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2017 NORTHEASTERN NATURALIST 24(3):267–288
Forest Bird Populations in Massachusetts:
Breeding Habitat Loss and Other Influences
Timothy J. Gardner1, Caroline R. Eagan1, and Robert I. Bertin1,*
Abstract - Our objective was to determine whether changes in populations of forest-interior
bird species were related to changes in extent of interior forest along Breeding Bird Survey
(BBS) census routes in Massachusetts. We first identified a suite of 28 forest-interior bird
species (FIA species), based on correlations between bird abundance (in 2003–2007) and extent
of interior forest (in 2005) along BBS routes. From this group, we eliminated 13 species
whose breeding habitats were described in the literature as including forest edge or second
growth, resulting in a more stringently defined subset of 15 (FIB) species. We quantified
the extent of forest and interior forest (>100 m from a forest edge) along BBS routes based
on digitized aerial photographs from 1971, 1985, and 1999. We also quantified changes in
abundance of the 28 forest bird species along BBS survey routes over the same time period.
Overall, changes in abundance of FIB species paralleled changes in extent of interior forest,
with 13 of 15 species showing positive correlations, 5 of which were significant. However,
substantial variation occurred among species, including conspicuous declines in Hylocichla
mustelina (Wood Thrush) and Piranga olivacea (Scarlet Tanager) and conspicuous increases
in Vireo solitarius (Blue-headed Vireo) and Setophaga coronata (Yellow-rumped Warbler).
Changes were not significantly related to either migratory status (Neotropical vs. other) or
nest location (ground vs. arboreal). Several differences could be attributed to species-specific
factors, such as reintroductions of Meleagris gallopavo (Wild Turkey) and Corvus corax
(Common Raven) or introduction of competitors, such as Haemorhous mexicanus (House
Finch) impacting Haemorhous purpureus (Purple Finch). Changes in some bird populations
seem to reflect forest succession, e.g., Hylatomus pileatus (Pileated Woodpecker), while others
are unexplained and may be due to changes on migratory routes or wintering grounds.
Overall, loss of interior forest is an important incremental factor in forest bird population declines,
although other factors had a greater impact in the period under study.
Introduction
Declines in abundance of forest bird species (particularly Neotropical migrants)
in eastern North America have been noted in numerous studies in recent decades
(Buchanan et al. 2016, Hall 1984, Sauer and Link 2011, Terborgh 1992, Whitcomb
et al. 1981). Because the extent and contiguity of forests in many parts of eastern
North America have also declined during this period, a causative relationship
has been suggested (Askins et al. 1990, Donovan and Flather 2002, Lynch and
Whigham 1984, Parker et al. 2005, Robinson and Wilcove 1994, Suarez-Rubio
and Lookingbill 2016), although it is widely recognized that habitat changes on the
wintering and migratory grounds or other factors might also be involved (Keller
and Yahner 2006, Robbins et al. 1989, Taylor and Stutchbury 2016).
1Biology Department, College of the Holy Cross, Worcester, MA 01610. *Corresponding
author - rbertin@holycross.edu.
Manuscript Editor: Peter Paton
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While a regional decline in populations of a given species in an area that is
undergoing widespread habitat loss is certainly consistent with a causative role for
such habitat losses in the population decline, it does not rule out other factors. More
compelling evidence can be provided if an association between these 2 variables
can be demonstrated in an area where both the extent of habitat loss and changes
in bird populations vary. Such analyses have become practical with the widespread
availability of long-term bird-census data (Sauer et al. 2014) and high-resolution
digitized land-use data covering the same time period.
Human activities in forested areas tend not only to reduce the extent of forest
but also fragment the remaining forest parcels. Fragmentation increases the ratio
of edge to interior forest, and forest edge differs from interior in various physical
factors and vegetation characteristics (Chen et al. 1993, Matlack and Litvaitis
1999, Williams-Linera 1990). In at least some circumstances, nest predation and
parasitism are also greater along forest edges than in interior forest (Brittingham
and Temple 1983, Chalfoun et al. 2002, Donovan et al. 1997, Gering and Blair
1999, Phillips et al. 2005). Paton (1994) concluded that edge-related increases in
nest predation extend less than 50 m into forest in most circumstances, although
some studies suggest edge effects extend greater distances (Brittingham and Temple
1983, Wilcove et al. 1986).
We use “forest-interior” birds to refer to species associated with large forest
tracts, realizing that such associations might reflect either area sensitivity or edge
avoidance (Parker et al. 2005, Villard 1998). We also note that designations of
forest-interior birds sometimes differ among studies (Askins et al. 1987, Dunford
and Freemark 2004, Mancke and Gavin 2000, Phillips et al. 2005, Villard 1998),
making empirical determination of such status desirable. Variation in responses of
forest-interior species to fragmentation have been noted, reflecting factors such
as migratory status, nesting location (ground vs. arboreal) and landscape context
(Chalfoun et al. 2002, Dunford and Freemark 2004, Hagan and Meehan 2002,
Lee et al. 2002, Lindenmayer et al. 2002, Sauer et al. 1996). Given these varied
responses, studies in multiple geographic areas and in different landscape contexts
(Richmond et al. 2012, Thompson et al. 2002) are essential to develop a comprehensive
picture of the effects of forest loss and fragmentation on forest birds.
New England forests have undergone dramatic changes in the past several hundred
years. As agriculture spread across the landscape, unbroken forest was reduced
in area and fragmented. In Massachusetts, deforestation peaked around 1860, when
nearly 70% of forest had been cleared (O’Keefe and Foster 1998). Abandonment
of agricultural land followed, with forest areas increasing in area and maturity and
becoming less fragmented. More recently, reduction and fragmentation of forests
have again increased, driven in most areas by increased use of land for residential
and commercial purposes and for infrastructure such as roads and utility corridors.
This study examines patterns of change in forest-interior bird species in Massachusetts
in relation to changes in the extent of forest and interior forest. We use
bird data from the Breeding Bird Survey (USGS 2016) and land-use data from
digitized aerial photographs from MassGIS (2016). Unlike some previous studies,
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we evaluated land-use changes in the specific zone of detection along the bird
census routes. By examining the association between changes in bird abundance
and extent of interior forest, we sought to shed light on the likely role of destruction
and fragmentation of breeding habitat vs. other factors in causing changes in
forest bird populations.
Our specific goals were to (1) identify a suite of forest-interior bird species
for Massachusetts, (2) estimate changes in forest and forest-interior habitat along
Breeding Bird Survey (BBS) routes in Massachusetts, (3) estimate changes in densities
of forest-interior birds and avian-nest predators and parasites along the same
BBS routes, (4) determine whether temporal changes in bird abundances and habitat
variables are correlated, and (5) determine whether changes in forest-interior
bird species are influenced by the regional habitat context.
Methods
Data sources
Data on bird abundances came from the North American Breeding Bird Survey
(USGS 2016). This survey has been conducted on hundreds of routes throughout
the country, starting in 1966. Routes are sampled by competent observers during
the height of the breeding season in June. Starting 0.5 h before sunrise, the observer
drives the route, stopping at 50 points 0.8 km apart. During each 3-minute stop,
the observer records all bird species heard as well as those seen within 400 m of
the sampling point (Sauer et al. 2014). We obtained data for the 27 Massachusetts
routes that had been surveyed over a period of at least 10 years each. Our measure
of bird abundance was the total number of individuals of a species recorded at all
stops on a route in a particular year.
We obtained data on land use along each bird route from a website maintained
by the Commonwealth of Massachusetts (MassGIS 2016). Land-use layers on this
website are based on digitized aerial photographs taken in 1971, 1985, 1999, and
2005. We used land-use categories 3 (forest) and 37 (forested wetland) to represent
forest. The Commonwealth hand-digitized land-use polygons in 1971, 1985, and
1999, but then switched to semi-automated methods in 2005, causing a slightly
more liberal interpretation of forest in 2005 than in preceding surveys. Thus, while
comparisons of land use among sites in any sampling year were valid, comparisons
of land use in 2005 with earlier years may be subject to error. Accordingly we relied
on the 1971–1999 data for assessing changes in forest area along individual routes.
Interior forest
We defined interior forest as forest at least 100 m from an edge bordering nonforested
habitat (e.g., Bayne and Hobson 1997), a distance that should exclude
edge-associated modifications in physical and vegetation variables and most
depredations of edge-associated species like corvids, cowbirds, or mammalian
predators. Defining interior forest using a larger buffer would cause very little
habitat in eastern Massachusetts to qualify as interior forest, despite the presence
in this area of breeding populations of forest-interior bird species.
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Identifying focal bird species
Because decisions about whether a bird species is a forest-interior species vary
in previous studies, we defined these groups ourselves, using empirical data from
Massachusetts supplemented by published habitat descriptions. The empirical
approach involved assessing correlations between bird abundance and extent of
interior forest along Massachusetts BBS routes. We began with all bird species
recorded along Massachusetts BBS routes. We deleted obvious non-forest species,
including those associated with grasslands, open habitat, and urban areas, as well as
shorebirds, waders, waterfowl, aerial insectivores, and (because they are difficult to
survey using BBS methods) hawks and owls. To minimize sampling error, we also
eliminated species recorded on fewer than 5 routes. We used 2005 habitat information
in this analysis, and included all 19 Massachusetts BBS routes that had data
for at least 3 of the 5 years from 2003 to 2007, to minimize the chance fluctuations
involved in a single year of data. For each bird route, we obtained the average number
of individuals detected along the route per year. We also calculated the extent
of interior forest within 400 m of the route (and presumably therefore within the
detection zone of the observers). We calculated a correlation coefficient between
these 2 variables across the 19 routes for each bird species. We also performed correlations
using interior forest within 200 m and 100 m of the route to guard against
the possibility that the detection zone was smaller than the nominal 400 m. The
trends were similar using these smaller buffers, but correlations were consistently
lower than when using 400 m, thus supporting the use of the 400-m detection band.
Species whose abundance showed significant positive correlation with the extent
of interior forest we designated FIA species (Table 1). We reviewed habitat
descriptions of each species in The Birds of North America (Rodewald 2015) and
eliminated from the list all species whose breeding habitats regularly extend into
forest edge or second growth. The species on this smaller list we designated FIB
species (Table 1). In addition, we tracked 3 nest predators (Corvus brachyrhynchos
[American Crow], Cyanocitta cristata [Blue Jay], Quiscalus quiscula [Common
Grackle]; see Table 1 for authorities) and 1 nest parasite (Molothrus ater [Brownheaded
Cowbird]), here referred to as agonistic species. Migratory status of each
species (permanent resident, short-distance migrant, Neotropical migrant) was determined
from Freemark and Collins (1992).
Landscape context
We assessed the relevance of landscape context to our analysis by examining
land use at 2 distances beyond the 400-m detection radius, also using 2005 land-use
data: 5 km and 10 km. However, the extent of forest in these wider bands around
the BBS routes generally was not useful in explaining changes in bird abundance,
and these data are not discussed further.
Extracting forest-cover data
We copied relevant shapefiles from MassGIS (2016) into ArcMap 10.1. For each
bird route in each of the 4 years (1971, 1985, 1999 and 2005), we created a layer of
forest habitat. Internal boundaries between adjacent forest parcels were dissolved
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Table 1. Focal bird species. Forest Interior B (FIB) species are those whose abundance is significantly correlated with forest cover in Massachusetts and
which Rodewald (2015) treats as a forest species. Other Forest Interior species are those whose abundance was significantly correlated with forest cover
in Massachusetts but which Rodewald (2015) indicates also frequent non-forest habitat. These species, together with FIB species, comprise FIA species.
Agonist species are those likely to pose problems for nesting species due to nest predation or brood parasitism. Sp ch refers to the change in bird abundance
between 1971 and 1999 as detected by BBS censuses. See Methods for formula. 2005 IF reports the correlation between 2003-2007 bird abundance and
extent of interior forest within 400 m of the survey route using 2005 land cover data. Change F and Change IF report the correlation between change in
bird abundance from 1971 to 1999 and change in extent of forest (F) or interior forest (IF), respectively, within 400 m of the survey route. Significant correlation
coefficients are denoted with asterisks. Neotropical migrants are noted with † and ground nesting species with ‡. [Table continued on next page.]
Common name Latin name Sp ch 2005 IF Change F Change IF
Forest Interior B Species
Black-and-white Warbler†‡ Mniotilta varia L. -18.2 0.524* 0.168 0.377
Blackburnian Warbler† Setophaga fusca Müller -13.4 0.673* 0.550 0.792*
Black-throated Blue Warbler† Setophaga caerulescens Gmelin -5.7 0.801* 0.106 0.332
Black-throated Green Warbler† Setophaga virens Gmelin 26.4 0.616* 0.403 0.572*
Blue-headed Vireo† Vireo solitarius Wilson 56.1 0.804* -0.067 -0.273
Dark-eyed Junco‡ Junco hyemalis L. -36.1 0.619* 0.774 0.890*
Hairy Woodpecker Picoides villosus L. +3.0 0.677* -0.087 0.005
Hermit Thrush‡ Catharus guttatus Pallas +8.7 0.652* 0.408 0.078
Ovenbird†‡ Seiurus aurocapilla L. +5.4 0.793* 0.508* 0.580*
Pileated Woodpecker Hylatomus pileatus L. +45.3 0.667* -0.314 -0.302
Red-eyed Vireo† Vireo olivaceus L. -18.7 0.645* 0.052 0.412
Scarlet Tanager† Piranga olivacea Gmelin -21.8 0.758* -0.163 0.208
Winter Wren‡ Troglodytes hiemalis Viellot +40.6 0.672* 0.501 0.426
Wood Thrush† Hylocichla mustelina Gmelin -41.7 0.661* -0.059 0.230
Yellow-rumped Warbler Setophaga coronata L. +58.8 0.705* 0.845* 0.812*
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Table 1, continued.
Common name Latin name Sp ch 2005 IF Change F Change IF
Other Forest Species
American Redstart† Setophaga ruticilla L. +6.7 0.701* 0.169 0.334
Common Raven Corvus corax L. +100.0 0.586* -0.024 -0.234
Least Flycatcher† Empidonax minimus Baird -55.1 0.612* -0.139 0.120
Magnolia Warbler† Setophaga magnolia Wilson +33.2 0.645* 0.338 0.813*
Nashville Warbler†‡ Oreothlypis ruficapilla Wilson -32.3 0.682* 0.114 0.358
Northern Waterthrush†‡ Parkesia noveboracensis Gmelin +15.3 0.566* 0.418 -0.212
Purple Finch Haemorhous purpureus Gmelin -50.4 0.649* 0.417 0.413
Rose-breasted Grosbeak† Pheucticus ludovicianus L. -10.2 0.753* 0.052 0.444
Ruby-throated Hummingbird† Archilochus colubris L. +85.7 0.730* -0.346 -0.227
Veery† Catharus fuscescens Stephens -1.9 0.897* -0.198 0.219
White-throated Sparrow‡ Zonotrichia albicollis Gmelin -78.0 0.625* -0.514 -0.092
Wild Turkey‡ Meleagris gallopavo L. +98.3 0.510* -0.074 -0.610
Yellow-bellied Sapsucker Sphyrapicus varius L. +22.1 0.505* 0.012 -0.009
Agonist Species
American Crow Corvus brachyrhynchos Brehm +36.9 0.071 -0.133 -0.261
Blue Jay Cyanocitta cristata L. -21.4 0.171 0.427 -0.075
Brown-headed Cowbird Molothrus ater Boddaert +5.1 -0.461 -0.079 0.246
Common Grackle Quiscalus quiscula L. -16.4 -0.544 -0.024 -0.349
Other Species Mentioned
House Finch Haemorhous mexicanus Müller
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so that these boundaries were not detected as forest edge. We created a forest layer
for each bird route and year by clipping the forest layer to within 400 m of the bird
route. A forest-interior layer was also created for each bird route and year by applying
a 100-m interior buffer to all forest patches in the dissolved forest layer and
clipping to within 400 m of the bird route.
Trends in forest cover and bird abundance
We calculated trends in interior forest along each bird route by regressing the
area of interior cover on date, using data for the years 1971, 1985, and 1999. The
regression coefficient was taken as the average annual rate of change in interior
forest over this period.
To prepare bird data for analysis, we first entered zero values for years in which
the route had been sampled but the bird species was not recorded. We analyzed
trends in bird abundance in relation to changes in extent of interior forest in 2 ways.
In one, we used only bird data from two 5-year time blocks, 1969–1973 and 1997–
2001. These years were chosen because they were centered on the earliest and latest
years for which we had comparable forest cover data (1971 and 1999, respectively).
We included only routes containing bird census data for at least 3 years in each time
block, to minimize sampling error. Although percent change would be an intuitive
measure of abundance change, the presence of zero abundances for several species
on particular routes during the earlier time period meant that this quantity could
not always be calculated. Thus, instead of using the 1969–1973 abundance in the
denominator, we used the average of the 1969–1973 and 1997–2001 abundances,
leading to the formula [N99 - N71] x 50) / ([N99 + N71] / 2), where N71 and N99 are the
average number of birds of that species recorded along the route in 1969–1973 and
1997–2001, respectively. The 50 in the numerator causes the theoretical range of
this index to lie between +100 and -100. Routes where a species was absent in both
time periods were excluded from the calculation. To be consistent, we calculated
an index of change in forest and an index of change in interior forest based on forest
cover data from 1971 and 1999 in the same manner. Each index was calculated
as ([F99 - F71] x 50) / ([F99 - F71] / 2), where F99 and F71 refer to the extent of forest
(or interior forest) in the years 1999 and 1971, respectively. We then performed
a correlation between the index of forest change and the index of change in bird
abundance for each bird species.
Our second approach involved regressing bird abundance on year for each bird
species on each route for the years 1967–2003, using the resulting regression coefficient
as the annual rate of change in bird abundance. To minimize sampling error,
we used only routes that had been censused a minimum of 10 times and bird species
whose presence had been recorded in a minimum of 5 censuses. We then calculated
a product-moment correlation coefficient between the regression coefficients for
bird abundance and the regression coefficients for extent of interior forest. Positive
correlations thus indicated an association between change in bird abundance and
change in extent of interior forest. In addition to examining the significance of individual
correlation coefficients, we sought evidence of an overall (across-species)
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pattern in the data by using a binomial test to determine whether positive correlation
coefficients were more frequent than expected by chance under the null hypothesis
that positive and negative coefficients were equally likely.
Finally, we made an independent assessment of changes in populations of focal
bird species using information from the one other comprehensive set of bird surveys
in Massachusetts over the past several decades. This work involved a pair of
breeding bird atlases (hereafter BBA) completed under the auspices of MassAudubon
(formerly the Massachusetts Audubon Society) in 1974–1979 and 2007–2011
(MassAudubon 2016). The atlases were based on occurrence data in roughly 1000
blocks distributed throughout the state, each consisting of 1/6th of a United States
Geological Survey 7.5-minute topographic map, about 25 km2 (10 mi2). Based on
block occupancy, population trends for different species in this study were reported
in 1 of 5 categories: strongly decreasing, likely decreasing, stable, likely increasing,
and strongly increasing. To assess the correspondence of these results with BBS
results, these 5 categories were assigned numbers -2, -1, 0, +1, +2, respectively, and
a nonparametric correlation (Spearman’s rho) was performed between these ranks
and the change indexes calculated for the 28 FIA bird species.
Forest extent and human population size
Because trends in interior forest seemed likely to be associated with changes in
human population, we sought to quantify the relationship between these 2 variables.
We obtained decadal county population data for 1970 and 2000 from the National
Bureau of Economic Research (NBER 2016). We then assigned each bird route to
1–3 counties as follows. If at least 90% of the route was confined to 1 county, then
the population change of this county was recorded. If the bird route entered several
counties, the average population change for those counties containing at least 10%
of the route was assigned to the route. We then ran a correlation between change
in interior forest between 1971 and 1999 and change in human population between
1970 and 2000 across all bird routes.
Results
Focal bird species
A total of 28 bird species (designated FIA species) showed significant positive correlations
with extent of interior forest within 400 m of the 19 BBS routes (Table 2).
Habitat descriptions in Rodewald (2015) suggested that 13 of these are often found
in habitats other than forest, such as thickets, gardens, and forest edges. Eliminating
these left 15 (FIB) species that both exhibited a significant correlation with extent of
forest interior habitat and were described as forest birds in Rodewald (2015).
Forest changes
Changes in the extent of forest and interior forest between 1971 and 1999 differed
dramatically among the BBS routes (Fig. 1). Change in total forest varied
from a decline of 0.4% to a decline of 24.8%. The greatest declines were in eastern
Massachusetts and on Cape Cod. Declines in interior forest followed the same
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geographical pattern as declines in total forest, but were larger, varying from 2.0%
to 69.5% (Fig. 1) The extent of loss would be even higher (up to 80%) if the denominator
in the percentages was the extent of 1971 forest rather than the average
of data from 1971 and 1999. Changes in interior forest were negatively correlated
with changes in human population density (R = -0.41, n = 25, P < 0.05).
Figure 1. Percent change in interior forest and total forest within 400 m of 25 breeding bird
survey routes in Massachusetts between 1971 and 1999.
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Changes in bird abundance
Between 1971 and 1999, population trends of different forest interior bird species
varied widely (Table 1). Among all forest-interior species, the mean change
was +8 (not significantly different from 0 by 1-tailed t-test: t = 0.93, P > 0.05, df =
27), but with wide variation from +100 for the Common Raven and +98 for the Wild
Turkey to -78 for the White-throated Sparrow. As noted below, populations of Corvus
corax (Common Raven), Meleagris gallopavo (Wild Turkey) and Haemorhous
purpureus (Purple Finch) are likely to be changing for reasons unrelated to landuse
change. Excluding these 3 species yielded a mean change index of +3, also not
significantly different from 0 (1-tailed t-test: t = 0.39, P > 0.05, df = 24).
Changes in bird abundance showed no significant relationship to migration status
or nest position. Mean change index was +1 for Neotropical migrants and +19
for short-distance migrants and residents (t = 1.07, P > 0.05, df = 26), (+8 for the
latter group if Common Raven, Wild Turkey, and Purple Finch are excluded). With
these 3 species excluded, the mean change index was -11 for ground nesters and
+11 for above-ground nesters (t = 1.37, P > 0.05, df = 23).
Table 2. Correlations between 2003–2007 bird abundances and extent of interior and total forest within
400 m of the census routes.
Species Interior forest 400 m Forest 400 m
American Redstart 0.701 0.548
Black-and-white Warbler 0.524 0.522
Blackburnian Warbler 0.673 0.427
Black-throated Blue Warbler 0.801 0.632
Black-throated Green Warbler 0.616 0.543
Blue-headed Vireo 0.804 0.735
Common Raven 0.586 0.459
Dark-eyed Junco 0.619 0.553
Hairy Woodpecker 0.677 0.730
Hermit Thrush 0.652 0.552
Least Flycatcher 0.612 0.615
Magnolia Warbler 0.645 0.531
Nashville Warbler 0.682 0.584
Northern Waterthrush 0.566 0.458
Ovenbird 0.793 0.742
Pileated Woodpecker 0.667 0.551
Purple Finch 0.649 0.565
Red-eyed Vireo 0.645 0.573
Rose-breasted Grosbeak 0.753 0.661
Ruby-throated Hummingbird 0.730 0.638
Scarlet Tanager 0.758 0.737
Veery 0.897 0.767
White-throated Sparrow 0.625 0.526
Wild Turkey 0.510 0.492
Winter Wren 0.672 0.518
Wood Thrush 0.661 0.601
Yellow-bellied Sapsucker 0.505 0.433
Yellow-rumped Warbler 0.705 0.658
Average 0.672 0.589
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Figure 2. Association
between changes
in abundance of
28 forest birds along
Massachusetts BBS
routes and estimated
changes from
MassAudubon’s
Bird Atlas project.
Bird Atlas changes
were given qualitatively
in the source
( M a s s A u d u b o n
2016) and were assigned
ranked numerical
values as
follows: -2 = strong decrease, -1 = likely decrease, 0 = stable, 1 = likely increase, 2 = strong
increase. The Spearman’s rho correlation between the two variables is 0.865.
Changes among agonist species were variable. The American Crow (+37) and
Brown-headed Cowbird (+5) increased, whereas the Blue Jay (-21) and Common
Grackle (-16) declined.
Changes in abundance of FIA bird species obtained from BBS data were highly
correlated with changes in block occupancy of the same species estimated from the
bird atlases (Spearman’s rho: Rs = 0.865, P < 0.001; Fig. 2).
Relationship between change in interior forest and change in bird abundance
For 20 of the 28 FIA bird species, a positive correlation coefficient existed
between change in abundance (1971–1999) and change in extent of interior forest
(Table 1). A binomial test (P = 0.018) reveals that it is unlikely to get this many
positive correlation coefficients by chance alone. For individual species, 6 correlation
coefficients were significant, all positive. Among the 15 FIB bird species,
13 had positive correlations, of which 5 were significant. A binomial test indicates
that the probability of having 13 of 15 coefficients positive by chance alone is
0.004. Patterns of change in bird abundance were marginally more strongly correlated
with changes in extent of interior forest than with changes in total forest
area. For the 15 FIB species, 11 showed higher (more positive) correlations with
changes in interior forest than changes in total forest. The likelihood of 11 or
more species showing such a pattern by chance is 0.059 (binomial test). Among
the 28 FIA species, 19 showed greater correlations with change in interior forest
than total forest. The likelihood of this or more extreme outcomes occurring by
chance is 0.044.
Among the 4 agonist species, none showed a population change that was significantly
related to the change in extent of either forest or interior forest (Table 1).
For both forest and interior forest, however, 3 of the 4 correlation coefficients were
negative, hinting at the possibility of population increases as forest extent declined.
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Discussion
Classifying forest-interior species
The suite of forest-interior bird species identified in this study shows considerable
overlap with sets of forest-interior species defined in other studies. For
example, 13 of the 15 FIB species in this study were included in at least 1 of the 3
studies by Askins et al. (1987), Dunford and Freemark (2004) and Freemark and
Collins (1992). All were reported as forest-interior species, except that Picoides
villosus (Hairy Woodpecker) and Vireo olivaceus (Red-eyed Vireo) were treated
as an interior/edge species by Dunford and Freemark (2004) and the Wood Thrush
was treated as an interior/edge species in all 3 studies. Most of the other 13 forest
species that contributed to the FIA group were not classified as interior species
in these other studies. A few that were treated as interior species in some studies
were not included in our FIB list because they are sometimes associated with early
successional habitats (Setophaga magnolia [Magnolia Warbler], Catharus fuscescens
[Veery]) or with forest openings and edge (Setophaga ruticilla [American
Redstart], Parkesia noveboracensis [Northern Waterthrush]) (R.I. Bertin, pers.
observ.; Rodewald 2015). Complete consensus on habitat classifications of species
is unlikely (and some real geographic variation undoubtedly exists), but our classification
seems reasonable in light of published information.
Changes in forest habitat
Several trends have been evident in Massachusetts forests in the past half
century. Forest area has declined markedly in some parts of the state, especially
eastern Massachusetts and Cape Cod, while it has been stable or increasing elsewhere
(MacConnell et al. 1991). The extent of interior forest has declined much
more rapidly than the area of total forest. The 5 most affected routes in this study,
all in eastern Massachusetts, each lost over 38% of interior forest between 1971
and 1999 (the percent loss is actually higher if expressed relative to 1971 rather
than relative to the average of the 2 years). It is important to note that changes
along BBS routes may differ from those in the broader landscape, because BBS
routes typically follow secondary roads. However, 2 lines of evidence suggest that
this bias is not substantial. First, studies comparing vegetation changes along BBS
routes elsewhere in the eastern United States to those in the broader landscape
typically find little difference (Bart et al. 1995, Keller and Scallan 1999). Second,
our results are consistent with those of other studies not tied to BBS routes (de la
Crétaz et al. 2010).
Although not quantified in this study, changes in forest successional status have
also occurred during the past several decades. Many agricultural fields were abandoned
in the late 1800s and early 1900s, replaced by early successional forest in
the next few decades. MacConnell and Niedzwiedz (1974) report that maturity of
central Massachusetts forests increased during the period 1951–1971. Similarly, de
la Crétaz et al. (2010) documented an increase in the area of Massachusetts forests
dominated by large-diameter trees between 1985 and 1998.
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Validity of bird abundance data
Most census data, including BBS data, are subject to various biases, and these
have been discussed extensively elsewhere (Harris and Haskell 2007, O’Connor
et al. 2000, Ralph and Scott 1981, Sauer et al. 1994). Possible issues include
inadequacy of roadside surveys for evaluating forest-interior species, influences
of habitat changes on detectability, interspecific differences in detectability, and
observer biases. Of most concern for this study are biases that would change over
time, either producing spurious changes or obscuring real changes in bird abundance.
Thus, interspecific differences in detectability are unlikely to affect our
major conclusions, which involve temporal trends in individual species. As noted
above, BBS surveys are unlikely to be entirely representative of changes across the
landscape, because census routes lie primarily along secondary roads. These areas
probably have suffered more habitat alteration than more remote areas, but less than
areas along major thoroughfares or in more-developed areas. However, other studies
in the Northeast suggest that changes along BBS routes were generally similar
to those taking place in the broader landscape (Bart et al. 1995, Keller and Scallan
1999). Furthermore, Harris and Haskell (2007) concluded that late successional
birds (e.g., those used in this study) were not subject to the same sorts of negative
biases that affected roadside censuses of early successional species.
Sauer et al. (1994) described a slight positive bias in BBS census data due to
the improving quality of observers. This factor would slightly inflate population
trends reported in this study. However, this bias is only likely to be relevant when
interpreting subtle trends in bird populations (Sauer et al. 1994). Here we focus
on conspicuous trends, and our main interest is whether these shifts are related to
changes in interior forest. Unless observer biases were large and correlated with
habitat changes along routes, which seems improbable, this bias seems unlikely to
have had a major influence on our results.
Bird census data are always subject to uncontrolled variables, such as weather
and differences in observer quality. However, the high correlation between the
BBS abundance data used in this study and independent BBA occupancy data (Rs =
0.865) suggests that BBS data are consistent with broader trends. This correlation
is particularly impressive given the difference in methodology (census blocks vs.
point counts), and the area and time period covered. Thus, we consider the BBS bird
abundance data to be a valid indicator of population changes along bird routes in
Massachusetts that can be extrapolated with caution to broader geographical areas.
Forest fragmentation and bird decline
Our major result is that changes in abundance of forest-interior bird species as a
group are associated with changes in the extent of interior forest but that this effect is
relatively weak, indicating that other factors are likely to have had larger effects on
populations of most species over the period examined. This general association between
bird abundance and extent of interior forest is shown by the fact that changes in
abundance of 13 of the 15 FIB species exhibited positive correlation coefficients with
change in abundance of interior forest, including 5 that were significant.
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The decline of forest bird species in reduced, fragmented, and/or isolated forest
plots has been demonstrated in many parts of eastern North America and is
predicted to continue (Askins et al. 1990, Betts et al. 2007, Brown et al. 2014,
Whitcomb et al. 1981). Reductions are often attributed to increased predation of
eggs and nestlings and/or greater nest parasitism by the Brown-headed Cowbird
(Brittingham and Temple 1983, Chalfoun et al. 2002). Nest parasitism seems unlikely
to be an important factor in the trends noted on Massachusetts BBS routes
because Brown-headed Cowbird populations were not correlated with the extent
of either forest or interior forest (Table 1). Kluza et al. (2000) similarly concluded
that cowbirds were unlikely to have caused declines of forest understory bird populations
in residential areas elsewhere in Massachusetts. Avian nest predators also
seem unlikely to be associated with bird population declines as population trends
in 2 of the 3 predator species were negative despite declines in forest extent. Mammalian
predators, including both native species and cats, could have been important
as populations of these species typically increase where intact forest is fragmented
by residential development (Phillips et al. 2005, Wilcove 1985)..
While the effects of breeding habitat on populations of forest interior birds in
Massachusetts were relatively minor during the study period, this does not mean
that efforts to maintain and restore interior forest habitat are unimportant. Gradual
declines and fragmentation in forest habitat lead to substantial changes over long
time periods, with detrimental effects for forest-interior birds. This relationship is
well illustrated by the changes in forest-interior bird species along the East Dennis
route, which experienced an interior-forest decline exceeding 69% during the study
period. Several forest bird species disappeared completely, including the American
Redstart, Catharus guttatus (Hermit Thrush), Veery, and Wood Thrush, and others
declined by more than 60%, including the Mniotilta varia (Black-and-white Warbler),
Seiurus aurocapilla (Ovenbird), Purple Finch, and Red-eyed Vireo. These
changes echo those reported during a 37-year period from a forested area altered
by residential development in Virginia (Aldrich and Coffin 1980). Similar changes
would be likely along other Massachusetts routes if they suffered comparable
losses of interior forest.
A second reason that our results do not justify a relaxation of conservation efforts
on the breeding grounds is that nesting success is undoubtedly influenced
by habitat quality. Several studies have suggested that pairing success and nest
productivity decline near forest edges or in fragmented forest (Donovan et al.
1995, Driscoll and Donovan 2004, Villard et al. 1993). Thus, even if the extent of
interior forest were unrelated to the numbers of individuals of forest-interior bird
species detected on the breeding grounds, the numbers of fledged offspring likely
would be lower in areas with less interior forest, reducing their value as forestbird
breeding habitat.
The primary driver of the decrease in interior forest over the period examined
has been conversion of forest for residential use and associated infrastructure
(MacConnell et al. 1991). Detrimental effects of the spread of housing into previously
forested areas have been predicted or demonstrated in several studies (Brown
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et al. 2014, Friesen et al. 1995, Kluza et al. 2000, Phillips et al. 2005, Suarez-Rubio
and Lookingbill 2016, Wood et al. 2014), and are usually attributed to increases in
nest predation and nest parasitism. Human population size is a reasonable proxy for
the residential footprint, and the association between extent of interior forest and
human population density is shown by the significant negative correlation between
these 2 variables. The areas of the state most affected in the interval under study
were in eastern Massachusetts, including Cape Cod and nearby islands. Between
1970 and 2000, human population increased by 57% in these areas, compared to
18% in central Massachusetts and just 8% in the western counties (NEBR 2016).
Because eastern Massachusetts is already heavily developed, subsequent population
increases will likely push westward, leading to declines of interior forests and
the species they support in central and perhaps western forests in the decades ahead.
Initial evidence of this population trend is shown by the 6.3% human population
increase between 2000 and 2010 in central Massachusetts, greater than the 3.7% increase
in eastern counties (NEBR 2016). Patterns of development often exacerbate
forest fragmentation. The common zoning practice of establishing moderately large
(0.5–3 acre; 0.25–1.5 ha) minimum lot sizes in semi-rural areas impacts more total
land and more interior forest than clustered development of the same number of
housing units (Steel 1999). Thoughtful zoning standards and land-use management
practices could help retain forest interior birds in the Massachusetts landscape.
Other factors affecting bird populations
As noted above, factors other than the decline in interior forest were responsible
for most bird population changes between 1971 and 1999 (Table 1). Wide variation
in population trends was documented among the 15 FIB species, and 5 species
showed substantial population increases (change index of +25 or higher) between
1971 and 1999, despite the decline of interior forest along every BBS route during
this period. Such interspecific differences in population trends are common (Germaine
and Vessey 1997, Mancke and Gavin 2000, Sauer et al. 2014). Evidence of
the importance of factors other than habitat decline comes from studies showing
substantial changes in forest bird populations in areas of relatively unfragmented
forest, perhaps related to successional changes, food availability, or habitat changes
on migration routes or wintering grounds (Ambuel and Temple 1982, Blodgett et
al. 2009, Holmes and Sherry 2001). Numerous authors have suggested that Neotropical
migrants are at particular risk for population declines (Sauer et al. 1996,
Whitcomb et al. 1981). However, among Neotropical migrants in our more stringently
defined FIB group, we see both apparent increases (+56 for Vireo solitarius
[Blue-headed Vireo], +26 for Setophaga virens [Black-throated Green Warbler])
and declines (-42 for Wood Thrush, -22 for Piranga olivacea [Scarlet Tanager]).
Notably, those species showing the greatest declines, including the 2 species mentioned
above, do not have changes that are significantly correlated with changes in
interior forest. This observation suggests that the population declines are regionwide
phenomena unrelated to the local extent of interior-forest breeding habitat.
Supporting this interpretation are the findings of Sauer and Droge (1992) that the
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Blue-headed Vireo and Magnolia Warbler showed significant population increases
for all BBS census routes between 1966 and 1988. In their work, the Wood Thrush
showed a significant decrease over this period, whereas the Scarlet Tanager showed
a significant increase for 1966–1988 but a significant decline for 1978–1988. The
lack of consistent declines among Neotropical migrants in our study parallels the
findings of Blodgett et al. (2009) at a site in western Massachusetts, an analysis of
11 BBS census routes in coastal Maine and New Hampshire by Witham and Hunter
(1992), and a general analysis of BBS results by Sauer et al. (1996). They are also
consistent with data collected during migratory periods at a bird-banding station in
coastal Massachusetts, which found greater declines among migratory species wintering
in the southeastern US than among those wintering in the Tropics (Hagan et
al. 1992). However, other studies have shown greater declines among Neotropical
migrants (Ambuel and Temple 1982, Hall 1984, Lynch and Whigham 1984, Robbins
et al. 1989, Terborgh 1992, Whitcomb et al. 1981). Resolution of this issue is
likely to require detailed understanding of the factors acting on particular species.
Among trends in individual species in our study, several have been noted in
other studies and, in some cases, have evident explanations. Wild Turkeys showed
one of the largest increases (+98; Table 1). This species was extirpated from Massachusetts
during the 1800s, followed by successful reintroduction in the 1970s and
subsequent population increases (EEA 2016). Hence, the large negative correlation
(r = ‑0.61) between Wild Turkey abundance and extent of interior forest very likely
results largely from the chance juxtaposition of 2 events: reintroduction and forest
decline. Common Ravens showed the largest percent increase of any species
in our data. These birds were extirpated from the state in the 1800s, and the first
Massachusetts Breeding Bird Atlas reported no confirmed breeding records during
the 1974–1979 census (MassAudubon 2016). Ravens subsequently spread into the
state from the north and west and are now confirmed breeders in all but the southeastern
portion of the state. Declines in the Purple Finch may be linked to spread of
Haemorhous mexicanus (House Finch), introduced from the western United States
(Wootton 1987).
Changes in the forest successional stage are likely to have influenced populations
of several species. Conspicuous increases in Hylatomus pileatus (Pileated
Woodpecker) numbers were seen in our study and along BBS routes more generally,
as well as in BBA occupancy data (Table 3). Changes along Massachusetts
routes are not due to increasing forest area because our data show that forest extent
along BBS routes decreased during this period. Instead, this increase is likely
to reflect increasing maturity of Massachusetts forests (de la Crétaz et al. 2010),
with larger trunk diameters important for both the food supply and appropriate
nesting sites for this species (Lemaître and Villard 2005, Savignac et al. 2000).
Increasing tree sizes may also explain increases in Sphyrapicus varius (Yellowbellied
Sapsucker) in Massachusetts and elsewhere (Table 3), by providing
more “suitable trees for feeding, drumming and nesting” (Blodgett et al. 2009).
Increasing forest maturity was suggested to underlie increases in populations
of the Black-throated Green Warbler and Setophaga coronata (Yellow-rumped
Warbler) in New Hampshire and western Massachusetts (Blodgett et al. 2009,
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Holmes and Sherry 2001) and may underlie similar trends in Massachusetts. The
conspicuous increase in Archilochus colubris (Ruby-throated Hummingbird)
along Massachusetts routes was mirrored in BBS and BBA data (Table 3). This
change was attributed by MassAudubon (2016) to the “proliferation of natural
beaver clearings and cultivated suburban gardens,” to which we might add the
popularity of hummingbird feeders.
Changes in vegetation diversity or composition, including successional changes,
could have influenced some bird species (Buchanan et al. 2016, Holmes and Sherry
Table 3. Population trends from 4 studies. Bird atlas results are from 2 projects in Massachusetts, one
in 1974–1979 and the other in 2007–2011 (MassAudubon 2016). BBS refers to Breeding Bird Survey
results as summarized in MassAudubon (2016). Blodgett et al. (2009) monitored an intact forest area
in western Massachusetts; numbers reported are slopes of population regression lines, with asterisks
denoting those significantly different from zero.
Common name Bird atlas BBS Blodgett et al. This study
Forest Interior B Species
Black-and-white Warbler Likely decrease Likely decrease -0.04 -18
Blackburnian Warbler Likely increase 0 0.02 -13
Black-throated Blue Warbler Likely increase 0 0.02 -6
Black-throated Green Warbler Likely increase 0 0.06* +28
Blue-headed Vireo Likely increase Likely increase -0.02 +56
Dark-eyed Junco 0 0 -0.16* -36
Hairy Woodpecker Likely increase 0 0.00 +3
Hermit Thrush Likely increase 0 -0.04* +9
Ovenbird Likely increase 0 -0.01 +5
Pileated Woodpecker Strong increase Strong increase - +45
Red-eyed Vireo 0 0 0.02* -17
Scarlet Tanager Stable Likely Decrease -0.02 -22
Winter Wren Strong increase 0 -0.02 +41
Wood Thrush Likely decrease Likely decrease 0.01 -42
Yellow-rumped Warbler Likely increase 0 0.01 +59
Other Forest Species
American Redstart Likely increase 0 -0.03* +7
Common Raven Strong increase 0 - +100
Least Flycatcher 0 Likely Decrease -0.08* -55
Magnolia Warbler Likely increase 0 0.13* +33
Nashville Warbler Strong decline Strong decline - -32
Northern Waterthrush Strong increase 0 -0.05* +15
Purple Finch Strong decline Likely decrease - -50
Rose-breasted Grosbeak Likely increase Likely decrease 0.00 -10
Ruby-throated Hummingbird Strong increase Likely increase - +86
Veery Likely increase 0 0.00 -2
White-throated Sparrow Strong decline Strong decline - -78
Wild Turkey Strong increase 0 - +98
Yellow-bellied Sapsucker Strong increase Likely increase 0.04* +15
Agonist Species
American Crow Stable Likely Increase 0.04 +33
Blue Jay Stable Likely Decrease -0.02 -21
Brown-headed Cowbird Likely increase 0 0.00 +5
Common Grackle Stable Likely Decrease - -17
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2001). Zonotrichia albicollis (White-throated Sparrow) numbers declined more
than those of any other species (-78), a pattern mirrored in coastal Maine, New
Hampshire, and Massachusetts (Hagan et al. 1992, Witham and Hunter 1992) and
in BBA data. MassAudubon (2016) concluded that habitat loss was “mostly to
blame”. More specifically, the maturing of forests in areas that were previously in
thicket and shrubland may be a factor, since the breeding habitat for this species
includes forests “with numerous openings”, “second growth”, and “brushy field
edges and overgrown pastures” (Rodewald 2015).
Other likely influences on abundance of forest birds include habitat changes on
migration routes or wintering grounds (Dugger et al. 2004, Keller and Yahner 2006,
Norris et al. 2004, Rappole and McDonald 1994, Sauer et al. 1996, Sillett et al.
2000, Taylor and Stutchbury 2016). Such effects would not be surprising given the
extensive destruction of Neotropical forests in some areas and the dramatic effects
on local populations of Neotropical migrants (Askins et al. 1992, Bradshaw et al.
2009, Rappole and Morton 1985).
Conclusion
The extent of forest and especially interior forest declined along BBS routes in
Massachusetts between 1971 and 1999. Populations trends in forest interior birds
varied widely among species during these decades. Overall, bird population trends
were significantly associated with forest trends, but the effect was weak. Other
factors, including species recovery from reduced populations, changes in forest
maturity, and events during migration and on wintering grounds, likely had much
greater effects during this period.
Acknowledgments
We thank Richard Lent for help in extracting data files from the Breeding Bird Survey
and the George I. Alden Trust for providing summer support to T.J. Gardner through an
Excellence in Career Related Undergraduate Education Award. Thanks to Peter Paton and
2 anonymous reviewers for their helpful comments on earlier versions of the manuscript.
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