest tree by all species.
Red-shafted Flicker; b = —0·037 ± 0·0061
Red-naped Sapsucker; b = —0·032 ± 0·0053
Williamson's Sapsucker; b = —0·027 ± 0·0057
Hairy Woodpecker; b = —0·017 ± 0·0039
Downy Woopecker; b = —0·015 ± 0·0041
Acorn Woodpecker; b = —0·014 ± 0·0042
5
4
3
2
b ± se = –0·045 ± 0·0090, P < 0·001
1985
1990
1995
2000
2005
2010
Fig. 6. Densities (pairs 10 ha 1) of (a) mountain chickadees
increased and (b) house wrens decreased relative to year based on
censuses of the same nine plots over a 29-year period.
Published 2014. This article is a U.S. Government work and is in the public domain in the USA., Journal of Applied Ecology, 52, 475–485
Population limitation in cavity-nesting birds
as also observed in other mixed habitat (Franzreb 1978). I
included densities of firs as foraging sites, house wren
densities as competitors for holes, and log-transformed
aspen density for nest-hole availability in a GLM. Aspen
density alone was the best model, with some support
(DAIC = 16) for an additional effect of fir density
(Table S1, Supporting information). Yet, the relationship
with aspen opposed expectations based on aspen limiting
nest-hole availability; chickadee density increased as aspen
became more limited in availability (b = 164 0324;
P < 0001). Woodpecker density, as possible hole creators,
was not included because chickadees increased as woodpecker densities declined, which also was opposite to
expectations based on nest-hole limitation. The increase in
chickadees with declining aspen may reflect the decreasing
density of deciduous trees in the canopy relative to the firs
on which they forage, as possibly suggested by the second
best model (Table S1, Supporting information). Indeed,
chickadee density was strongly correlated with the logtransformed ratio of fir to aspen density across years
(Fig. 7a), and this relationship provided the strongest
model (AICc = 24924; see Table S1, Supporting information).
For the decline in house wrens, I included total woodpecker density as nest-hole creators, density of mountain
chickadees because they can compete for holes, and total
density of deciduous woody stems in the understorey
Mountain chickadee density
(pairs 10 ha–1)
(a)
1995
2013
5
4
3
1·4
1·6
1·8
log10 Canopy fir density/aspen density
(11-m-radius plot)
1987
2013
5
4
3
0
10
20
30
40
50
60
70
80
0
90
10
0
12
0
b ± se = –0·015 ± 0·0046, P = 0·001
11
House wren density
(pairs 10 ha–1)
6
2
because it influences foraging habitat. Only understorey
woody stem density was supported for explaining the
decline in house wren density (Table S2, Supporting information; Fig. 7b). House wren density was not correlated
with the decline in aspen over the 18 years that aspen
density was measured (r = 004, P = 09).
RESPONSES OF VEGETATION AND CAVITY-NESTING
BIRDS TO ELK EXCLOSURE EXPERIMENTS
Understorey vegetation recovered strongly in elk exclosures (Martin & Maron 2012). By 2013, densities of understorey woody stems differed strongly between
exclosure and control drainages (n = 150 plots 9 two
treatments = 300 plots of 5 m radius) based on a multivariate GLM (Pillai’s Trace = 0217, F6,293 = 136,
P < 0001) that indicated that aspen, maple, locust and
oak were all more abundant in exclosures (all P < 0001),
while small firs (P = 054) and pines (P = 012) did not
differ between treatments. In contrast, none of the canopy
and subcanopy trees (aspen, maple, locust, oak, fir and
pine) differed in densities between the exclosure and control drainages (n = 300 plots of 113 m radius) in 2013
(Pillai’s Trace = 0029, F6,293 = 14, P = 020).
Elk exclosures did not affect densities of woodpeckers
(Fig. 8a; treatment 9 Before–After: t1070 = 045, P = 058;
treatment: t739 = 025, P = 080; Before–After: t103 =
118, P = 026) or mountain chickadees (Fig. 8b; treatment 9 Before–After: t1070 = 112, P = 027; treatment:
t1039 = 052, P = 060; Before–After: t112 = 201,
P = 007). House wrens exhibited a strong density response
to understorey vegetation recovery in elk exclosures
(Fig. 8c; treatment 9 Before–After: t709 = 489, P < 0001;
treatment: t492 = 070, P = 049; Before–After: t104 =
054, P = 060).
DENSITY RESPONSES OF HOUSE WRENS TO
2 b ± se = 1·68 ± 0·310, P < 0·001
0·4
0·6
0·8
1·0
1·2
(b)
481
Number of understory deciduous tree stems
(5-m-radius plot)
Fig. 7. Population correlates of (a) mountain chickadees with the
ratio of canopy firs to aspen tree densities and (b) house wrens
with understorey deciduous woody stem densities. Each point is
1 year.
EXPERIMENTAL INCREASES IN NEST HOLES
Nest boxes were added at a density of 13 boxes 10 ha 1
for the last 2 years of study, when aspen was least available (Figs 1b,c and 2a), to test whether nest sites limited
populations. Thirty-seven percentage of house wren pairs
nested in boxes, which represented 27% of available
boxes, on the four box-addition plots. House wrens also
used most of the available nest boxes in nearby meadow
habitat where natural cavities are largely unavailable,
although this habitat type is not included in this study.
Both sets of observations demonstrate that house wrens
readily use the nest boxes. Densities of house wrens, however, did not respond to the nest-box additions (Fig. 8d;
treatment 9 Before–After: t260 = 088, P = 039; treatment: t101 = 167, P = 013; Before–After: t23 = 016,
P = 089). Only 3–4 house wrens nested per box-addition
plot (Fig. 8d), leaving 9–10 boxes and large areas unoccupied per plot. These results indicate that availability of
nest holes was not limiting house wren densities.
Published 2014. This article is a U.S. Government work and is in the public domain in the USA., Journal of Applied Ecology, 52, 475–485
482 T. E. Martin
Total woodpecker density
(pairs 10 ha—1)
(a)
Mountain Chickadee density
(pairs 10 ha—1)
(b)
8
Exclosures erected
Control plots
Exclosure plots
4
2
0
6
5
4
3
2
1
0
10
House wren density
(pairs 10 ha—1)
CHICKADEES
House wrens decreased their use of aspen for nesting
across years (Fig. 3), as aspen density declined over years
(Figs 1b,c and 2a). The primary alternative nest site was
natural cavities offered by cracks in maple (Fig. 4a).
GLM results (Table S3, Supporting information) indicated that house wrens increased their proportional use of
maple across years (Fig. 9a) and with increased house
wren density (Fig. 9b) on control plots.
House wrens also increased their use of maple for nesting across years on exclosure plots (Fig. 9a). However,
use of maple in exclosures substantially exceeded use of
maple on control plots in the later years (Fig. 9a). The
much greater use of maple in exclosures was correlated
with the increased density (i.e. Fig. 8c) of house wrens in
exclosures (Fig. 9c), while accounting for year effects
(b = 324 0940) that reflect the decline in aspen availability (i.e. Fig. 1b, see above).
Mountain chickadees also decreased their use of aspen
across years, although to a lesser extent than house wrens
(Fig. 3). Like house wrens, maple was the primary alternative nest site (Fig. 4b). GLM analyses that included
year, chickadee density, house wren density and total
woodpecker density (Table S4, Supporting information)
yielded a best-fit model that only included year
(b = 0786 0197, P < 0001) for explaining use of trees
other than aspen for nesting (as in Fig. 3).
6
(c)
8
6
Discussion
4
2
0
(d)
Boxes
added
Control plots
Box addition plots
5
House wren density
(pairs 10 ha—1)
NEST-SITE USE AND DENSITY IN WRENS AND
4
3
2
1
0
2004
2006
2008
2010
2012
Fig. 8. Densities (pairs 10 ha 1) of (a) total woodpeckers, (b)
mountain chickadees and (c) house wrens on exclosures (fenced
to exclude elk in fall 2004, n = 3 plots) vs. control (n = 10 plots)
plots. (d) In the case of house wrens, boxes were added (nest
boxes added at 13 boxes 10 ha 1 in 2012, n = 4 plots), leaving six
control plots. Years prior to exclosure (2004) or box addition
(<2012) provide pre-treatment comparisons.
Aspen availability has declined dramatically in this system
across the last three decades of study (Figs 1 and 2), as
seen across North America and other parts of the world
associated with climate change and climate-induced elk
browsing, among other causes (Romme et al. 1995;
White, Olmsted & Kay 1998; Latva-Karjanmaa, Penttil€
a
& Siitonen 2007; Rehfeldt, Ferguson & Crookston 2009;
Brodie et al. 2012; Martin & Maron 2012). The loss of
aspen in this mixed forest system has had very different
impacts on populations of cavity-nesting bird species.
The most negative impact was on woodpeckers (Fig. 5).
All six woodpecker species showed very high specialization on aspen for nesting (>95% of nests, Fig. 3), and all
showed significant population declines over time, correlated with the decline in aspen availability (Fig. 5). Total
woodpecker density declined to nearly one-third of the
peak density. The population decline became steeper relative to aspen availability as aspen availability became
increasingly limited (Fig. 5c). Other aspects of the habitat,
such as understorey deciduous woody stem densities, also
declined over time (Fig. 2, Martin & Maron 2012). Yet,
this habitat change clearly did not contribute to the
woodpecker decline given that recovery of the understorey
Published 2014. This article is a U.S. Government work and is in the public domain in the USA., Journal of Applied Ecology, 52, 475–485
Population limitation in cavity-nesting birds
100
80
(c)
Control
Exclosures
60
40
20
0 b ± se = 1·41 ± 0·128, P < 0·001
1985
1990
1995
2000
2005
Control
Exclosures
40
b ± se = 6·18 ± 1·575,
P < 0·001
Relative % of wren nests
in maple
% of wren nests in maple
(a)
483
20
0
–20
b ± se = 5·93 ± 1·014, P < 0·001
–40
2010
–4
Relative house wren density (pairs 10 ha–1)
–2
0
2
4
Relative house wren density (pairs 10 ha–1)
Relative % of wren nests in maple
(b)
15
10
5
0
–5
–10
b ± se = 4·46 ± 1·118, P < 0·001
–1·5
–1·0
–0·5
0·0
0·5
1·0
1·5
Relative house wren density (pairs 10 ha–1)
Fig. 9. Use of maple for nesting by house wrens over time and relative to house wren density on control and exclosure plots. (a) Aspen
was preferred in early years (Fig. 3), but house wrens increasingly used canyon maple across years on control and exclosure (2005–2013)
plots. (b) House wrens increased their use of maple for nesting as relative (i.e. controlled for year effects) house wren density increased
on control plots. (c) House wrens increased their use of maple for nesting as relative (i.e. controlled for year effects) house wren density
increased across control and exclosure plots during the period after exclosures were erected (2005–2013).
in elk exclosures, when canopy trees like aspen did not
change, had no effect on woodpecker densities (Fig. 8).
Thus, loss of aspen in these high-elevation riparian systems that are embedded in a sea of pine upland has serious negative consequences for diverse woodpecker
populations that exhibit high specialization on aspen for
nest sites.
Mountain chickadees, on the other hand, increased
strongly over time (Fig. 6a), despite their preferred nest
tree (aspen) declining in availability over this time period
(Figs 1 and 2). This increase may reflect an increase in the
absolute (Fig. 2a) and relative (Fig. 7a) density of firs,
their preferred foraging substrate. Reduced competition
for nest holes because of declining house wren densities
(i.e. Fig. 6b) might also have contributed to the population
increase. Chickadee and wren densities were negatively
correlated (r = 050, P = 0007) across years. Yet, both
species were sometimes observed nesting in different holes
in the same aspen tree suggesting that competition for
holes between them might be weak. The increase in chickadee densities over time demonstrated that nest-hole availability was not limiting their populations, likely facilitated
by their decreased use of aspen (Fig. 3) and increased use
of maple and other nest sites (Fig. 4).
The increased use of maple nest sites by house wrens
across years (Fig. 9a) as aspen declined (Figs 1 and 2)
and in years of higher population density (Fig. 9b)
together indicated that aspen was increasingly limited in
availability for nesting over time. This increased use of
maple trees for nesting was not a result of increasing
numbers of maple trees, as maple also declined over time
(Fig. 2). Instead, wrens were clearly general in their nestsite use (Johnson 1998) and increased their use of maple
as aspen became more limited in availability. Ultimately,
plasticity in nest-site selection by house wrens prevented
declining aspen from limiting populations, as demonstrated by the lack of population responses to nest-box
additions (Fig. 8d).
Nest-site plasticity also allowed house wrens to take
advantage of recovered habitat conditions created by elk
exclusion. Populations strongly increased on exclosure
plots (Fig. 8c) despite no increase in cavity availability;
only young plants, which do not provide cavities,
increased on exclosures (Figs 1d and 2; Martin & Maron
Published 2014. This article is a U.S. Government work and is in the public domain in the USA., Journal of Applied Ecology, 52, 475–485
484 T. E. Martin
2012). Instead, house wrens achieved higher densities on
exclosures by increasing their use of maple for nesting to
the point that maple use far exceeded use of aspen on exclosures in contrast to control plots (Fig. 9a). This greater
use of maple reflected both the decline and low availability of aspen across years (i.e. Figs 1 and 2) and the much
greater density of house wrens on exclosures (Figs 8c and
9c). In short, their nest-site plasticity allowed increased
population densities to take advantage of the presumably
higher quality habitat in elk exclosures despite few aspen
nest sites (Fig. 3), further verifying that resource selection generality prevented cavity availability from limiting
populations.
These results demonstrate two points of importance for
land managers. First, it is important to experimentally test
environmental causes of population limits. It would have
been easy to conclude that declining aspen was the cause
of house wren declines, given that aspen was their preferred nest site and nest-site availability commonly limits
populations of cavity-nesting organisms (see Newton
1994). As a result, the coarsely correlated decline of wrens
with understorey foraging habitat (Fig. 7b) might have
been overlooked. Yet, the elk exclosure and nest-box
addition experiments demonstrated that understorey foraging habitat was the clear limit on populations. Nest-site
selection plasticity ameliorated one potential population
constraint arising from habitat change (i.e. reduced aspen
cavities), but the population was still limited by broader
habitat degradation associated with climate change and
elk browsing. Thus, careful experiments that ascertain the
environmental factors that limit populations can strongly
benefit targeted management actions.
Secondly, the degree of resource selection specialization
and spatial scale of habitat selection had important, but
not universal, consequences for population responses to
climate-induced habitat change. Woodpeckers in this system, with their very high specialization on aspen, showed
very strong population declines with the decline in their
preferred nest tree. Yet, specialization was not a perfect
predictor of population trends. Mountain chickadees had
intermediate, but still relatively high, specialization on
aspen for nesting and their populations still increased
despite the decrease in aspen. Mountain chickadees and
house wrens apparently were able to avoid nest-site limits
on their populations through more general nest-site use.
In the case of house wrens, changes in the understorey
created limits on their populations. These limits, however,
may be ameliorated by differential resource selection at
the larger habitat scale of plots (i.e. individual snowmelt
drainages). House wrens preferentially selected the higher
quality habitat represented by elk exclosures over lower
quality habitat represented by unfenced plots (Fig. 8c).
Such resource selectivity can reduce broader population
impacts of environmental change by increasing the proportion of the population in high-quality habitat. Yet,
species can differ in their ability to discern and respond to
variation in habitat quality (Lloyd et al. 2005). For exam-
ple, ovenbirds Seiurus aurocapillus reduce their use of ecological traps associated with landscape fragmentation,
while wood thrush Hylocichla mustelina appear to be
poorer at making such resource selection distinctions, and
these differences in resource selectivity are associated with
large differences in their population trajectories at regional scales (Lloyd et al. 2005). Such differences in resource
selectivity along with differences in plasticity will yield
large differences in adaptability of species to rapid and
broad environmental change associated with climate and
landscape changes. At the same time, the ability of
resource selectivity to ameliorate degrading environmental
conditions will also depend on good quality habitat being
available for occupation by the broader population.
In aspen forests, management actions that protect and
bolster recruitment of aspen, to replace dying canopy
trees, may be critical for creating such quality habitat and
mitigating the negative effects of aspen loss on biodiversity. Moreover, the differential use of differing life classes
for nesting by differing bird species (Fig. 4) emphasizes
the need for transitional age/life classes and the importance of managing for shifting age classes.
Ultimately, understanding the potential population
impacts and sensitivity of species to environmental change
will depend on understanding their specialization in
resource needs. Species, such as woodpeckers, that have
high resource use specialization are clearly vulnerable to
declines in that resource (Figs 3 and 5). Species that are
more general in resource selection, such as mountain
chickadees and house wrens, may be better able to withstand environmental changes. Yet, even strong resource
use generality may be insufficient to offset the population
impacts of large environmental degradation, such as the
broad habitat change seen here, if suitable alternative
habitat is not available. Understanding the mechanisms of
population limitations and adaptability of species to environmental change needs to be a critical focus of future
studies. More importantly, land management programmes
that target protection and restoration of aspen habitats
are critical given their importance to cavity-nesting organisms and biodiversity.
Acknowledgements
I thank J. C. Oteyza, J. LaManna, P. Llambıas, J. Mouton, J. Rivers and
two anonymous reviewers for helpful comments. I also thank numerous
research assistants for help in locating and monitoring nests and measuring
vegetation, as well as the U.S. Forest Service Blue Ridge Ranger Station
for logistical support. This research was supported by grants from the U.S.
Geological Survey Climate Change Research Program and the National
Science Foundation (DEB-1241041, IOS-1349178). This work was conducted under University of Montana IACUC #059-10TMMCWRU. Any
use of trade, firm or product names is for descriptive purposes only and
does not imply endorsement by the U.S. Government.
Data accessibility
Data are available from the Dryad Digital Repository: http://dx.doi.org/
10.5061/dryad.tr25v (Martin 2014).
Published 2014. This article is a U.S. Government work and is in the public domain in the USA., Journal of Applied Ecology, 52, 475–485
Population limitation in cavity-nesting birds
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Supporting Information
Additional Supporting Information may be found in the online version
of this article.
Fig. S1. Map of study plots, drawn to scale.
Table S1. Generalized linear model results for explaining annual
variation in density of mountain chickadees.
Table S2. Generalized linear model results for explaining annual
variation in density of house wrens.
Table S3. Generalized linear model results for explaining annual
variation in proportional use of maple for nesting by house
wrens.
Table S4. Generalized linear model results for explaining annual
variation in proportional use of maple for nesting by mountain
chickadees.
Published 2014. This article is a U.S. Government work and is in the public domain in the USA., Journal of Applied Ecology, 52, 475–485