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Large, infrequent disturbance on a regulated river: response of floodplain forest birds to the 2011 Missouri River flood ESZTER C. MUNES,1,4, MARK D. DIXON,1 DAVID L. SWANSON,1 CHRISTOPHER L. MERKORD,2 AND

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ADAM R. BENSON3

1 Department of Biology, University of South Dakota, Vermillion, South Dakota 57069 USA Geospatial Sciences Center of Excellence, South Dakota State University, Brookings, South Dakota 57007 USA 3 20861 467th Avenue, Brookings, South Dakota 57006 USA

Citation: Munes, E. C., M. D. Dixon, D. L. Swanson, C. L. Merkord, and A. R. Benson. 2015. Large, infrequent disturbance on a regulated river: response of floodplain forest birds to the 2011 Missouri River flood. Ecosphere 6(11):212. http://dx.doi.org/10.1890/ES15-00007.1

Abstract. Floodplain forests are dynamic habitats that support a high diversity and abundance of birds. Periodic flood disturbance is important in the establishment and maintenance of the heterogeneous mosaic of vegetation communities across the riverine landscape. Human suppression of disturbance regimes has been implicated in the decline of bird species in these systems. Because few large rivers are not subject to flood control by dams and levees, opportunities to study avian responses to flood disturbance are limited. A large magnitude, long-duration flood event on the Missouri River, USA, during the summer of 2011 provided an opportunity to quantify post-flood changes in forest bird densities and species richness relative to pre-flood conditions on a riverine floodplain impacted by decades of flow regulation. We surveyed 75 forest sites on two segments of remnant floodplain forest along the Missouri National Recreational River (MNRR) in southeastern South Dakota and northeastern Nebraska and examined changes in density for 35 breeding landbird species from pre-flood (2009–2010) to post-flood (2012–2014) periods. We used a repeated measures ANOVA design to test the effects of year on average densities of birds and nesting guilds and confidence intervals to determine changes in densities of individual species and species richness. 19 of 35 focal species declined significantly one year after the flood (2012), but abundances of ten species recovered to pre-flood densities or higher within two years. In 2012, density declines of six species and density increases of two species were significantly correlated with a decrease in woody vegetation density and percent shrub cover. Average bird densities and the density of shrub nesters rebounded in 2013 to pre-flood levels and continued to increase through 2014. There were no significant changes in species richness at the level of forest habitat types between sampling years. Our results demonstrate short-term resilience of floodplain bird species to a major disturbance despite declines in early successional habitat and minimal recovery of woody vegetation. Key words: avian ecology; bird habitat; cottonwood; floodplain; forest succession; riparian forest. Received 5 January 2015; revised 30 April 2015; accepted 6 May 2015; published 9 November 2015. Corresponding Editor: D. P. C. Peters. Copyright: Ó 2015 Munes et al. This is an open-access article distributed under the terms of the Creative Commons Attribution License, which permits unrestricted use, distribution, and reproduction in any medium, provided the original author and source are credited. http://creativecommons.org/licenses/by/3.0/ 4

Present address: 6808 West Yakima Avenue, Yakima, Washington 98908 USA.

E-mail: [email protected]

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western North America, as it is in these generally more arid habitats where riparian zones are particularly important to birds (e.g., Carlisle et al. 2009). Along the Missouri River, plains cottonwood (Populus deltoides Marsh. subsp. monilifera (Ait.) Eckenw.; hereafter cottonwood) forests support a high richness of bird species (Swanson 1999) and serve as critical breeding and stopover habitat, particularly for Neotropical migrants (Carter and Barker 1993), which make up about one-half of the nesting species observed in the Missouri River forests (Liknes et al. 1994). Many species, such as the Warbling Vireo (Vireo gilvus), are strongly associated with cottonwood habitats (Rumble and Gobeille 2004), and are found in greater densities there than in other woodland habitat types (Gentry et al. 2006). Riparian forests are also important for species such as the Eastern Wood-Pewee (Contopus virens), Red-headed Woodpecker (Melanerpes erythrocephalus), Bell’s Vireo (Vireo bellii ), and Ovenbird (Seiurus aurocapillus), which have shown regional population declines (Sauer et al. 2014). Historically, the recruitment of cottonwood forests on the Missouri River was correlated with annual, bi-modal spring and summer flooding, followed by summer low-flows (Johnson et al. 1976, Scott et al. 1997, Galat et al. 2005). Since the 1890s, land conversion to agriculture has decreased the areal extent of floodplain forests by nearly 50% along unchannelized floodplain segments of the Missouri River in the northern Great Plains and much additional forested land has been lost to inundation by reservoirs (Dixon et al. 2012). Conversion of forest to agriculture has been compounded by reduced cottonwood recruitment, a phenomenon that has been widely observed on meandering river floodplains in western North America following construction of dams (Johnson et al. 1976, Rood and Mahoney 1990, Friedman et al. 1998, Rood et al. 2005). Declines in sandbar formation related to reductions in river dynamism and sediment transport have limited new recruitment of cottonwood and other early successional woody species on the Missouri River (Johnson 1992, Dixon et al. 2012, Johnson et al. 2012). The demography of cottonwood forest stands remaining along the Missouri River is shifting towards older stands, which become increasingly dominated by late-successional tree species (Johnson 1992, Dixon et al.

INTRODUCTION Flooding is an important aspect of large river systems that drives patch dynamics and initiates ecological processes necessary for vegetation dispersal and establishment (Jones et al. 1994), thereby affecting the structure and composition of riparian ecosystems (Junk et al. 1989, Hupp and Osterkamp 1996). Natural flooding regimes are marked by periodic extreme hydrological events (large and infrequent disturbances; LIDs) that reshape channel and habitat configuration, as well as smaller and more frequent events that help sustain existing riparian ecosystems. The effects of and recovery from LIDs is spatially and temporally non-uniform (Parsons et al. 2006) and dependent on components that include event frequency, intensity, severity, size, rotation period, return interval, and biotic residuals (Turner et al. 1998). As the ecological importance of LIDs is increasingly recognized (Brawn et al. 2001, Turner 2010), so is the potential for their restorative capacity to degraded ecosystems. In regulated rivers, however, the post-flood recovery trajectory is dictated by a complex interaction between flooding and anthropogenic influences (Reich and Lake 2014). By attenuating flooding magnitude and frequency, dams and other flood control infrastructure greatly impact riparian vegetation and the biota along rivers in North America (Rood and Mahoney 1990, Friedman et al. 1998, Marston et al. 2005, Andersen et al. 2007) and elsewhere (Kingsford 2000, Mumba and Thompson 2005). However, LIDs are likely to occur even on regulated rivers (e.g., Allen 1993), particularly under projections of climate change (Milly et al. 2002, Palmer et al. 2008, IPCC 2012). A major flood on the Missouri River, USA, in the summer of 2011 provided an opportunity to investigate and quantify the responses of floodplain forests and their bird communities to a large and infrequent disturbance after six decades of comparative stasis due to flow regulation. In the Great Plains region of North America, floodplain forests support higher diversity and abundance of landbird species relative to other habitat types (Knopf et al. 1988, Finch and Ruggiero 1993, Knutson et al. 1995, Best et al. 1996). This pattern of high avian abundance in floodplain forests is applicable over much of v www.esajournals.org

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2012, Johnson et al. 2012) and by upland or exotic species of shrubs and small trees in the understory (Dixon et al. 2010, Greene and Knox 2014). Reduced rates of cottonwood forest recruitment under regulated flows are likely to decrease landscape heterogeneity and forest area, and thus reduce overall biodiversity in the decades to come (Johnson 1992). The Missouri River flood of 2011 was an extreme hydrological event that resulted from high antecedent soil moisture, heavy Rocky Mountains and Great Plains snowpacks, and extreme spring rains in the upper Missouri River basin. The combination of factors led to a record 59.7 cubic kilometers (km3) of runoff between March and July, 2011. In June, release of water from Gavins Point Dam in Yankton, South Dakota reached 4,551 cubic meter per second (m3s1), which was more than double the previous post-dam record of 1,982 (m3s1) in 1997 and the highest since the flood of record (peak of 11,213 m3s1) in 1952 (Fig. 1A). Pre-dam floods produced a large spike in the hydrograph which then tapered off quickly, but controlled releases by the U.S. Army Corps of Engineers in summer and fall 2011 caused inundation for a much longer period, nearly three months in some areas (USACE 2012; Fig. 1B). As indicators of ecological change, alterations of bird abundances and distributions can reveal the effects of anthropogenic stresses and natural disturbances, such as the impact of forest loss on ecosystem health (Canterbury et al. 2000, Scott et al. 2002). Declines in terrestrial bird abundances in North America have been partially attributed to the modification of natural disturbance regimes and successional processes (Askins 2000). Conservation of avifauna requires an understanding of relationships between avian habitat requirements and the processes that create and maintain these habitats (Askins 2000, Brawn et al. 2001). With the availability of pre-flood bird survey data (Benson 2011) from two unchannelized reaches of the Missouri River, we resurveyed the same sites post-flood to quantify and compare vegetation and bird densities for each period. We repeated bird surveys for three breeding seasons post-flood to measure the recovery trajectory. Few studies have measured the effect of floodinduced habitat changes on bird species or v www.esajournals.org

communities (e.g., Knopf and Sedgwick 1987, Knutson and Klaas 1997), especially relative to pre-flood conditions. Bird densities and species richness often decrease during and immediately after disturbance events (Knutson and Klaas 1997) with various recovery responses in the subsequent years following disturbance, including both resiliency and continued depression, often dependent on the severity of the disturbance (Knopf and Sedgwick 1987, Waide 1991, Bontrager et al. 1995, Barlow and Peres 2004). In addition, different bird species respond differently to disturbance events, with some being relatively unaffected, but others showing longerterm impacts (Knopf and Sedgwick 1987, Askins and Ewert 1991, Torres and Leberg 1996, Woinarski and Recher 1997). Thus, avian responses to LIDs are difficult to generalize, in large part because of few studies having appropriate before-after data sets to address these questions. In the present study, we quantified the overall response of the forest bird community to the 2011 flood by estimating bird densities and species richness before (2009 and 2010) and after (2012, 2013, and 2014) the flood across a range of floodplain forest ages. Because vegetation structure is an important feature driving habitat selection for birds and influences bird abundances (MacArthur and MacArthur 1961, MacArthur et al. 1962, Karr and Roth 1971, Cody 1981, James and Wamer 1982, Miller et al. 2004), we expected changes in bird density and species richness to result from flood-induced changes in vegetation, particularly in early successional forest that was impacted the most by flooding (Dixon et al. 2015). Concurrently, we hypothesized that bird species that specialize in young forests, such as Bell’s Vireo and Willow Flycatcher (Empidonax traillii; Knutson et al. 2005, Thogmartin et al. 2009), as well as shrub nesters as a guild, would experience greater declines than mature forest birds or those that nest in cavities or the forest canopy. Because replacement of comparable landbird habitat via recruitment and succession takes 1–2 decades, we expected bird densities to remain low throughout the sampling period.

METHODS Study area The study area consists of floodplain forests 3


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Fig. 1. Historical flows at Gavins Point Dam, Yankton, South Dakota (A) between 1930 and 2014, and (B) during three high water years (1944, 1949, 2011). Graphs use data from Gavins Point Dam flow releases (USACE) from August 1955 to present and data from the nearby Yankton USGS gage (#06467500) from 1930 to 1955.

morphology of the MNRR resemble those of the pre-dam river, their geomorphological and ecological characteristics are modified by the presence of dams and reservoirs (Jacobson et al. 2010). The 39-mile (63 km) segment begins at Fort Randall Dam and extends downstream to Lewis and Clark Reservoir, ending approximately at the confluence of the Niobrara and Missouri rivers. Farther downstream, the 59-mile (95 km) segment begins at Gavins Point Dam, situated approximately three miles west of Yankton,

with open and closed canopies and early successional shrublands (hereafter, forests) along two (‘‘39-mile’’ and ‘‘59-mile’’ or segments 8 and 10, respectively) segments of the Missouri River within the Missouri National Recreational River (MNRR) in southeastern South Dakota and northeastern Nebraska (Fig. 2). The MNRR is managed by the National Park Service under the Wild and Scenic Rivers Act, in cooperation with the United States Army Corps of Engineers. While remnant floodplain forests and channel v www.esajournals.org

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Fig. 2. Missouri National Recreational River study segments. Segments 8 and 10 represent the ‘‘39-mile’’ and ‘‘59-mile’’ reaches of the MNRR, respectively.

South Dakota, and continues southeast to Ponca, Nebraska. No dams occur below the 59-mile segment, although the lower 1,200 km has been channelized for barge navigation. The vegetation communities at all study sites consist of riparian forests of varying age and composition. Historically on the Missouri River, recruitment of cottonwood occurred on bare, moist surfaces established by lateral migration of the river channel (Johnson 1992) and, on geologically constrained reaches, by overbank flooding and deposition (Scott et al. 1997). Early successional vegetation, characterized by young cottonwood and willow (Salix interior, S. amygdaloides, and S. lutea), occurred closest to the river channel, followed by older gallery forests of large-diameter cottonwoods farther v www.esajournals.org

from and higher above the channel. Over time, large cottonwoods were gradually replaced by green ash (Fraxinus pennsylvanica), box elder (Acer negundo), and American elm (Ulmus americana) to form ‘‘post-cottonwood’’ equilibrium forest communities, as cottonwoods cannot reproduce successfully under a forest canopy (Johnson et al. 1976, Hesse et al. 1988, Johnson et al. 2012). Currently, the dominant shrubs of the understory of mature forests are roughleaf dogwood (Cornus drummondii ), common buckthorn (Rhamnus cathartica), and eastern red-cedar (Juniperus virginiana) (Dixon et al. 2010). Colonization of these stands by red-cedar has primarily occurred over the last 55 years, after completion of adjacent upstream dams (Greene and Knox 2014). Younger (post-dam, ,55 years old) non5


MUNES ET AL. Table 1. Forest age (relative to 2012) and type categories used for bird and vegetation sampling. Forest category

Age

We conducted bird surveys pre-flood in 2009 and 2010 (Benson 2011) and post-flood in 2012, 2013, and 2014 using the same sampling points and point count survey methodology. To ensure that mostly breeding species were detected, surveys took place between late May and early July in all years, with two visits to each site per season. For the post-flood surveys, each site was visited once by E. Munes and once by a trained assistant in each field season. One visit occurred during the first half of the sampling period (approximately May 25–June 15), and the second visit during the latter half (approximately June 16–July 5). Repeat visits at an individual site took place at least 10 days apart. We visited 2–4 sites each morning, began point counts at dawn, and concluded by 0930 hr. We surveyed each point for ten minutes and recorded each individual bird seen or heard, along with its species identification, estimated distance (in meters) from observer, and time of first detection. We used laser rangefinders to estimate distance when possible. Data were collected on a handheld digital voice recorder and then transcribed into a Microsoft Access database.

Dominant tree species

Pole cottonwood Intermediate cottonwood Mature cottonwood Old cottonwood

15–29 30–55

Cottonwood Cottonwood

56–119 .119

Pole non-cottonwood

15–29

Intermediate non-cottonwood Post-cottonwood

30–55

Cottonwood Cottonwood; some ash, elm, and boxelder Willow, Russian olive, and eastern red-cedar Willow, Russian olive, and eastern red-cedar Ash, elm, and boxelder

.55

Bird surveys

cottonwood forests are composed primarily of willow, Russian olive (Elaeagnus angustifolia), and eastern red-cedar. Forests, agricultural lands, residential development, anthropogenic woodlots (e.g., shelterbelts, windbreaks), and grasslands are interspersed within the floodplain of the study area. We sampled 75 stands, with two bird survey points per stand. Most locations were chosen from the 121 sites surveyed for vegetation along the MNRR in a study by Dixon et al. (2010). Survey points were placed at least 250 meters apart and 50 meters from a forest edge, when possible. We did not set a minimum distance between stands, but maintained the preceding rules for survey points for stands that were adjacent to one another. Survey points were located in forest patches stratified by a combination of stand type (cottonwood, non-cottonwood, post-cottonwood) and age (15–29, 30–55, 56–119, .119 years), relative to 2012 (Table 1). We determined stand ages through historical aerial photography and field reconnaissance (Dixon et al. 2012). In some cases, we consolidated these forest categories into fewer classes for analysis (e.g., young forest, ,30 years; mature, .55 years across all forest types). We did not sample forests less than 15 years old, because they do not provide the habitat characteristics required by the bird species of interest and had not been sampled prior to the flood. Of our 75 study sites, 18 out of 19 pole-aged (15–29 years) sites, 16 out of 18 intermediate-aged (30–55 years) sites, and 7 out of 38 mature/old (.55 years) sites experienced flooding.


Vegetation surveys We surveyed vegetation structure and composition within forest stands containing bird survey sites during pre-flood (2006–2009) and post-flood (2012–2013) time periods using survey methods from Dixon et al. (2010; http://digitalcommons. unl.edu/usarmyceomaha/78/). We sampled 73 sites pre-flood and re-sampled 52 of these sites post-flood. Each site was sampled for trees (diameter at breast height 10 cm) using either 15-m radius circular plots (12 per stand) or the point-centered quarter method (40 points per stand; Cottam and Curtis 1956). Shrub density and cover were sampled using the line strip method (Lindsey 1955), with 2 3 10 m belt transects, at 12 points per stand.

Data analysis We estimated bird densities and species richness, tested for differences in density or species richness among years and according to stand types, and evaluated relationships between changes in vegetation structure and bird densities. Density calculations were limited to bird 6


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species that met the following criteria: (1) were woodland or forest obligates, (2) had a minimum of 80 detections during the five survey years, (3) were known to breed in South Dakota and Nebraska, (4) did not exhibit extensive flocking behavior, and (5) were not detected primarily by flyovers. Density estimates were calculated for the 35 species that met these criteria. Statistical significance was set a priori at a 0.05 for all statistical tests. Because of hybridization in South Dakota (Tallman et al. 2002) and difficulty in differentiating between Eastern (Pipilo erythrophthalmus) and Spotted (Pipilo maculatus) towhees, these two species were grouped as ‘‘Rufoussided’’ Towhee for analysis. Density estimates.—We used Program Distance 6.2 (Thomas et al. 2010) to estimate densities for each of the focal bird species across the two study segments by applying distance sampling methods (Buckland et al. 2001) to correct for imperfect detection. We fit global detection functions using data pooled across all years and sites and estimated densities, with 95% confidence intervals, for each species for each year, pre- and postflood periods, and year/forest categories. We calculated ‘‘pre-flood’’ density estimates from the pooled density estimate of 2009 and 2010. For each species, we pooled distance data into sets of distance bins to counteract heaping at particular distance values, and excluded the farthest 10% of detections from analyses. Distance bins varied by species, determined by the number of detections that occurred within each distance interval. We fit models to the binned data using either of two key functions (halfnormal and hazard rate) and one of three series expansions (cosine, hermite polynomial, and simple polynomial), which represented detection probability as a function of distance from the survey point. We chose the model with the lowest Akaike’s information criterion (AIC) value to estimate density. To achieve a better fitting model to our data, we recalculated distance bins if the chi-squared goodness-of-fit test generated a p value of ,0.10. Bird density changes.—We determined interannual changes in bird density for each of the 35 focal bird species using the density estimates and 95% confidence intervals generated by Program Distance. We considered pairwise differences between years to be significant if neither v www.esajournals.org

confidence interval overlapped the mean of the other year. Next, we assessed pre- to post-flood changes in mean bird densities by species using repeated measures ANOVAs, with density of each focal species as the response variable and year (‘‘preflood,’’ 2012, 2013, 2014) or year and forest category as predictor variables (Table 1). We treated each species and its density estimate in a given year as a replicate to calculate a mean density value for all species across the avian community for each year of the study. This approach is essentially equivalent to describing the annual density for an ‘‘average species’’ in the community as a whole. In separate tests, we measured the effect of year on mean densities across all focal species, changes in mean densities of species within three nesting guilds (ground/ shrub, canopy, and cavity nesters; classified using Ehrlich et al. 1988), and changes in mean species densities within each of the seven forest categories. For significant ( p 0.05) ANOVA results, we then used Tukey-Kramer post-hoc tests to identify which treatment groups differed from one another in mean species densities. The Tukey-Kramer procedure adjusts for the number of comparisons to conserve the overall experimentwise Type I error rate of the ANOVA. The two ground nesting species, ‘‘Rufous-sided’’ Towhee and Ovenbird were included with the shrub-nesting guild for analysis. All tests were run in SAS Enterprise Guide 4.3 OnDemand for Academics (Littell et al. 1996). Species richness.—We estimated species richness for each of the seven forest categories and sampling years. This included all forest bird species, not just the 35 focal species for which we computed density estimates. To minimize counting species from non-forest habitats, only birds detected within 50 meters of the point were used in species richness estimates. We excluded species that do not breed in floodplain forests within the study area, such as migrants and waterfowl. We computed species richness using a rarefaction analysis in EstimateS 9 (Colwell 2013), based on individual-abundance data. Within each forest category, we used confidence intervals to compare species richness among sampling years. Avian reponses to vegetation changes.—We measured the association between absolute and 7


MUNES ET AL. Table 2. List of 35 focal species, nesting guild assignment, and density trends. Density trends: (1) lower in 2012 with densities increasing in 2013 and/ or 2014, but remaining lower than pre-flood levels in 2014, (2) lower in 2012 followed by a rebound in 2013 and/or 2014 to pre-flood levels or higher, (3) lower in 2012 and remaining lower throughout all post-flood, (4) increasing across sample years, and (5) no change or trend. Species

Nesting guild

Density trend

Mourning Dove Red-headed Woodpecker Red-bellied Woodpecker Downy Woodpecker Hairy Woodpecker Northern Flicker Eastern Wood-Pewee Willow Flycatcher Great Crested Flycatcher Eastern Kingbird Bell’s Vireo Warbling Vireo Red-eyed Vireo Blue Jay Black-capped Chickadee White-breasted Nuthatch House Wren Wood Thrush American Robin Gray Catbird Brown Thrasher Ovenbird Common Yellowthroat American Redstart Yellow Warbler ‘‘Rufous-sided’’ Towhee Chipping Sparrow Song Sparrow Northern Cardinal Rose-breasted Grosbeak Indigo Bunting Brown-headed Cowbird Orchard Oriole Baltimore Oriole American Goldfinch

Canopy Cavity Cavity Cavity Cavity Cavity Canopy Shrub Cavity Canopy Shrub Canopy Canopy Canopy Cavity Canopy Cavity Canopy Canopy Shrub Shrub Ground Canopy Canopy Shrub Ground Shrub Shrub Shrub Shrub Shrub Shrub Shrub Canopy Shrub

2 2 4 4 4 5 1 3 2 5 1 4 5 1 5 5 1 2 4 4 1 3 2 4 2 2 4 5 2 3 5 2 2 1 4

was calculated as the combined densities of shrub and tree stems per hectare. We tested the hypothesis that there is an association between changes in vegetation and relative change in bird density using each of the following vegetation variables: proportional change in stem density, proportional change in percent shrub cover, absolute change in stem density, and absolute change in shrub cover.

RESULTS During the five-year survey period, we detected 32,615 individual birds from 121 species. Each of the 35 focal species was detected at one or more sites during each of the five years of the surveys. We categorized 13 species as shrub nesters, 12 species as tree nesters, eight as cavity nesters, and two as ground nesters (Table 2).

Changes in avian density and species richness There were significant differences in density estimates between years for 34 of 35 species (Table 2). Among the patterns of variation in density change among sampling years, the most common included: (1) significantly lower densities in 2012 with densities increasing in 2013 and/ or 2014, but remaining lower than pre-flood levels in 2014 (6 species; Fig. 3A–C); (2) significantly lower densities in 2012 followed by a rebound in 2013 and/or 2014 to pre-flood levels or higher (10 species; Fig. 3D); (3) significantly lower densities in 2012 and remaining significantly lower throughout all post-flood years (3 species; Fig. 3E); and (4) significantly increasing densities across sampling years, a pattern found in nine species, including three woodpecker species, American Goldfinch (Spinus tristis), and American Robin (Turdus migratorius; Fig. 3F). Compared to pre-flood, 2012 densities were significantly lower for 19 (54%) species, higher for 10 (29%) species, and not significantly different for 6 (17%) species. Willow Flycatcher, Brown Thrasher (Toxostoma rufum), and Ovenbird showed the greatest declines from pre-flood densities to 2012, with 70%, 66%, and 57% reductions in density, respectively. Densities of 9 of 35 species decreased by at least 50%. The median change in density among the focal species was 20%. Species whose density was lower in 2012 relative to pre-flood were associ-

proportional changes in vegetation structure and proportional changes in density for each of the 35 focal bird species from pre-flood to 2012, using the Pearson’s product moment correlation. Proportional change in bird density was calculated as the difference between post-flood density (2012) and pre-flood density (2009–2010 average), divided by the pre-flood density. We calculated woody stem density and percent shrub cover for sites corresponding with bird points that were sampled for vegetation in both pre-and post-flood periods (n ¼ 52). Stem density v www.esajournals.org

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Fig. 3. Inter-annual changes in mean bird density (95% CI) for (A) Bell’s Vireo, (B) House Wren, (C) Baltimore Oriole, (D) Orchard Oriole, (E) Ovenbird, and (F) American Robin. Letters indicate significant differences between means ( p 0.05).

ated with both young and mature forests and occurred in all nesting guilds. Fourteen of the 19 species that had low densities in 2012 increased significantly from 2012 to 2013 and 10 species recovered to pre-flood levels or higher by 2014. Mean bird species density across the 35 focal species (bird community) was significantly lower in 2012 than pre-flood ( p ¼ 0.0180), but pre-flood density did not differ significantly from 2013 ( p ¼ 0.9724) and 2014 ( p ¼ 0.0879) densities (Fig. 4). Among nesting guilds, only the shrub nesting guild showed a significantly lower mean species density from pre-flood to 2012 ( p ¼ 0.0123). Ten of the 15 species (67%) in the shrub-nesting group had significantly lower densities, three v www.esajournals.org

Fig. 4. Inter-annual variation in mean densities (6 SE) of bird species (n ¼ 35). Letters indicate significant differences between means ( p 0.05).

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MUNES ET AL. Table 3. Inter-annual changes in mean density (individuals per hectare) by forest category (with age in years in parentheses), with standard error. CW ¼ cottonwood, NCW ¼ non-cottonwood, PCW ¼ post-cottonwood. Letters indicate significantly different means among years ( p 0.05). Forest category CW(15–29) CW(30–55) CW(56–119) CW(.119) NCW(15–29) NCW(30–55) PCW(.55)

Pre-flood

2012

2013

2014

0.69A 6 0.17 0.78A 6 0.20 0.69AB 6 0.14 0.71AB 6 0.14 0.63A 6 0.15 0.83AB 6 0.20 0.60AB 6 0.12

0.48B 6 0.18 0.61A 6 0.20 0.61A 6 0.14 0.60A 6 0.10 0.54A 6 0.13 0.62B 6 0.15 0.49B 6 0.09

0.56A 6 0.13 0.78A 6 0.16 0.78AB 6 0.14 0.79B 6 0.14 0.61A 6 0.15 0.69B 6 0.20 0.60AB 6 0.11

0.72B 6 0.15 0.96B 6 0.21 0.86B 6 0.15 0.90C 6 0.15 0.70A 6 0.17 0.93C 6 0.24 0.70A 6 0.12

(20%) had significantly higher densities, and two (13%) showed no significant change. Three of eight (38%) cavity nesters had significantly lower densities, four (50%) had significantly higher densities, and one (13%) showed no significant change. Six of 12 (50%) canopy nesters were significantly lower, three (25%) were significantly higher, and three (25%) showed no significant change. Mean species densities of the shrubnesting guild did not differ significantly, however, from pre-flood to 2013 or 2014 ( p ¼ 0.7875 and p ¼ 0.7313, respectively). Mean densities of cavity and canopy nesting species did not differ significantly between pre-flood and any of the post-flood years.

In 2012, mean densities of focal bird species densities in pole-aged cottonwood stands were significantly lower ( p ¼ 0.0044) than pre-flood levels. Densities rebounded in 2013, and were not significantly different from pre-flood levels ( p ¼ 0.1595). In 2014, mean species densities were higher in all forest categories relative to pre-flood, although the difference was only significant for intermediate ( p ¼ 0.0405) and old cottonwood ( p ¼ 0.0264; Table 3). American Redstart (Setophaga ruticilla), Eastern Kingbird (Tyrannus tyrannus), Indigo Bunting (Passerina cyanea), and Orchard Oriole (Icterus spurius) had higher densities in forests .55 years old in 2012, 2013, and/or 2014 relative to pre-flood (e.g., Fig. 5).

Fig. 5. Inter-annual changes in bird density (95% CI) among forest strata from pre-flood to 2012, 2013, and 2014 for American Redstart. CW ¼ cottonwood, NCW ¼ non-cottonwood, PCW ¼ post-cottonwood.


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MUNES ET AL. Table 4. Inter-annual changes in mean species richness by forest category (with age in years in parentheses). CW ¼ cottonwood, NCW ¼ non-cottonwood, PCW ¼ post-cottonwood. There were no significant differences at p 0.05. Confidence intervals (95% CI) are in parentheses. Forest category CW(15–29) CW(30–55) CW(56–119) CW(.119) NCW(15–29) NCW(30–55) PCW(.55)

Pre-flood

2012

2013

2014

31.7 (26.2–37.1) 42.8 (38.5–47.2) 39.0 (32.8–45.2) 42.5 (37.6–47.5) 25.6 (21.0–30.1) 28.8 (23.7–33.8) 40.6 (36.2–45.1)

29.9 (25.3–34.4) 42.3 (37.3–47.3) 38.6 (35.7–41.4) 43.0 (38.0–48.0) 24.4 (20.9–27.9) 27.1 (21.6–32.6) 42.9 (40.2–45.7)

28.0 (23.7–32.2) 39.3 (34.0–44.5) 40.1 (38.1–42.2) 40.2 (35.0–45.5) 27.6 (21.3–34.0) 24.5 (19.9–29.0) 38.7 (36.2–41.2)

29.0 (25.2–32.8) 40.6 (38.4–42.8) 42.0 (34.9–49.0) 40.9 (37.5–44.4) 26.8 (22.9–30.6) 26.3 (22.2–30.3) 43.0 (39.2–46.7)

Among all sampling years and forest categories, species richness estimates varied between 24.4 (non-cottonwood pole, 2012) and 43.0 (postcottonwood, 2014). Species richness was higher in cottonwood and post-cottonwood than noncottonwood forests. Averaged across all sampling years, species richness was highest in old cottonwood forests. There were no significant inter-annual differences between species richness estimates within each forest category (Table 4).

headed Cowbird (Molothrus ater), and ‘‘Rufoussided’’ Towhee (Table 5). Significant negative correlations between change in abundance and magnitude of vegetation change were detected for Blue Jay (Cyanocitta cristata) and Warbling Vireo, but positive correlations (i.e., bird changes were in the same direction as vegetation changes) occurred for the six other species. All four vegetation variables were positively correlated with changes in ‘‘Rufous-sided’’ Towhee density.

Vegetation changes

DISCUSSION

Live woody stem densities declined significantly from pre-flood to 2012 in pole-aged and intermediate forests, by 78% and 49%, respectively (Fig. 6A). Similarly, percent shrub cover declined by 74% in pole-aged forest and by 59% in intermediate forest (Fig. 6B). There were no significant changes in woody stem density or shrub cover in mature (.55 years old) forests. Pre-to-post flood changes in absolute or relative stem density and/or percent shrub cover were significantly correlated with relative (proportional) change in bird density for 8 out of 35 species including Bell’s Vireo, Gray Catbird, Brown-

Our study suggests that floodplain forest bird communities along the Missouri River are resilient to the short-term effects of a large, infrequent flood disturbance, such as the flood of 2011. Despite significant initial declines, abundances of most species rebounded within two years after the flood. Post-flood population trends in our study were similar to two other studies that investigated the effects of flooding on landbirds in the United States. Knutson and Klaas (1997) found a significant decline in bird abundance and richness during (1993) and one year after

Table 5. Pearson’s Correlation Coefficient between proportional changes in bird density and change in vegetation. Only statistically significant ( p 0.05) results are shown. *p 0.05, **p , 0.01, ***p , 0.001, ****p , 0.0001. Stem density Species

Relative

Mourning Dove Bell’s Vireo Warbling Vireo Blue Jay House Wren Gray Catbird ‘‘Rufous-sided’’ Towhee Brown-headed Cowbird

Percent shrub cover Absolute

Relative

Absolute

0.28* 0.71**** 0.36* 0.29* 0.28* 0.35* 0.62**** 0.39**


0.49** 0.34*

11

0.59**** 0.50***

0.45**


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Fig. 6. Changes in stem density (A) and shrub cover (B) (6 SE) among three forest age classes from pre-flood to post flood at 52 sites with pre- and post-flood vegetation sampling.

(1994) a major flood on the Mississippi River, relative to pre-flood levels (1992). On the South Platte River in eastern Colorado, Knopf and Sedgwick (1987) found lower densities of Brown Thrasher and ‘‘Rufous-sided’’ Towhee one year post-flood, and a subsequent rebound for Brown Thrasher in year two. In contrast to our study, Knopf and Sedgwick found no declines of canopy species, such as ‘‘Northern’’ (Icterus galbula) and Orchard orioles, during either postflood year. v www.esajournals.org

Examples of initial declines followed by rapid recolonization (,5 years) of disturbed sites by birds have been documented for other large disturbance events, including fire (Bontrager et al. 1995, Woinarski and Recher 1997, Barlow and Peres 2004) and hurricanes (Waide 1991), although effects were measured at varying scales and for bird species with different dispersal capabilities. Waide (1991), Bontrager et al. (1995), and Barlow and Peres (2004) suggest that birds are displaced into temporary refugia by 12


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disturbance events, and may re-colonize the area after a period of initial declines. We saw some evidence for this in our study, with some species that normally favor early successional habitats exhibiting short-term increases in abundance in older, less impacted stands. Askins and Ewert (1991) and Waide (1991) found that feeding guild affected the magnitude of density change after Hurricane Hugo. Similarly, Torres and Leberg (1996) and Woinarski and Recher (1997) suggest that niche breadth and habitat specificity will differentiate species responses to disturbance. In our study, species nesting in shrubs or on the ground and those that primarily use early successional forested habitats appeared to have been impacted the most by the 2011 flood. Local vegetation structure and composition are thought to be important influences on bird abundance (Miller et al. 2004), as even species that are widely distributed exhibit clustering of abundance within sites of desirable resources (Maurer 1999). The 2011 flood removed much of the shrub layer in nearly every young forest stand and inundated the understory for a prolonged period of time during the breeding season in seven of the 40 mature stands. Our study provides some evidence that density declines in 2012 were influenced by changes in stem density and shrub cover, although significant effects only occurred for approximately onefourth of the bird species analyzed. The significant declines across the shrub-nesting guild and two ground-nesting species partially supports the hypothesis that flood effects to the forest understory would result in decreases in these bird species. Post-flood changes in vegetation structure only partially support observed changes in bird density. Vegetation structure remained similar from 2012 to 2013 and 2014 (i.e., sparsely vegetated pole-aged sites, open understory at intermediate-aged sites, etc.; C. Boever, unpublished data), but bird density rebounded to preflood levels or higher, even in the shrub nesting guild. Many canopy and cavity-nesting species found predominantly in mature forests had significantly lower densities in 2012, despite little change in the vegetation in those stands from flooding. In our study, only Baltimore Orioles (Icterus galbula) showed an aversion to mature forest sites that were flooded the previous year v www.esajournals.org

(Munes 2014). Potential explanations for this response are unclear. We did not detect changes in species richness of forest birds in our seven forest categories. Refugia at unflooded sites may have buffered the effects of species richness decreases at flooded sites. While not significant, species richness in pole-aged forests was lower in 2012 than each of the other sampling years, which would be expected due to the extensive vegetation removal across this habitat type. We detected more nonforest bird species, such as Piping Plovers and Least Terns, at our sites during the post-flood sampling period than before the flood, and observed habitat shifts in forest bird occupancy. Young forest bird species, such as American Redstart, American Goldfinch, and Orchard Oriole were found in greater numbers in old cottonwood forests after the flood, a trend presumably driven by changes in resource availability in early successional stands. Shortterm changes in species composition related to disturbance-generated habitat heterogeneity have been reported in studies of windthrow ˙ (Lain et al. 2008, Zmihorski and Durska 2011) and fire disturbance (Choi et al. 2014). Our findings suggest that there were likely multiple drivers influencing bird density and distribution during and after the flood. Depressed food availability and fewer suitable nesting sites in 2011 may have posed an ‘‘ecological crunch’’ for species, resulting in fewer resources, higher intraspecific competition, and lower bird abundances (Wiens 1977). Cool and wet weather, which predominated in 2011, can depress food availability (i.e., insect emergence) and nesting success (Finch 1991, Custer et al. 1996, Flaspohler 1996, Knutson and Klaas 1997). During floods in 1995 and 1996 on the lower Missouri River, Galat et al. (1998) found that emergent insect production was greatly reduced. Similarly, riparian corridors along the Missouri River and nearby woodlots showed an approximately 50% reduction in arthropod biomass in the spring of 2011, relative to 2010 and 2012 (M. Liu and D. L. Swanson, unpublished data). Like flooding, droughts may also affect resource availability (Cody 1981, Smith 1982), although multi-year events are more likely to result in changes to bird abundances (Albright et al. 2010). The entire duration of the 2012 breeding 13


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season was much warmer and drier than average, with extreme drought conditions developing through the month of June (NOAA 2012). Gray (1993) demonstrated that decreases in arthropods and insectivorous birds (flycatchers and gleaner species) were highly correlated during a drought year in Kansas, USA. In our study, density estimates showed significant mean decreases in the same species groupings in 2012 relative to pre-flood (Flycatchers: Great Crested Flycatcher [Myiarchus crinitus], Eastern Kingbird, Eastern Wood-Pewee, Willow Flycatcher, and four gleaners: Black-capped Chickadee (Poecile atricapillus), Common Yellowthroat [Geothlypis trichas], House Wren [Troglodytes aedon], and Bell’s Vireo). Drought has been linked to fewer active nests (Lindsey et al. 1997) as well as reduced nesting success (Nice 1957), leading to a reduction of recruitment in the following year (DeSante and Geupel 1987). The significant increases of the flycatcher group, as well as 14 focal species, from 2012 to 2013, do not reflect drought impact on recruitment. Finally, changes in predator/prey dynamics during the flood could have contributed to bird density declines in 2012, as birds and their potential predators may have been concentrated on sites unaffected by flooding (i.e., most forest stands .55 years) in 2011. Potential predators of birds that are present in the area include snakes, skunks (Kostecke et al. 2001), red fox, raccoons (Schmidt 2003), coyotes, rodents, corvids (including Blue Jay and American Crow [Corvus brachyrhynchos]; Haskell 1995), and Brown-headed Cowbird, an obligate brood parasite (Shaffer et al. 2003). Increased contact between predator and prey species was observed during a large flood on the Mississippi River in 1993, when animals were forced onto higher ground by rising waters (Allen 1993). Such shifts in predator territories may predominantly impact ground and shrub nesters, as nest height can be a predictor of nest success (Gentry et al. 2006) and cavity nesters are inherently less vulnerable to predation (Martin and Li 1992). We did not detect an increase in Blue Jays, American Crows, or Brown-headed Cowbirds on sites greater than 55 years old, suggesting it is unlikely that bird predators or brood parasitism contributed to a decline in overall bird density. v www.esajournals.org

Managing disturbances Our data suggest that to effectively maintain bird populations along the Missouri River, shortterm management actions following disturbance, such as the flood of 2011, are probably not necessary. Proactive long-term management actions are, however, likely required. Declines of ‘‘disturbance-dependent’’ bird species have been documented by Askins (2000), Brawn et al. (2001), and Hunter et al. (2001). Disturbance management poses challenges to reconciling the conflicting objectives of social welfare and ecosystem function (NRC 1992, Dale et al. 1998), but the emulation of natural disturbance agents is necessary to mitigate habitat losses in floodplains, shrublands, grasslands, savannahs, and boreal forests (Brawn et al. 2001). Along the Missouri River, alterations to disturbance regimes and sediment transport from flow regulation has limited recruitment of pioneer woody species such as cottonwood and willow during the last few decades (Johnson et al. 1976, Johnson 1992, Scott et al. 2002, Dixon et al. 2012). These chronic effects of flow regulation may also reduce landscape diversity (Johnson 1992) and limit creation of early successional habitats used by many riparian bird species such as Yellow Warbler (Setophaga petechial ), Bell’s Vireo, and Orchard Oriole, some of which have shown regional declines (Sauer et al. 2014). Yet, large disturbance events are still likely to occur episodically, even on highly managed systems, with expectations for greater frequencies and intensities in the future under climate change (Milly et al. 2002, Palmer et al. 2008). Despite the inevitability of disturbance events, they are seldom incorporated into ecosystem management plans (Bayley 1995). Dale et al. (1998) recommends managing post-disturbance succession by aiding, removing, or augmenting particular ecological components or processes. For birds, they suggest the addition of nest boxes and perches to supplement lost habitat. In the short-term, bird species in our study showed rapid recovery despite flood-related changes in habitat, suggesting that such direct interventions would be unnecessary, although longer term monitoring and demographic studies may be needed to verify that population recovery has occurred. Changes in abundance do not necessarily reflect functional changes in habitat suit14


MUNES ET AL. 43(11):732–737. Andersen, D. C., D. J. Cooper, and K. Northcott. 2007. Dams, floodplain land use, and riparian forest conservation in the semiarid upper Colorado River basin, USA. Environmental Management 40:453– 475. Askins, R. A. 2000. Restoring North America’s birds. Yale University Press, New Haven, Connecticut, USA. Askins, R. A. and D. N. Ewert. 1991. Impact of hurricane Hugo on bird populations on St. John, US Virgin Islands. Biotropica 23(4):481–487. Barlow, J. and C. A. Peres. 2004. Ecological responses to El Ni˜no–induced surface fires in central Brazilian Amazonia: management implications for flammable tropical forests. Philosophical Transactions of the Royal Society of London B 359(1443):367–380. Bayley, P. B. 1995. Understanding large river: floodplain ecosystems. BioScience 45(3):153–158. Benson, A. 2011. Effects of forest type and age class on songbird populations across a cottonwood successional gradient along the Missouri River. Thesis. University of South Dakota, Vermillion, South Dakota, USA. Best, L. B., K. E. Freemark, B. S. Steiner, and T. M. Bergin. 1996. Life history and status classifications of birds breeding in Iowa. Journal of the Iowa Academy of Science 103:34–45. Bontrager, D. R., R. A. Erickson, and R. A. Hamilton. 1995. Impacts of the October 1993 Laguna Canyon fire on California gnatcatchers and cactus wrens. Pages 69–76 in J. E. Keely and T. Scott, editors. Brushfires in California: ecology and resource management. International Association of Wildland Fire, Fairfield, Washington, USA. Brawn, J. D., S. K. Robinson, and F. R. Thompson. 2001. The role of disturbance in the ecology and conservation of birds. 1. Annual Review of Ecology and Systematics 32(1):251–276. Buckland, S. T., D. R. Anderson, K. P. Burnham, J. L. Laake, D. L. Borchers, and L. Thomas. 2001. Introduction to distance sampling. Oxford University Press, Oxford, UK. Canterbury, G. E., T. E. Martin, D. R. Petit, L. J. Petit, and D. E. Bradford. 2000. Bird communities and habitat as ecological indicators of forest condition in regional monitoring. Conservation Biology 14(2):544–558. Carlisle, J. D., S. K. Skagen, B. E. Kus, C. V. Riper, K. L. Paxtons, and J. F. Kelly. 2009. Landbird migration in the American West: recent progress and future research directions. Condor 111(2):211–225. Carter, M. F., and K. Barker. 1993. An interactive database for setting conservation priorities for western neotropical migrants. Pages 120–144 in D. M. Finch, P. Stangel, and W. Peter, editors. Status and management of neotropical migratory birds.

ability for nesting birds, so resiliency in abundance may or may not be accompanied by resiliency in productivity. Future research is needed to address the relationship between abundance and nesting success across different successional habitats in our study area, as well as in response to infrequent, large disturbances in these habitats. To ensure the long-term persistence of floodplain forest bird communities on the regulated Missouri River, however, management actions should focus on restoration of early successional riparian forests by facilitating vegetation recruitment through emulating natural flow and sediment regimes, through planting when and where process-based restoration is not feasible, by retaining areas of natural recruitment that occur after periodic floods, and by exploring the restoration potential of novel habitats such as reservoir deltas (Johnson 2002, Johnson et al. 2015, Volke et al. 2015).

ACKNOWLEDGMENTS Funding was provided via contracts #W912HZ-12-20009 (2012-2015) and #W912DQ-07-C-0011 (2007-2010) from the US Army Corps of Engineers, Wildlife Diversity Small Grants Program from the South Dakota Game Fish and Parks, and the Plains and Prairie Potholes Landscape Conservation Cooperative from the United States Fish and Wildlife Service. Additional funding was provided by the Missouri River Institute and the National Park Service. Thanks to the graduate students and field technicians from the University of South Dakota who collected bird and vegetation data, Chris Boever for summarizing vegetation data, and Malia Volke for producing the study area map. Fig. 2A was modified from a figure by Wayne Nelson-Stastny (USFWS, Yankton, South Dakota, USA). This research would not be possible without the gracious cooperation of private landowners, the Yankton Sioux Tribe, Karl Mundt National Wildlife Refuge, and National Park Service (Missouri National Recreational River).

LITERATURE CITED Albright, T. P., A. M. Pidgeon, C. D. Rittenhouse, M. K. Clayton, B. D. Wardlow, C. H. Flather, P. D. Culbert, and V. C. Radeloff. 2010. Combined effects of heat waves and droughts on avian communities across the conterminous United States. Ecosphere 1(5):art12. Allen, W. H. 1993. The great flood of 1993. BioScience


15


MUNES ET AL. September 21–25, 1992, Estes Park, Colorado. General Technical Report RM-229. Rocky Mountain Forest and Range Experiment Station, USDA Forest Service, Fort Collins, Colorado, USA. Choi, C. Y., E. J. Lee, H. Y. Nam, W. S. Lee, and J. H. Lim. 2014. Temporal changes in the breeding bird community caused by post-fire treatments after the Samcheok forest fire in Korea. Landscape and Ecological Engineering 10(1):1–12. Cody, M. L. 1981. Habitat selection in birds: the roles of vegetation structure, competitors, and productivity. BioScience 31(2):107–113. Colwell, R. K. 2013. EstimateS: Statistical estimation of species richness and shared species from samples. Version 9. User’s guide and application. http://purl. oclc.org/estimates Cottam, G. and J. T. Curtis. 1956. The use of distance measures in phytosociological sampling. Ecology 37(3):451–460. Custer, T. W., R. K. Hines, and C. M. Custer. 1996. Nest initiation and clutch size of Great Blue Herons on the Mississippi River in relation to the 1993 flood. Condor 98(2):181–188. Dale, V. H., A. E. Lugo, J. A. MacMahon, and S. T. A. Pickett. 1998. Ecosystem management in the context of large, infrequent disturbances. Ecosystems 1(6):546–557. DeSante, D. F. and G. R. Geupel. 1987. Landbird productivity in central coastal California: the relationship to annual rainfall, and a reproductive failure in 1986. Condor 89(3):636–653. Dixon, M. D., C. J. Boever, V. L. Danzeisen, C. L. Merkord, E. C. Munes, M. L. Scott, W. C. Johnson, and T. C. Cowman. 2015. Effects of a ‘‘natural’’ flood event on the riparian ecosystem of a regulated large-river system: the 2011 flood on the Missouri River, USA. Ecohydrology 8(5):812– 824. http://dx.doi.org/10.1002/eco.1613 Dixon, M. D., W. C. Johnson, M. L. Scott, and D. Bowen. 2010. Status and trend of cottonwood forests along the Missouri River. Report no. 78. United States Army Corps of Engineers, Omaha District, Omaha, Nebraska, USA. Dixon, M. D., W. C. Johnson, M. L. Scott, D. Bowen, and L. A. Rabbe. 2012. Dynamics of plains cottonwood (Populus deltoides) forests and historical landscape change along unchannelized segments of the Missouri River, USA. Environmental Management 49(5):990–1008. Ehrlich, P. R., D. S. Dobkin, and D. Wheye. 1988. The birder’s handbook: a field guide to the natural history of North American birds. Simon and Schuster, New York, New York, USA. Finch, D. M. 1991. House Wrens adjust laying dates and clutch size in relation to annual flooding. Wilson Bulletin 103(1):25–43. Finch, D. M. and L. F. Ruggiero. 1993. Wildlife habitats


and biological diversity in the Rocky Mountains and northern Great Plains. Natural Areas Journal 13(3):191–203. Flaspohler, D. J. 1996. Nesting success of the Prothonotary Warbler in the Upper Mississippi River bottomlands. Wilson Bulletin 103(3):457–466. Friedman, J. M., W. R. Osterkamp, M. L. Scott, and G. T. Auble. 1998. Downstream effects of dams on channel geometry and bottomland vegetation: regional patterns in the Great Plains. Wetlands 18(4):619–633. Galat, D. L. et al. 1998. Flooding to restore connectivity of regulated, large-river wetlands. BioScience 48(9):721–733. Galat, D. L., C. R. Berry, E. J. Peters, and R. G. White. 2005. Missouri River basin. Pages 427–480 in A. C. Benke and C. E. Cushins, editors. Rivers of North America. Elsevier, Oxford, UK. Gentry, D. J., D. L. Swanson, and J. D. Carlisle. 2006. Species richness and nesting success of migrant forest birds in natural river corridors and anthropogenic woodlands in southeastern South Dakota. Condor 108(1):140–153. Gray, L. J. 1993. Response of insectivorous birds to emerging aquatic insects in riparian habitats of a tallgrass prairie stream. American Midland Naturalist 129(2):288–300. Greene, S. L., and J. C. Knox. 2014. Coupling legacy geomorphic surface facies to riparian vegetation: Assessing red cedar invasion along the Missouri River downstream of Gavins Point dam, South Dakota. Geomorphology 204(1):277–286. Haskell, D. G. 1995. A reevaluation of the effects of forest fragmentation on rates of bird-nest predation. Conservation Biology 9(5):1316–1318. Hesse, L. W., C. W. Wolfe, and N. K. Cole. 1988. Some aspects of energy flow in the Missouri River ecosystem and a rationale for recovery. Pages 13– 29 in N. G. Benson, editor. The Missouri River, the resources, their uses, and values. Special Publication 8. American Fisheries Society, North Central Division, USA. Hunter, W. C., D. A. Buehler, R. A. Canterbury, J. L. Confer, and P. B. Hamel. 2001. Conservation of disturbance-dependent birds in eastern North America. Wildlife Society Bulletin 29(2):440–455. Hupp, C. R., and W. R. Osterkamp. 1996. Riparian vegetation and fluvial geomorphic processes. Geomorphology 14(4):277–295. Intergovernmental Panel on Climate Change [IPCC]. 2012. Managing the risks of extreme events and disasters to advance climate change adaptation. Special Report of the Intergovernmental Panel on Climate Change. Cambridge University Press, Cambridge, UK. Jacobson, R. B., C. M. Elliott, and B. L. Huhmann. 2010. Development of a channel classification to evaluate

16


MUNES ET AL. potential for cottonwood restoration, lower segments of the middle Missouri River, South Dakota and Nebraska. United States Geological Survey Scientific Investigations Report 2010–5208. United States Geological Survey, Reston, Virginia, USA. James, F. C. and N. O. Wamer. 1982. Relationships between temperate forest bird communities and vegetation structure. Ecology 63(1):159–171. Johnson, W. C. 1992. Dams and riparian forests: case study from the upper Missouri River. Rivers 3(4):229–242. Johnson, W. C. 2002. Riparian vegetation diversity along regulated rivers: contribution of novel and relict habitats. Freshwater Biology 47(4):749–759. Johnson, W. C., R. L. Burgess, and W. R. Keammerer. 1976. Forest overstory vegetation and environment on the Missouri River floodplain in North Dakota. Ecological Monographs 46(1):59–84. Johnson, W. C., M. D. Dixon, M. L. Scott, L. Rabbe, G. Larson, M. Volke, and B. Werner. 2012. Forty years of vegetation change on the Missouri River floodplain. BioScience 62(2):123–135. Johnson, W. C., M. A. Volke, M. L. Scott, and M. D. Dixon. 2015. The dammed Missouri: prospects for recovering Lewis and Clark’s river. Ecohydrology 8(5):765–771. http://dx.doi.org/10.1002/eco.1534 Jones, R. H., R. R. Sharitz, P. M. Dixon, D. S. Segal, and R. L. Schneider. 1994. Woody plant regeneration in four floodplain forests. Ecological Monographs 64(3):345–367. Junk, W. J., P. B. Bayley, and R. E. Sparks. 1989. The flood pulse concept in river-floodplain systems. Canadian Special Publication of Fisheries and Aquatic Sciences 106(1):110–127. Karr, J. and R. R. Roth. 1971. Vegetation structure and avian diversity in several New World areas. American Naturalist 105(945):423–435. Kingsford, R. T. 2000. Ecological impacts of dams, water diversions and river management on floodplain wetlands in Australia. Austral Ecology 25(2):109–127. Knopf, E. L., R. R. Johnson, T. Rich, E. B. Samson, and R. C. Szaro. 1988. Conservation of riparian ecosystems in the United States. Wilson Bulletin 100(2):272–284. Knopf, F. L. and J. A. Sedgwick. 1987. Latent population responses of summer birds to a catastrophic, climatological event. Condor 89(4):869–873. Knutson, M. G., J. P. Hoover, and E. E. Klaas. 1995. The importance of floodplain forests in the conservation and management of neotropical migratory birds in the Midwest. Pages 168–188 in E. R. Thompson, editor. Management of Midwestern landscapes for the conservation of Neo-tropical migratory birds. General Technical Report NC-187. USDA Forest Service, North Central Forest Experiment Station,


Detroit, Michigan, USA. Knutson, M. G., and E. E. Klaas. 1997. Declines in abundance and species richness of birds following a major flood on the Upper Mississippi River. Auk 114:367–380. Knutson, M. G., L. E. McColl, and S. A. Suarez. 2005. Breeding bird assemblages associated with stages of forest succession in large river floodplains. Natural Areas Journal 25(1):55–70. Kostecke, R. M., G. M. Linz, and W. J. Bleier. 2001. Survival of avian carcasses and photographic evidence of predators and scavengers. Journal of Field Ornithology 72(3):439–447. Lain, E. J., A. Haney, J. M. Burris, and J. Burton. 2008. Response of vegetation and birds to severe wind disturbance and salvage logging in a southern boreal forest. Forest Ecology and Management 256(5):863–871. Liknes, E. T., K. L. Dean, and D. L. Swanson. 1994. Avian diversity, density, and breeding status in a riparian community in southeastern South Dakota. Proceedings of the South Dakota Academy of Science 73:83–100. Lindsey, A. A. 1955. Testing the line-strip method against full tallies in diverse forest types. Ecology 36(3):485–494. Lindsey, G. D., T. K. Pratt, M. H. Reynolds, and J. D. Jacob. 1997. Response of six species of Hawaiian forest birds to a 1991-1992 El Ni˜no drought. Wilson Bulletin 109(2):339–343. Littell, R. C., G. A. Milliken, W. W. Stroup, and R. D. Wolfinger. 1996. SAS system for mixed models. SAS Institute, Cary, North Carolina, USA. MacArthur, R. H. and J. MacArthur. 1961. On birds species diversity. Ecology 42(3):594–598. MacArthur, R. H., J. W. MacArthur, and J. Preer. 1962. On bird species diversity. II. Prediction of bird censuses from habitat measurements. American Naturalist 96(888):167–174. Marston, R. A., J. D. Mills, D. R. Wrazien, B. Bassett, and D. K. Splinter. 2005. Effects of Jackson Lake Dam on the Snake River and its floodplain, Grand Teton National Park, Wyoming, USA. Geomorphology 71(1):79–98. Martin, T. E., and P. Li. 1992. Life history traits of openvs. cavity-nesting birds. Ecology 73(2):579–592. Maurer, B. A. 1999. Untangling ecological complexity: the macroscopic perspective. University of Chicago Press, Chicago, Illinois, USA. Miller, J. R., M. D. Dixon, and M. G. Turner. 2004. Response of avian communities in large-river floodplains to environmental variation at multiple scales. Ecological Applications 14(5):1394–1410. Milly, P. C. D., R. Wetherald, K. A. Dunne, and T. L. Delworth. 2002. Increasing risk of great floods in a changing climate. Nature 415(6871):514–517. Mumba, M., and L. Thompson. 2005. Hydrological and

17


MUNES ET AL. ecological impacts of dams on the Kafue Flats floodplain system, southern Zambia. Physics and Chemistry of the Earth 30(6):442–447. Munes, E. C. 2014. Short-term and projected long-term responses of floodplain bird communities 524 to the 2011 flood on the Missouri River. Thesis. University of South Dakota, Vermillion, South Dakota, USA. National Climatic Data Center [NOAA]. 2012. State of the climate: drought for June 2012. http://www. ncdc.noaa.gov/sotc/drought/2012/6 National Research Council Panel [NRC]. 1992. Restoration of aquatic ecosystems, science, technology, and public policy. National Academy Press, Washington, D.C., USA. Nice, M. M. 1957. Nesting success in altricial birds. Auk 74(3):305–321. Palmer, M. A., C. A. Reidy Liermann, C. Nilsson, M. Flo¨rke, J. Alcamo, P. S. Lake, and N. Bond. 2008. Climate change and the world’s river basins: anticipating management options. Frontiers in Ecology and the Environment 6(2):81–89. Parsons, M., C. A. McLoughlin, M. W. Rountree, and K. H. Rogers. 2006. The biotic and abiotic legacy of a large infrequent flood disturbance in the Sabie River, South Africa. River Research and Applications 22(2):187–201. Reich, P., and P. S. Lake. 2014. Extreme hydrological events and the ecological restoration of flowing waters. Freshwater Biology. http://dx.doi.org/10. 1111/fwb.12508 Rood, S. B. and J. M. Mahoney. 1990. Collapse of riparian poplar forests downstream from dams in western prairies: probable causes and prospects for mitigation. Environmental Management 14(4):451– 464. Rood, S. B., G. M. Samuelson, J. H. Braatne, C. R. Gourley, F. M. Hughes, and J. M. Mahoney. 2005. Managing river flows to restore floodplain forests. Frontiers in Ecology and the Environment 3(4):193– 201. Rumble, M. A. and J. E. Gobeille. 2004. Avian use of successional cottonwood (Populus deltoides) woodlands along the middle Missouri River. American Midland Naturalist 152(1):165–177. Sauer, J. R., W. A. Link, J. E. Fallon, K. L. Pardieck, and D. J. Ziolkowski. 2014. The North American breeding bird survey 1966–2013: summary analysis and species accounts. North American Fauna 79:1– 32. Schmidt, K. A. 2003. Nest predation and population declines in illinois songbirds: a case for mesopredator effects. Conservation Biology 17(4):1141– 1150. Scott, J. M., P. J. Heglund, M. L. Morrison, J. B. Haufler, M. G. Raphael, W. A. Wall, and F. B. Samson, editors. 2002. Predicting species occurrences: issues


of accuracy and scale. Island Press, Washington, D.C., USA. Scott, M. L., G. T. Auble, and J. M. Friedman. 1997. Flood dependency of cottonwood establishment along the Missouri River, Montana, USA. Ecological Applications 7(2):677–690. Shaffer, J. A., C. M. Goldade, M. F. Dinkins, D. H. Johnson, L. D. Igl, and B. R. Euliss. 2003. Brownheaded Cowbirds in grasslands: their habitats, hosts, and response to management. Prairie Naturalist 35(3):145–186. Smith, K. G. 1982. Drought-induced changes in avian community structure along a montane sere. Ecology 63(4):952–961. Swanson, D. L. 1999. Avifauna of an early successional habitat along the middle Missouri River. Prairie Naturalist 31(3):145–164. Tallman, D. T., D. L. Swanson, and J. S. Palmer. 2002. Birds of South Dakota. Third edition. South Dakota Ornithologists’ Union, Aberdeen, South Dakota, USA. Thogmartin, W. E., M. Gallagher, N. Young, J. J. Rohweder, F. Durbian, and M. G. Knutson. 2009. Avian assemblages in the lower Missouri River floodplain. Wetlands 29(2):552–562. Thomas, L., S. T. Buckland, E. A. Rexstad, J. L. Laake, S. Strindberg, S. L. Hedley, J. R. B. Bishop, T. A. Marques, and K. P. Burnham. 2010. Distance software: design and analysis of distance sampling surveys for estimating population size. Journal of Applied Ecology 47(1):5–14. Torres, A. R. and P. L. Leberg. 1996. Initial changes in habitat and abundance of cavity-nesting birds and the Northern Parula following hurricane Andrew. Condor 98(3):483–490. Turner, M. G. 2010. Disturbance and landscape dynamics in a changing world. Ecology 91(10):2833–2849. Turner, M. G., W. L. Baker, C. J. Peterson, and R. K. Peet. 1998. Factors influencing succession: lessons from large, infrequent natural disturbances. Ecosystems 1(6):511–523. United States Army Corps of Engineers [USACE]. 2012. Post 2011 flood event analysis of Missouri River mainstem flood control storage. United States Army Corps of Engineers, Omaha District, Omaha, Nebraska, USA. Volke, M. A., M. L. Scott, W. C. Johnson, and M. D. Dixon. 2015. The ecological significance of emerging deltas in regulated rivers. BioScience. 65(6 ):598–611. http://dx.doi.org/10.1093/biosci/ biv040 Waide, R. B. 1991. The effect of Hurricane Hugo on bird populations in the Luquillo experimental forest, Puerto Rico. Biotropica 23(4):475–480. Wiens, J. A. 1977. On competition and variable environments. American Scientist 65(5):590–597.

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MUNES ET AL. contrasting management types on two distinct taxonomic groups in a large-scaled windthrow. European Journal of Forest Research 130(4):589– 600.

Woinarski, J., and H. Recher. 1997. Impact and response: a review of the effects of fire on the Australian avifauna. Pacific Conservation Biology 3(3):183. ˙ Zmihorski, M., and E. Durska. 2011. The effect of

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