Results of 20 years of experimental forest management on breeding birds in Ozark forests of Missouri, USA

Forest Ecology and Management 310 (2013) 747–760

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Results of 20 years of experimental forest management on breeding birds in Ozark forests of Missouri, USA Dana L. Morris a,b,⇑, Paul. A. Porneluzi b, Janet Haslerig c, Richard L. Clawson d, John Faaborg a a

Division of Biological Sciences, 105 Tucker Hall, University of Missouri, Columbia, MO 65211, USA Division of Science, Mathematics, and Computer Science, 411 Central Methodist Square, Central Methodist University, Fayette, MO 65248, USA c Missouri Department of Conservation, Resource Science, Missouri Department of Conservation, 2901 W. Truman Blvd., Jefferson City, MO 65109, USA d Missouri Department of Conservation, Resource Science, 3500 East Gans Road, Columbia, MO 65201, USA b

a r t i c l e

i n f o

Article history: Received 6 June 2013 Received in revised form 11 September 2013 Accepted 11 September 2013 Available online 7 October 2013 Keywords: Avian diversity Bird density Nest survival Ozark forests Even-aged management Uneven-aged management

a b s t r a c t Understanding the relationship between forest management and bird populations requires understanding the effects of silvicultural practices on avian demography at large spatio-temporal scales. The Missouri Ozark Forest Ecosystem Project (MOFEP) is a long-term, large-scale manipulative experiment testing the effects of even-aged (3–13 ha cuts over 10–15% of the site; n = 3), uneven-aged (0.03– 3.14 ha cuts over 57% of the site; n = 3), and no harvest forest management on ecosystem level responses. We report on the effects of these management systems on the density and reproductive success of 11 songbird species from 5 years of pre-harvest (1991–1995) to 14 years of post-harvest (1997–2010). Density of four of the five mature forest species were lower after harvest in all management types and did not return to pre-harvest density, even in no harvest sites. Ovenbird (Seiurus aurocapilla) and Wood Thrush (Hylocichla mustelina) responded most negatively to even-aged management in the early post-harvest period (1997–2003) where density was significantly lower than in no harvest sites. Kentucky Warbler (Geothlypis formosa) density increased on uneven-aged and even-aged sites during early post-harvest, but returned to pre-harvest density on both management types by 14 years post-harvest. Acadian Flycatcher (Empidonax virescens) and Worm-eating Warbler (Helmitheros vermivorus) had lower density in all treatments post-harvest. Among the six early-successional species, density of Blue-winged Warbler (Vermivora cyanoptera), Hooded Warbler (Setophaga citrina), Indigo Bunting (Passerina cyanea) White-eyed Vireo (Vireo griseus) and Yellow-breasted Chat (Icteria virens) were significantly higher in even-aged and uneven-aged sites than in no harvest sites after harvest. Density of Prairie Warbler (Setophaga discolor) were significantly higher only in even-aged management after harvest. Prairie Warbler, Hooded Warbler, and White-eyed vireo appeared on the study sites following harvest while Indigo Bunting, Prairie Warbler, and Yellow-breasted Chat were absent from the study area 14 years post-harvest. Nest survival of mature forest and early-successional species did not change significantly from the pre- to the late post-harvest period or with forest management. Brood parasitism rates remained low from pre-harvest to late post-harvest, but parasitism rates were higher for early-successional species (4%) than mature forest species (1%). Although forest management had variable effects on species, we suggest a modified version of even-aged management could maximize benefits to early-successional species while minimizing decreases in mature forest bird species in central hardwood forests. Rather than the current prescription to harvest 10–15% of the mature stands every 15 years, we recommend harvesting approximately half the number of mature stands with a shorter re-entry period of 8–10 years. Ó 2013 Elsevier B.V. All rights reserved.

Abbreviations: MOFEP, Missouri Ozark Forest Ecosystem Project; EAM, evenaged Management; UAM, uneven-aged management; NH, no harvest; BBS, Breeding Bird Survey; WNV, West Nile Virus. ⇑ Corresponding author. Address: Division of Science, Mathematics, and Computer Science, 411 Central Methodist Square, Central Methodist University, Fayette, MO 65248, USA. Tel.: +1 660-248-6678; fax: +1 660-248-6398. E-mail addresses: [email protected] (D.L. Morris), ppornelu@ centralmethodist.edu (Paul. A. Porneluzi), [email protected] (J. Haslerig), [email protected] (R.L. Clawson), [email protected] (J. Faaborg). 0378-1127/$ - see front matter Ó 2013 Elsevier B.V. All rights reserved. http://dx.doi.org/10.1016/j.foreco.2013.09.020

1. Introduction Eastern deciduous forests support 34 obligate breeding bird species (NABCI, 2011) and populations of some of these species have declined by as much as 25% (4.9–42.9% decline; n = 25) in recent decades (Sauer and Link, 2011). Although eastern forests have recovered from vast historic deforestation, more recent fire

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suppression and increased herbivory have resulted in loss of understory and diverse structural components (Abrams, 2003; McShea et al., 2007). Increased fragmentation has occurred in recent years largely due to urbanization (Riitters et al., 2002; Riitters and Coulston, 2005). During this period, populations of shrubland bird species have also declined (Askins, 1993). These declines have been linked to the loss of early-successional forests due to forest maturation and parcelization of forest tracts that impedes comprehensive forest management (Askins, 2001; Brooks, 2003). Sustainable forest management seeks to promote a balance between conserving biological diversity, soil, and water resources while producing harvestable timber (Flader, 2004). Understanding the relationship between forest management and bird responses is key to securing diverse and resilient bird populations. Past studies have generally measured relative abundance or nest success over short periods or at the patch or edge scale which can give inconsistent results regarding fragmentation effects (Stephens et al., 2003; Lampila et al., 2005; Smith et al., 2011). Research on the effects of forest management needs to be long-term, large-scale and include well-replicated pre-harvest and post-harvest data to tease apart effects of environmental stochasticity from forest harvest (Stephens et al., 2003; Faaborg et al., 2010). To better facilitate conservation planning, large-scale studies should include the spatial extent of habitat influences on fitness to distinguish between site-level and population-level responses to forestry (Faaborg et al., 2010). Therefore, rigorous experimental research on abundance and reproductive success is needed to guide the level of timber harvest and forest management that balances wildlife habitat needs and economic benefits of timber harvesting. Additionally, although publicly-owned eastern forests support only 15% of the distribution of the 34 obligate breeding bird species (NABCI, 2011), these parcels of land tend to occur in larger blocks than privately-owned forests that could enable coordination and management at a larger scale. Publicly-owned forests are reaching harvestable age and sustainable forest management is needed to provide critical habitat and refugia for obligate forest breeding birds and to mitigate losses to urban development and degradation on private lands (NABCI, 2011). Bird community composition is influenced by processes occurring at multiple spatial scales (Faaborg et al., 1995; Boulinier et al., 2001; Guenette and Villard, 2005; McDermott and Bohall Wood, 2010). Forest structure can influence the occurrence of early-successional and mature forest species at the stand- or microhabitat scale (McDermott and Bohall Wood, 2010). At both the local and landscape scales, breeding birds exhibit demographic responses to forest fragmentation that can ultimately lead to population declines via lower nesting success (Robinson et al., 1995; Porneluzi and Faaborg, 1999), higher brood parasitism by the Brown-headed Cowbird (Molothrus ater, reviewed by Chace et al., 2005), and decreased pairing success (Gibbs and Faaborg, 1990; Bayne and Hobson, 2001; Lee et al., 2002). However, responses to landscape disturbances are often mediated by regional landscape characteristics and forest type (Stephens et al., 2003; Lloyd et al., 2005; Smith et al., 2011). Additionally, increased edge habitat created by fragmentation can attract nest predators (Dijak and Thompson, 2000; Chalfoun et al., 2002a), but predator responses to edge can be species-specific and scale-dependent (Cox et al., 2012a). Managers are often tasked to elaborate management plans for diverse forests that support a high diversity of bird species and to maximize biodiversity via rotational harvests (Raeker et al., 2010). Harvesting typically creates suitable habitat for earlysuccessional bird species (Annand and Thompson, 1997; King

et al., 2001; Thompson and DeGraaf, 2001; Gram et al., 2003) and provides post-breeding habitat for mature forest and earlysuccessional species (Anders et al., 1998; Pagen et al., 2000; Vitz and Rodewald, 2006; McDermott and Bohall Wood, 2010). However, intensive management can have negative effects on mature forest species that avoid areas of high edge density (King and DeGraaf, 2000; Villard et al., 2007; Wallendorf et al., 2007). Forest management often requires a combination of regenerative cuts and stand improvement or thinning. Some mature forest species respond favorably to timber stand improvement (Goodale et al., 2009), while the edge-sensitive Ovenbird (Seiurus aurocapilla) avoids areas treated with timber stand improvement (Wallendorf et al., 2007). Moreover, habitat alteration can have important effects on mature forest species regardless of distance to edge (Guenette and Villard, 2005; Becker et al., 2011). Yet, many studies suggest that silvicultural methods that achieve regeneration without excessive thinning and minimizing disturbance due to road maintenance and reentry have the potential to maximize biodiversity of mature forest and early-successional bird species (Keller et al., 2003; Guenette and Villard, 2005; Twedt and Somershoe, 2009; McDermott and Bohall Wood, 2010). In this study we evaluated the long-term consequences of three forest management alternatives to breeding birds in the Missouri Ozark Forest Ecosystem Project, a forest management experiment on public lands (Brookshire and Shifley, 1997). A no harvest treatment allowed tracking of birds in response to forest age and other natural processes in the absence of harvest. Even-aged management harvested 10–15% of a site where mature stands were harvested from discrete stands 3–25 ha in area using a 15year re-entry period. Within EAM stands, residual trees were also removed to promote regeneration of shade intolerant species. zUneven-aged management harvested a similar volume of timber but the cuts were spread over a greater area (41–69% of a site), also over a 15-year re-entry period. Gram et al. (2003) evaluated the impacts on breeding birds during the first three years of post-treatment data. Here, we investigated one complete 15-year entry cycle and assessed impacts during early post-harvest and late post-harvest to compare the effects of these three management alternatives. Based on results from studies testing effects of UAM on Ovenbirds (Vanderwel et al., 2007; Haché and Villard, 2010), we predicted density of mature forest species would decline in proportion to the percent of basal area removed during harvest with treatment effects lasting until mature forest regenerates to pre-harvest levels. We predicted declines would be greater in UAM because the effects of perforation, gaps created by groupand single-tree selection cuts, were distributed across a greater area than with EAM. For early-successional species, we predicted a larger increase in species’ density with EAM than UAM because post-harvest ground cover increased at a higher rate in clearcuts than in group-selection cuts (Zenner et al., 2006), thus providing more suitable breeding habitat. We expected the low intensity of UAM and EAM harvests (both removed 16% of basal area) and extensive forest cover in the region (84%) would buffer any severe increases in nest predation rates post-harvest (Thompson et al., 2002; Gram et al., 2003). However we still predicted a moderate decline in nest survival in harvested sites compared to NH sites because of the potential for regenerating forest to benefit some species of nest predators (Chalfoun et al., 2002b). Similarly, because of the vast forest cover (84%) and lack of large agricultural openings in this study, we predicted no significant increases in brood parasitism rates post-harvest (Thompson et al., 2002; Chace et al., 2005). However, given that Brown-headed Cowbirds would likely be more abundant in harvested than non-harvested sites (Annand

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and Thompson, 1997), we predicted parasitism rates would be higher for early-successional species than mature forest species. 2. Methods 2.1. Study area The MOFEP study area consists of 9 sites that average 400 ha in size (Brookshire and Shifley, 1997). The study sites were located in Carter, Reynolds, and Shannon counties in the forested Ozark hills of south-central Missouri (91°010 –91°130 W and 37°000 –37°120 N, Fig. 1). White oak (Q. alba), Black Oak (Q. velutina) and Scarlet Oak (Q. coccines) dominated the MOFEP sites, followed by Shortleaf Pine (P. echinata), Post Oak (Q. stellata), Mockernut Hickory (C. tomentosa), Black Hickory (C. texana), and Pignut Hickory (C. glabra) (Shifley and Kabrick, 2000). At the start of this project in 1990, most overstory trees were 50–70 years old (Shifley and Kabrick, 2000; see Shifley and Brookshire, 2000 for more details). Natural and anthropogenic disturbances such as fire, logging, and agriculture have affected the MOFEP sites over the past 1000 years (Guyette and Larsen, 2000). Prior to the purchase of the MOFEP study sites by the Missouri Department of Conservation (site 6, 1925; sites 3–5, 1938; sites 1–2, 1944; sites 7–9, 1952), the land was managed for timber. As a criterion for inclusion in this study, sites had to be largely free from manipulation for at least 40 years before the start of MOFEP in 1990 (Brookshire and Shifley, 1997). These 9 sites were grouped into 3 blocks based on spatial proximity. Each block contained 3 sites randomly assigned to EAM, UAM or no harvest (NH) treatments, resulting in a randomized complete block design. Each site was a collection of approximately 70 smaller stands, averaging 5 ha in size (Brookshire and Shifley, 1997). Timber harvests and thinning, to increase growing space

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for residual trees, were applied to stands within sites. The NH sites were not harvested and wildfires were suppressed. The NH sites serve as indicators of successional processes as the forest matures over the life-span of this 100-year study. The Missouri Department of Conservation supervised harvest of the sites from May 1996 to May 1997. In the EAM treatments, approximately 10–15% of the total forest area was clearcut. Clearcuts were 3–13 ha in size, resulting in seven to nine clearcut stands per EAM site (Fig. 1; Brookshire and Shifley, 1997). In addition, foresters thinned 5–24% of each site to promote tree growth of selected sizes. The goal of EAM and thinning was to create a specific distribution of tree size classes in study sites: 10% in regeneration, 20% in small trees (trees 6– 14 cm diameter at breast height [dbh]), 30% in poletimber (14– 29 cm dbh, and 40% in sawtimber (>29 cm dbh; Kabrick et al., 2002). In UAM treatments, foresters harvested trees from a combination of small-group and single-tree selection cuts across 41–69% of each site. Small-group cuts ranged from 21 to 43 m in diameter, depending on aspect. Five percent of the harvested area per UAM site was treated with small-group cuts (153–267 small-group cuts per site). Foresters used single-tree selection cuts to obtain a balance of size classes with 1.5 times more small trees than large trees in the next size class. Even-aged and uneven-aged management treatments removed similar amounts of biomass (3.2 m2/ha, or 16% of basal area), but resulted in different spatial configurations. Thus, in EAM sites, foresters harvested from about 15% of the forest area, and large blocks of cut or thinned forest were interspersed in a matrix of uncut forest. In UAM sites, foresters harvested from about 41–69% of each site, where small areas of cut or thinned forest were scattered throughout uncut forest (Fig. 1; Kabrick et al., 2002). In both

Fig. 1. Map of the study sites in the Missouri Ozark Forest Ecosystem Project. Nine sites, grouped into 3 blocks were randomly assigned to no-harvest, even-aged management or uneven-aged management treatments.

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EAM and UAM sites, a patch of approximately 10% of each site was designated as a permanent reserve and left uncut. Rotation lengths for EAM and UAM sites are approximately 100 years with a 15-year re-entry where different mature stands are harvested with each reentry. 2.2. Data collection We mapped breeding bird territories from mid-May through June in all sites prior to harvest (1991–1995) and after harvest (1997–2003; 2008–2010). For each site, we sampled four subplots (45 ha each) that were visited 10 times per breeding season. On each visit, trained field assistants spent 3–4 h, beginning at dawn, spot-mapping (Anonymous, 1970) one entire subplot. Individual detections of singing males were recorded on enlarged topographic maps of the subplot (map scale 1:3330 m). Field assistants surveyed subplots in rotation to minimize observer bias. To identify territories, we created yearly composite maps for each species per site (all plots combined), with all detections color-coded by census date. We defined territories based on a cluster of at least three detections from three different survey dates, counter-singing, and presence of nests. We counted partial territories as fractions when a site boundary intersected a territory, but this rarely occurred because most sites were bounded by a road or drainage. Mean species density per site was expressed as the mean number of territories per 100 ha. We located and monitored nests while spot-mapping and through deliberate searching from mid-May through July every year from 1991 to 2010, except 1996 when harvesting occurred. We marked nest locations with flagging tape placed 5–10 m from the nest with orientation and identifying information (Ralph et al., 1993). We checked nests every 3–5 days to determine their fate and noted predation and parasitism events (Martin and Geupel, 1993). The majority of nests were on the ground, in shrubs, or in the subcanopy given that canopy nests were too difficult to find and monitor. 2.3. Statistical analyses We limited our analyses to selected focal species based on reliability of detection with spot-mapping and nest height. We selected five mature forest species [Acadian Flycatcher (Empidonax virescens) Kentucky Warbler (Geothlypis formosa), Ovenbird, Worm-eating Warbler (Helmitheros vermivorus) and Wood Thrush (Hylocichla mustelina)] and six early-successional species [Bluewinged Warbler (Vermivora cyanoptera), Hooded Warbler (Setophaga citrina), Indigo Bunting (Passerina cyanea), Prairie Warbler (Setophaga discolor), White-eyed Vireo (Vireo griseus) and Yellowbreasted Chat (Icteria virens)]. To test for treatment, period, and their interaction effects on density of each species we used mixed-model ANOVA with block as a random effect. We grouped the years of the study into categories representing pre-harvest (1991–1995), early post-harvest (1997–2003), and late post-harvest (2008–2010). We used a normal distribution for Ovenbird, Acadian Flycatcher, and Wormeating Warbler and a Poisson distribution for Kentucky Warbler, Wood Thrush, Blue-winged Warbler, White-eyed Vireo and Yellow-breasted Chat which showed non-normal distributions. We specified a negative binomial distribution for Hooded Warbler, Indigo Bunting and Prairie Warbler which showed overdispersed distributions. We used planned contrasts to determine if density differed between periods and treatments within periods. We did not perform Bonferoni corrections of alpha because sample size was small in this experiment (n = 9 sites) and we were interested in detecting an effect if one existed (i.e., minimizing Type II error). Hooded Warbler, Prairie Warbler, and White-eyed Vireo were not

detected on the study area during the pre-harvest period and, therefore, we tested only for differences between early and late post-harvest. For Prairie Warbler, we only tested for a treatment effect because there were no detections during pre-harvest or late post-harvest periods. Using the same analytical approach and predictors, we tested for effects on the relative abundance of Blue Jays (Cyanocitta cristata) and the brood parasitic Brown-headed Cowbird. Because neither of these species defends territories during the breeding season, we did not delineate territory boundaries. Instead, for each species we summarized mean detections per site from spot-map data. We used the logistic exposure method (Shaffer, 2004) to model nest survival as a function of a priori hypotheses concerning harvest treatment and time period. For these analyses, time period was categorized as pre-harvest (1991–1995), early post-harvest (1997–2003) and late post-harvest (2004–2010). The sampling unit with this method is the interval between nest checks. The effective sample size for each species was derived from the model likelihood and consisted of the sum of the number of days that all nests were known to have survived and the number of intervals that ended in failure (Rotella et al., 2004) and is presented for each species. Again, we were interested in the effects of timber harvest and time period on nest survival, but temporal effects such as nest stage and Julian date can strongly influence nest survival (Cox et al., 2012a). To retain parsimony in the final models, we first ran a set of 5 models including single variables nest stage (incubation or nestling) and ordinal date (with and without a quadratic term) and an additive model with both variables to see which best fit the data for each species (Tables A1 and A2). We then used the variable(s) from the top-ranked model as a null temporal model and included it as a covariate in all survival models to avoid confounding effects of nest stage or date. For each species we developed 3 candidate models that included a null temporal model, time period alone, and a global model with treatment, period and their interaction. We included treatment in models only when combined in an interaction with period because inferences based on treatment alone would not be informative. In addition to nest stage or date, we included block as a fixed effect in each model to account for our study design. Block was not included as a random effect because model convergence was problematic. Due to small sample sizes, nest observations for Prairie Warbler were not included and White-eyed Vireo and Yellow-breasted Chat were combined for nest survival analysis and species was included in the model. Both species nest in shrubs and use similar microhabitat so combining them was biologically sound and allowed us to use the samples. We rightcensored (removed final observation) nests with unknown fate or nests that failed due to disturbance by researchers to reduce bias in survival estimates (Manolis et al., 2000). We excluded nests that were found but never observed with eggs or young. We used generalized linear models with binomial distributions and logit link functions in GENMOD (SAS Institute, 2008) to fit the logistic-exposure models (Shaffer, 2004). We used the second-order Akaike information criterion (AICc; Burnham and Anderson, 2002) for model selection and calculated differences in AICc values and Akaike weights to evaluate relative support for candidate models. Models with low AICc scores, DAIC < 4, and high Akaike weights represent best-fit models. To account for model selection uncertainty (Burnham and Anderson, 2002), we reported 95% confidence intervals (CI) for parameter estimates based on model-averaging over the entire candidate model set. We calculated species-specific nest survival probabilities over the entire nest period (period survival) in terms of the effects, by raising daily survival to a power equal to the average length of the nest cycle for each species. Because we were

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interested in the treatment by period interaction, we used modelbased methods (Shaffer and Thompson, 2007) to calculate period survival estimates by treatment and time period for each species while holding the remaining covariates at their mean or for categorical variables at levels representing the proportions of observations at each category level (Shaffer and Thompson, 2007). To examine the probability of parasitism between treatment, period, their interaction, and species group (mature forest vs. early-successional) we used generalized linear models specified in GLIMMIX (SAS Institute, 2008) with a negative binomial distribution. Block was included as a nuisance variable to reduce statistical confounding. 3. Results 3.1. Density Density of mature forest species (except Kentucky Warbler) was lower in all treatments following experimental cutting in 1996 (Fig. 2). Declines from preharvest to early post-harvest (Fig. 2) in NH, EAM, and UAM were 18.8%, 26.0%, 18.5% for Acadian

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Flycatcher; 35.7%, 67.7%, 52.5% for Ovenbird; 24.3%, 34.1%, 17.9% for Worm-eating Warbler; and 69.2%, 34.8%, 1.3% for Wood Thrush, respectively. Planned contrasts show late post-harvest density of Ovenbird, Wood Thrush and Worm-eating Warbler have not returned to pre-treatment density (Table 1), whereas Acadian Flycatcher and Wood Thrush have returned to pre-treatment density in UAM only. The contrast between pre-harvest and late post-harvest density for Acadian Flycatcher was non-significant but it showed the same trend. Ovenbirds had significantly lower density in EAM compared to NH sites during the early post-harvest period. Wood Thrush showed a significant treatment by period interaction where density declined in NH and EAM sites following harvest but increased somewhat in NH and UAM sites in later years (Table 1). Kentucky Warbler density remained low in NH sites but increased significantly on EAM and UAM sites during the early post-harvest period (Table 1); during the late post-harvest period, density declined but were higher in UAM sites than in EAM or NH sites (Fig. 2). Density of Acadian Flycatcher and Worm-eating Warbler did not differ by treatment. Hooded Warbler, Prairie Warbler, and White-eyed Vireo were not present on the study sites or were detected at very low density

Fig. 2. Mean density (±SE) of mature forest species in relation to experimental timber harvesting on the Missouri Ozark Forest Ecosystem Project. Shown are five years of pretreatment data (1991–1995), seven years of early post-treatment data (1997–2003) and three years of late post-treatment data (2008–2010).

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Table 1 Mixed model analysis of variance results of even-aged and uneven-aged forest management (treatment) on density of mature forest birds. Significant effects are in bold. Source

Ndf,Ddfa

Ndf,Ddf

F

Treatmentb Periodc Treatment * period

Acadian Flycatcher 2,3 1.01 0.462 2,3 13.10 0.033 4,9 0.58 0.685

Kentucky 2,3 2,3 4,9

Warbler 22.75 0.015 3.26 0.176 3.40 0.058

Contrastsd PRE vs. EP PRE vs. LP EP vs. LP NH EP vs. EA EP NH LP vs. EA LP NH EP vs. UA EP NH LP vs. UA LP EA EP vs. UA EP EA LP vs. UA LP

1,3 1,3 1,3 1,9 1,9 1,9 1,9 1,9 1,9

0.014 0.057 0.347 0.922 0.953 0.626 0.222 0.559 0.242

1,3 1,3 1,3 1,9 1,9 1,9 1,9 1,9 1,9

1.74 2.38 6.22 23.03 0.15 35.29 6.80 4.91 5.79

Treatment Period Treatment * period

Ovenbird 2,3 2,3 4,9

4.78 40.14 1.29

0.116 0.007 0.344

Worm-eating Warbler 2,3 1.34 0.398 2,3 18.32 0.021 4,9 1.25 0.357

Contrasts PRE vs. EP PRE vs. LP EP vs. LP NH EP vs. EA EP NH LP vs. EA LP NH EP vs. UA EP NH LP vs. UA LP EA EP vs. UA EP EA LP vs. UA LP

1,3 1,3 1,3 1,9 1,9 1,9 1,9 1,9 1,9

79.92 22.15 6.79 12.86 3.51 3.37 0.20 3.07 2.04

0.003 0.018 0.080 0.006 0.094 0.099 0.667 0.114 0.187

1,3 1,3 1,3 1,9 1,9 1,9 1,9 1,9 1,9

Treatment Period Treatment * period

Wood Thrush 2,3 0.47 2,3 21.30 4,9 9.54

0.661 0.017 0.002

Contrasts PRE vs. EP PRE vs. LP EP vs. LP NH EP vs. EA EP NH LP vs. EA LP NH EP vs. UA EP NH LP vs. UA LP EA EP vs. UA EP EA LP vs. UA LP

1,3 1,3 1,3 1,9 1,9 1,9 1,9 1,9 1,9

0.011 0.018 0.482 0.011 0.087 0.113 0.435 0.167 0.026

F

25.68 9.07 1.24 0.01 0.00 0.25 1.72 0.37 1.57

32.29 24.22 0.62 10.27 3.69 3.08 0.67 2.26 7.12

P

34.32 16.70 0.42 0.90 0.47 0.00 0.83 0.91 2.55

P

0.278 0.220 0.088 0.001 0.712 <0.001 0.028 0.054 0.039

0.009 0.026 0.564 0.367 0.510 0.997 0.385 0.366 0.144

a The Ndf and Ddf are the numbers of degrees of freedom in the numerator and denominator, respectively, for calculating the F statistic. b The treatment main effect compares even-aged management (EA), uneven-aged management (UA), and no-harvest (NH) sites. c The period main effect compares data from 5 years of pre-treatment (1991– 1995), 7 years early post-treatment (1997, 1998, 1999, 2000, 2001, 2002, 2003), and 3 years late post-treatment (2008, 2009, 2010). d Period contrasts compare data between pre-treatment (PRE) and early posttreatment (EP), pre-treatment and late post-treatment (LP), and early post-treatment and late post-treatment. Treatment * period contrasts compare treatment effects within each post-treatment time period (early post-treatment = EP, late post-treatment = LP).

prior to harvest in 1996 (Fig. 3). Most early-successional species, except Prairie Warbler and Blue-winged Warbler, showed significant increases in density in the early post-harvest period in EAM and UAM compared to NH sites (Table 2). Density of Prairie Warbler was significantly higher in EAM, but showed no difference in density between UAM and NH management and were only detected during the early post-harvest period. Blue-winged Warbler was detected at relatively low density across the study area, but consistently had higher density in UAM and showed a moderate increase in EAM in the early post-harvest period. White-eyed Vireo and Yellow-breasted Chat had higher density in EAM than UAM sites. Most early-successional species exhibited marked declines

in density from the early post-harvest period (up to 7 years postharvest) to the late post-harvest period (12–14 years post-harvest), although this difference was significant only for Indigo Buntings (Table 2). By 2010, Indigo Bunting, Prairie Warbler, and Yellowbreasted Chat were absent from the study sites. The number of detections of Blue Jays declined from the pre-harvest period to the post-harvest period in all treatments (Table 3; Fig. 4). During the early and late post-harvest periods, the number of detections of Blue Jays was higher in NH and EAM treatments than in UAM treatments. Brown-headed Cowbird detections followed a similar temporal pattern (Table 3; Fig. 4). However, there were fewer detections in NH treatments than either harvested treatment during the early post-treatment period, but there was no treatment effect during the late post-harvest period (Table 3; Fig. 4). 3.2. Reproductive success Reliable data for our focal species were available from 2254 nests (1592 and 662 nest from mature forest and early-successional species, respectively). For mature forest species, except Acadian Flycatcher, a null model containing stage only was most supported and indicates nest survival did not vary with treatment or period (Table 4). For Acadian Flycatcher, the global model was most supported (Table 4), but when averaged across all models, treatment and period had negligible effects on nest survival (Fig. 5). The null temporal models were most supported for Hooded Warbler and Indigo Bunting (Table 5), suggesting no difference in nest survival in relation to treatment and time period. For Whiteeyed vireo and Yellow-breasted Chat combined, the most supported model was the global model that included the temporal effect of nest stage. The treatment by period interaction showed a trend for higher survival in EAM than UAM sites in the early postharvest period but no difference in survival between treatments during the late post-harvest period (Fig 6). The probability of brood parasitism did not differ with treatment, period, or treatmentxperiod interaction (P > 0.05), but varied between mature forest and early-successional species (F1,2 = 17.81, P = 0.05). The proportion of parasitized nests was higher for earlysuccessional species than mature forest species (mean = 0.05 ± 0.01 vs. 0.01 ± 0.003, respectively). 4. Discussion 4.1. Density 4.1.1. Mature forest species Contrary to our prediction, changes in density of all mature forest species, except Kentucky Warbler, were greater than the proportion of the area harvested. Our prediction that decline would be greater in UAM was not supported and Ovenbirds and Wood Thrushes responded most negatively to EAM. That Wood Thrush did not decline in UAM from pre-harvest to early post-harvest and then increased modestly in UAM during the late post-harvest period could reflect their preference for interior edge for nesting (Becker et al., 2011; Evans et al., 2011). Density of most mature forest species has not yet recovered to preharvest levels, even though basal area in EAM and UAM reached pre-harvest levels by 14 years post-harvest and increased by 14% in NH sites since the start of the study (Shifley, unpublished. data). The greater range of habitat alteration tolerated by mature forest species in other studies (Guenette and Villard, 2005; Vanderwel et al., 2007; Haché and Villard, 2010) could be attributed to higher regional population density that sustained recruitment to post-harvest

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Fig. 3. Mean density (±SE) of early-successional species in relation to experimental timber harvesting on the Missouri Ozark Forest Ecosystem Project. Shown are five years of pre-treatment data (1991–1995), seven years of early post–treatment data (1997–2003) and three years of late post–treatment data (2008–2010).

plots and buffered precipitous local declines. For example, Haché and Villard (2010) reported pre-harvest density of Ovenbirds that were twice as high as in this study and recruitment to partial-harvested stands was sustained by a high proportion of young males. Higher densities could also be attributed to more productive habitat and higher food abundance (Haché et al., 2013). The greater than expected decrease in mature forest species across all treatments, including NH, could be partly explained by several non-mutually exclusive processes. First, it is likely that habitat suitability has changed over time for species like Wood Thrush that favor the mid-story canopy for nesting (Evans et al., 2011). The NH sites and reserves within the harvested sites contained mature trees that were 70–90 years old by the 15th year post-harvest. The decline of Wood Thrush on NH sites could be explained by increasing successional age as Wood Thrushes have disappeared from undisturbed mature second growth forests in the northeast (Holmes and Sherry, 2001). Additionally, red oak species (Quercus section Lobatae) declined on the study sites from 1992 to 2002 due to mortality and this caused substantial crown dieback

(Kabrick et al., 2008). Cumulative basal area of snags was higher in the NH treatment than the EAM and UAM treatments (5.1 m2/ ha, 3.7 m2/ha, 3.8 m2/ha, respectively; S. Shifely, unpublished data). This difference is likely associated with the presence of more red oak species group trees in the NH sites because harvesting in EAM and UAM treatments targeted red oak species for removal due to their vulnerability to mortality. This dieback could have caused small forest openings in the no harvest sites that were avoided by mature forest species during subsequent settlement periods. However, we believe changes in population density due to forest succession ought to be gradual, and does not explain lower density in NH sites immediately after harvest. Second, the removal of trees could have removed available food resources for insectivorous songbirds. Summerville (2011) found 40% fewer Lepidopteran species in sites managed with EAM and UAM than in control sites following experimental harvest in Indiana. Edge effects can also create drier conditions (Chen et al., 1999) that are unfavorable to deep litter invertebrates favored by Ovenbirds (Burke and Nol 1998). Similarly, Haché et al. (2013) showed a

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Table 2 Mixed model analysis of variance results of even-aged and uneven-aged forest management (treatment) on density of early successional birds. Significant effects are in bold. Source

Ndf,Ddfa

Treatmentb Periodc Treatment * period

Blue–winged Warbler 2,3 11.18 0.041 2,3 1.10 0.437 4,9 1.33 0.331

Hooded Warbler 2,3 19.09 0.019 1,2 11.18 0.079 2,6 6.13 0.035

Contrastsd PRE vs. EPe PRE vs. LP EP vs. LP NH EP vs. EA EP NH LP vs. EA LP NH EP vs. UA EP NH LP vs. UA LP EA EP vs. UA EP EA LP vs. UA LP

1,3 1,3 1,3 1,9 1,9 1,9 1,9 1,9 1,9

0.277 0.996 0.999 0.039 0.691 0.017 0.153 0.240 0.997

– – 1,2 1,6 1,6 1,6 1,6 1,6 1,6

– – 11.18 28.16 9.41 24.84 10.93 1.35 0.07

– – 0.079 0.002 0.022 0.002 0.016 0.289 0.807

Treatment Period Treatment * period

Indigo Bunting 2,3 9.36 2,3 72.41 4,9 4.97

0.051 0.003 0.021

Prairie Warbler 2,3 30.76 – – – –

0.010 – –

Contrasts PRE vs. EP PRE vs. LP EP vs. LP NH EP vs. EA EP NH LP vs. EA LP NH EP vs. UA EP NH LP vs. UA LP EA EP vs. UA EP EA LP vs. UA LP

1,3 1,3 1,3 1,9 1,9 1,9 1,9 1,9 1,9

0.001 0.940 0.003 0.001 0.833 <0.001 0.056 0.688 0.076

– – – 1,3 – 1,3 – 1,3 –

– – – 0.019 – 0.171 – 0.007 –

Treatment Period Treatment * period Blockf

White-eyed Vireo 2,3 0.01 1,2 0.00 2,6 3.17 2,3 2.04

0.991 0.985 0.115 0.276

Yellow-breasted Chat 2,3 0.37 0.721 2,3 11.14 0.041 4,9 3.11 0.073 2,3 8.78 0.056

Contrasts PRE vs. EP PRE vs. LP EP vs. LP NH EP vs. EA EP NH LP vs. EA LP NH EP vs. UA EP NH LP vs. UA LP EA EP vs. UA EP EA LP vs. UA LP

– – 1,2 1,6 1,6 1,6 1,6 1,6 1,6

0.985 0.022 0.978 0.038 0.977 0.041 0.213

1,3 1,3 1,3 1,9 1,9 1,9 1,9 1,9 1,9

F

1.75 0.16 0.89 5.77 0.17 8.43 2.48 1.58 0.00

107.14 0.01 68.10 42.96 0.05 48.73 4.82 0.17 4.02

– – 0.00 9.47 0.00 7.05 0.00 6.69 1.94

P

Ndf,Ddf

F

– – – 20.79 – 3.22 – 42.39 –

22.28 0.00 0.00 42.95 0.00 33.31 0.00 19.58 0.01

P

0.018 0.978 0.968 <0.001 0.9822 <0.001 0.982 0.002 0.939

a The Ndf and Ddf are the numbers of degrees of freedom in the numerator and denominator, respectively, for calculating the F statistic. b The treatment main effect compares even-aged management (EA), uneven-aged management (UA), and no-harvest (NH) sites. c The period main effect compares data from 5 years of pre-treatment (1991– 1995), 7 years early post-treatment (1997, 1998, 1999, 2000, 2001, 2002, 2003), and 3 years late post-treatment (2008, 2009, 2010). d Period contrasts compare data between pre-treatment (PRE) and early posttreatment (EP), pre-treatment and late post-treatment (LP), and early post-treatment and late post-treatment. Treatment * period contrasts compare treatment effects within each post-treatment time period (early post-treatment = EP, late post-treatment = LP). e Hooded warbler, White-eyed vireo, and Prairie warbler were not detected on the study sites prior to treatments, so pre-treatment period was not included in analysis. f Due to model convergence problems when block was a random effect, block was included as a fixed effect for White-eyed Vireo and Yellow-breasted Chat.

negative effect of UAM on litter invertebrates. However, others have found clearcutting benefits litter invertebrates (Keller et al., 2003), including ants (Palladini et al., 2007), which are favored by Ovenbirds (Porneluzi et al., 2011). Additionally, mature forest species are known to use early-successional habitat during breeding (Pagen et al., 2000; Major and Desrochers 2012) suggesting the availability of alternative food sources in young forests. Whether logging or crown dieback negatively affected food availability for

Table 3 Mixed model analysis of variance results of forest management on density of Blue Jays and Brown-headed Cowbirds. The treatment main effect compares even-aged management (EA), uneven-aged management (UA), and no-harvest (NH) sites. The period main effect compares data from 5 years of pre-treatment (1991–1995), 7 years early post-treatment (1997–2003) and 2 years late post–treatment (2008–2010). Source

Ndf,Ddfa

F

P

Treatment Period Treatment * period

Blue Jay 2,3 2,3 4,9

6.88 25.75 1.13

0.076 0.013 0.399

Contrastsb PRE vs. EP PRE vs. LP EP vs. LP NH EP vs. EA EP NH LP vs. EA LP NH EP vs. UA EP NH LP vs. UA LP EA EP vs. UA EP EA LP vs. UA LP

1,3 1,3 1,3 1,9 1,9 1,9 1,9 1,9 1,9

24.24 46.61 9.81 2.48 0.06 11.23 4.78 3.20 3.79

0.016 0.006 0.052 0.149 0.813 0.008 0.056 0.107 0.083

Treatment Period Treatment * period

Brown-headed Cowbird 2,3 3.30 2,3 64.23 4,9 5.05

0.174 0.003 0.021

Contrasts PRE vs. EP PRE vs. LP EP vs. LP NH EP vs. EA EP NH LP vs. EA LP NH EP vs. UA EP NH LP vs. UA LP EA EP vs. UA EP EA LP vs. UA LP

1,3 1,3 1,3 1,9 1,9 1,9 1,9 1,9 1,9

0.014 0.001 0.004 0.001 0.586 0.001 0.229 0.987 0.484

26.86 127.33 65.70 21.63 0.32 21.83 1.66 0.00 0.53

a The Ndf and Ddf are the numbers of degrees of freedom in the numerator and denominator, respectively, for calculating the F statistic. b Period contrasts compare data between pre-treatment (PRE) and early posttreatment (EP), pre-treatment and late post-treatment (LP), and early post-treatment and late post-treatment. Treatment * period contrasts compare treatment effects within each post-treatment time period (early post-treatment = EP, late post-treatment = LP).

mature forest birds and could have altered habitat suitability and density of these species in the no harvest sites is unclear. Lower densities in NH sites during both post-harvest periods could have resulted from broad-scale regional population declines of forest obligate species and Neotropical migrants (Sauer and Link, 2011). However, Breeding Bird Survey (BBS) data show no declines for these focal species in Missouri (Sauer et al., 2011). Researchers studying other components of MOFEP also recorded significant changes on the NH sites during the post-treatment years. Plant species richness and the abundance of small mammals, reptiles, and amphibians decreased on the NH sites (Kabrick et al., 2002). The decreases on the NH sites in several different groups of organisms immediately following the harvest treatments may be the result of forestry on adjacent treated sites or an environmental event such as drought coincided with the MOFEP experiment. In no harvest sites, abundance of mature forest species has remained 35% lower than pre-harvest levels for 14 years. It seems unlikely that this immediate and long-lasting decrease is explained by a coincidental environmental change. It is possible that a change in regional timber harvest has occurred and resulted in regional changes that could impact distribution or settlement patterns of migratory songbirds (Boulinier et al., 2001; Thompson et al., 2002). While Missouri’s total acreage of forestland and growing stock volume has increased over the last 50 years, annual net growth has declined by about 100 million cubic feet from 1999–2003 to 2004–2008 and annual removals have increased about 65 million cubic feet in that same period

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MOFEP takes place. Results following the second round of harvest in 2011 may provide additional insight into the temporal and spatial scales over which timber harvest impacts or thresholds can be observed. At this time, we cannot rule out that the impacts extend over several kilometers and several hundred hectares.

Fig. 4. Mean detections (±SE) of Brown-headed Cowbirds and Blue Jays in relation to experimental timber harvesting on the Missouri Ozark Forest Ecosystem Project. Shown are five years of pre-harvest data (1991–1995), seven years of early posttreatment data (1997–2003) and two years of late post-treatment data (2008– 2010).

4.1.2. Early-successional species Prior to harvest, only three of the six early-successional focal species occurred on the study sites. Early-successional species colonized the harvested sites within 1–2 years after treatment and peaked in density 6–8 years post-harvest (P. Porneluzi, unpublished data), a pattern also observed in other regions (Thompson and DeGraaf, 2001; Keller et al., 2003; Schlossberg and King, 2009). By 12–14 years post-harvest, Indigo Bunting, Prairie Warbler and Yellow-breasted Chat had disappeared from the study sites, but Hooded Warbler, Blue-winged Warbler, and White-eyed Vireo were still detected at low density. Most early-successional species were more abundant in sites treated with EAM than UAM, especially the White-eyed Vireo and Prairie Warbler (see also Annand and Thompson, 1997). The small group selection cuts in our study (21–43 m in diameter) and single tree selection cuts were likely less suitable for Prairie Warbler, White-eyed Vireo, and Yellow-breasted Chat, which prefer early-successional patch sizes >2 ha (Brito-Aguilar, 2005; Shake et al., 2012). Contrary to previous studies (Annand and Thompson, 1997; Eng et al. 2011), we found higher density of Hooded Warbler in clearcuts than in group-selection cuts during the early post-harvest period. It appears that Hooded Warbler prefer small clearcuts, (<25 ha; see also Moorman et al., 2002) and Brito-Aguilar, 2005). Tree regeneration following timber harvest provides important habitat for early-successional species by increasing stem density and leaf area (Keller et al., 2003). Zenner et al. (2006) reported regeneration of woody and vegetative growth was higher in clearcuts than in group selection cuts 5 years post-harvest on our study sites, indicating a likely driver behind our results. 4.2. Nest survival

(Raeker et al., 2010). Raeker et al. (2010) expressed concern for potential overharvest in Carter, Reynolds, and Shannon counties where harvest pressure is greatest, the same three counties where

Nest survival did not decline as predicted in harvested sites and there was little variation in relation to treatment or time period for

Table 4 AIC results for nest survival of mature forest species in the Missouri Ozark Forest Ecosystem Project (1991–2010). The model parameters are described in the text. n = number of observation intervals for each species. K = number of parameters in each model. DAIC = the difference between the AIC values for the most-supported model and the given model. wi = weight of evidence for each model.

a

Modela

k

DAIC

wi

Period + treat + (period  treat) + stage + date + date2 Period + stage + date + date2 Null (stage + date + date2)

Acadian Flycatcher (n = 9701) 14 2159.28 8 2159.65 6 2159.65

0.00 0.36 0.36

0.38 0.31 0.31

Null (stage) Period + stage Period + treat + (period  treat) + stage

Kentucky Warbler (n = 396) 4 114.30 6 117.07 9 117.40

0.00 2.77 3.10

0.68 0.17 0.14

Null (stage) Period + stage Period + treat + (period  treat) + stage

Ovenbird (n = 1480.5) 4 5 12

0.00 1.95 7.48

0.71 0.27 0.02

Null (stage) Period + stage Period + treat + (period  treat) + stage

Worm-eating Warbler (n = 3111) 4 439.20 6 442.42 12 451.40

0.00 3.24 12.35

0.83 0.17 0.00

Null (stage) Period + stage Period + treat + (period  treat) + stage

Wood Thrush (n = 3111) 4 758.18 6 761.04 12 767.92

0.00 2.85 9.73

0.80 0.19 0.01

Block was included as a nuisance variable in each model.

AIC

582.27 584.25 589.78

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Fig. 5. Model-based predictions of period nest survival (±95% CI) for mature forest species as a function of treatment and time period in the Missouri Ozark Forest Ecosystem Project, Missouri, USA, 1991–2010.

Table 5 AIC results for nest survival of early successional species in the Missouri Ozark Forest Ecosystem Project (1991–2010). The model parameters are described in the text. n = number of observation intervals for each species. k = number of parameters in each model. DAIC = the difference between the AIC values for the most-supported model and the given model. wi = weight of evidence for each model.

a b

Modela

k

DAIC

wi

Null (date) Period + date Period + treat + (period  treat) + date

Hooded Warbler (n = 814) 4 239.76 5 240.75 7 243.21

0.00 0.98 3.45

0.56 0.34 0.10

Null (stage + date) Period + stage + date Period + treat + (period  treat) + stage + date

Indigo Bunting (n = 4108) 5 982.37 7 985.65 13 992.46

0.00 3.29 10.16

0.83 0.16 0.01

Period + treat + (period  treat) + stage + species Species + stage Null (stage) Period + stage

Other Sppb (n = 1167) 8 5 4 5

0.00 1.53 2.08 3.94

0.51 0.24 0.18 0.07

AIC

268.64 270.17 270.72 272.58

Block was included as a nuisance variable in each model. Due to small sample sizes, observations for White-eyed Vireo and Yellow-breasted Chat were combined and the variable species was added as a fixed effect.

late-successional or early-successional species, similar to several recent studies showing no effect of select cutting on nest survival (LeBlanc et al., 2011; Haché et al., 2013). The only trend supporting this prediction was slightly higher nest survival for Acadian Flycatcher in no harvest sites compared to EAM and UAM. The reduction of mast-crop production associated with harvest could have been great enough to trigger resource-pulse events (Schmidt et al., 2008; Leblanc et al., 2011). However, it is unknown whether nest predators such as raptors, corvids, and snakes in our study area (see Cox et al., 2012b for detail) show important numerical re-

sponses as a function of resource-pulses. Moreover, little is known about the effects of forestry on most of these predator species (but see King et al. 1998). It is not likely that the predator community was greatly altered by the first harvest entry in this contiguously forested landscape (Oehler and Litvaitis, 1996). Mesopredators that might be expected to increase in response to early succession habitat (Chalfoun et al., 2002b) appear to be less important as nest predators in our study area (Cox et al., 2012b). In our study, the only nest predator that we surveyed, Blue Jay, declined in number of detections from pre-harvest to the late post-harvest period.

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Fig. 6. Model-averaged predictions of period nest survival (±95% CI) for early-successional species as a function of treatment and time period in the Missouri Ozark Forest Ecosystem Project, Missouri, USA, 1991–2010.

According to BBS data, Blue Jay populations showed a negative population trend (1%, 95% CI: 2.5, 0.5) between 2000 and 2010 in Missouri (Sauer et al., 2011). Blue Jay declines could coincide with the appearance of West Nile Virus (WNV) in the United States in 1999. Blue Jays suffer high mortality rates when exposed to the virus (Komar et al., 2003; but see LaDeau et al., 2007). As predicted, probability of parasitism did not increase during either post-harvest period and brood parasitism rates were higher for early-successional species than for mature forest species. In harvested sites, Brown-headed Cowbirds could have benefited from increased perch sites and visibility (Clotfelter, 1998; Hauber and Russo 2000) facilitating host detection (Banks and Martin, 2001; but see Saunders et al., 2003). The temporal decline in Brown-headed Cowbird detections in our study coincides with regional population declines (Sauer et al., 2011) and declining parasitism rates in the Midwest (Cox, W.A. unpublished data). Perhaps the decline in Cowbird detections on our sites could be driven also by the temporal decline in host density, but with parasitism rates less than 5%, it is unlikely that host nests are limiting. 4.3. Management implications Our results suggest that even in contiguous forests, openings created by EAM and UAM can result in the decline of some mature forest species and can reduce carrying capacity for many years. Of the mature forest species, Ovenbirds responded most negatively to harvest regimes and declined in NH sites, indicating even low levels of forestry could impact their populations (Vanderwel et al., 2007). Both harvest regimes provided habitat for several early-successional species that were not present prior to harvest, although they responded most positively to EAM. The peak of abundance occurred 7 years after harvest with most shrub-dependent species disappearing by 12 years after harvest (see also Goodale et al., 2009). To better retain populations of mature forest and early-successional species, EAM should be implemented on fewer stands with a shorter re-entry period of 8-10 years to maintain peak abundance of early-successional species (Schlossberg and King 2009; McDermott and Bohall-Wood, 2009) Assuming harvest levels are reduced by half, then it should take half the time to regenerate basal area and thus, a shorter re-entry period could

be a sustainable approach in central hardwoods. Wood Thrush, however, will likely need some modest level of UAM to provide more suitable breeding habitat. Here we have shown timber harvest of a sustainable intensity within a largely forested landscape does not negatively affect reproductive success. Timber harvest impacts birds by leading to changes in abundance that result from changes in habitat suitability. The population decline observed in mature forest species across harvested and NH sites indicates that the value of this forested region as a population source has likely diminished for these species (Robinson et al., 1995; Porneluzi and Faaborg, 1999). Continued monitoring of bird populations on these study sites, as well as monitoring harvesting trends in the greater Ozark region, will be necessary to determine the relative importance of local versus regional habitat on bird populations. Our results can be used to generate population models projecting the impact of alternative forest management plans. Those predictions could be tested against the data after the harvest of additional mature trees during the second 15-year re-entry on the MOFEP sites that occurred in 2011. This should provide important additional insights into the broader impact of harvest effects within this large forested landscape. The Missouri Ozark population is at the western end of the breeding range of most species monitored in this study. This could explain why this population may not be as resilient as others (Flaspohler et al., 2001; Haché and Villard, 2010; Newell and Rodewald, 2012). We anticipate comparison of results with similar experiments such as the Hardwood Ecosystem Experiment in Indiana (Malloy and Dunning, 2013) to expand evaluation of forest management systems on breeding birds across the spectrum of the central hardwood forests. Acknowledgements This project would not have been possible without the hard work of >250 dedicated interns over the years. A. Cox and S. Wolken provided helpful comments on an early draft and we thank 2 referees for their helpful suggestions. Missouri Department of Conservation funded this study. The National Fish and Wildlife Foundation and the Missouri Federal Aid in Wildlife Restoration Project W-13-R contributed additional funds for parts of the study.

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Table A1 AIC results from modeling of nest stage and Julian date on nest survival of mature forest species in the Missouri Ozark Forest Ecosystem Project. k = number of parameters in each model. DAIC = the difference between the AIC values for the most-supported model and the given model. wi = Akaike weight for each model. Models with the lowest DAIC values and highest wi values have the most support. Model

K

Stage, date (date)2 Date (date)2 Stage date Date Stage

Acadian Flycatcher (n = 9701) 4 2159.79 0.00 3 2162.91 3.11 3 2165.59 5.79 2 2165.89 7.09 2 2166.89 26.98

AIC

DAIC

0.77 0.16 0.05 0.03 0.00

Stage Date Stage, date Date (date)2 Stage, date (date)2

Kentucky 2 2 3 3 4

Warbler (n = 396) 124.85 125.63 126.79 126.88 128.31

0.00 0.78 1.94 2.03 3.46

0.38 0.26 0.15 0.14 0.07

Stage Stage, date Stage, date, (date)2 Date (date)2 date

Ovenbird 2 3 4 3 2

(n = 1480.5) 486.20 487.53 487.71 487.83 488.18

0.00 1.33 1.50 1.62 1.97

0.36 0.18 0.17 0.16 0.13

Stage Date Date (date)2 Stage, date Stage, date (date)2

Worm-eating Warbler (n = 3111) 2 437.59 0.00 2 437.72 0.13 3 438.92 1.33 3 439.59 2.01 4 440.79 3.20

0.33 0.31 0.17 0.12 0.07

Stage Stage, date Date Stage, date (date)2 Date (date)2

Wood Thrush (n = 3111) 2 755.61 3 757.37 2 758.57 4 758.80 3 759.73

0.51 0.21 0.11 0.10 0.07

0.00 1.76 2.96 3.18 4.12

wi

Table A2 AIC results from modeling of nest Stage and Julian date on nest survival of earlysuccessional species in the Missouri Ozark Forest Ecosystem Project. n = number of observation intervals for each species. k = number of parameters in each model. DAIC = the difference between the AIC values for the most-supported model and the given model. wi = Akaike weight for each model. Models with the lowest DAIC values and highest wi values have the most support. Model

K

AIC

DAIC

wi

Date Date (date)2 Stage, date Stage Stage, date (date)2

Hooded Warbler (n = 814) 2 235.75 3 237.76 3 237.76 2 237.84 4 239.78

0.00 2.01 2.01 2.09 4.03

0.45 0.16 0.16 0.16 0.06

Stage, date Stage Stage, date, (Date)2 Date Date (date)2

Indigo Bunting (n = 4108) 3 979.36 2 980.03 4 980.64 2 991.07 3 991.68

0.00 0.67 1.28 11.71 12.32

0.44 0.32 0.23 0.00 0.00

Stage Stage, date (date)2 Stage, date Date (date)2 Date

Other Sppa (n = 1167) 2 267.85 4 268.99 3 269.19 3 273.30 2 274.19

0.00 1.14 1.34 5.44 6.34

0.46 0.26 0.23 0.03 0.02

a Due to small sample sizes, observations for White-eyed Vireo and Yellowbreasted Chat were combined for analysis and the variable species was added as a fixed effect.

Appendix A See Tables A1 and A2.

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