Southeastern Naturalist W.A. Kroschel, S.L. King, and R.F. Keim 2016 42 Vol. 15, Special Issue 9 Tree Regeneration by Seed in Bottomland Hardwood Forests: A Review Whitney A. Kroschel1,*, Sammy L. King2, and Richard F. Keim1 Abstract - Bottomland hardwood forests (BLH) are found in temperate, humid regions of the southeastern US, primarily on alluvial floodplains adjacent to rivers. Altered hydrology in rivers and floodplains has caused changes in stand development and species composition of BLHs. We hypothesize that the driving mechanisms behind these changes are related to the regeneration process because of the complexity of recruitment and the vulnerability of species at that age in development. Here we review the state of our understanding regarding BLH regeneration, and identify potential bottlenecks throughout the stages of seed production, seed dispersal, germination, establishment, and survival. Our process-level understanding of regeneration by seed in BLHs is rudimentary, thus limiting our ability to predict the effects of hydrologic alterations on species composition. By focusing future research on the appropriate stages of regeneration, we can better understand the sources of forest-community transitions across the diverse range of BLH systems. Introduction Bottomland hardwood forests (BLH) are found in temperate, humid regions of the southeastern US, primarily on alluvial floodplains adjacent to rivers. Historic BLHs were a product of the natural hydrologic and geomorphic processes associated with their adjacent rivers. Over time, repeated flooding and erosional and depositional events created a dynamic landscape that supported extensive floral and faunal diversity (Allen et al. 2001, Hodges 1997, Wharton et al. 1982). BLHs are extensive in the Lower Mississippi Alluvial Valley (LMAV), where about 10 million ha of BLHs originally existed (Hefner and Brown 1985, National Research Council 1982). With the development of agriculture within and around the LMAV, ~80% of BLH area was cleared for field crops (MacDonald et al. 1979, US Department of the Interior 1988). Similarly, East Texas lost over 60% of historic BLH ecosystems but still retains ~664,860 ha along major rivers and tributaries such as the Neches, Trinity, and Sabine rivers (Allen 1997, Elliott et al. 2014, Frye 1987). In addition to direct conversion, BLHs have been affected by anthropogenic alterations to the hydrology of southeastern rivers and floodplains. Since the late 1800s, modifications to rivers for flood control and navigation purposes have dramatically altered the natural hydrologic and geomorphic processes of most BLH floodplains (Allen et al. 2001, Biedenharn and Watson 1997, Dynesius and 1School of Renewable Natural Resources, Louisiana State University, Baton Rouge, LA 70803. 2US Geological Survey, Louisiana Cooperative Fish and Wildlife Research Unit, LSU School of Renewable Natural Resources, Baton Rouge, LA 70803. *Corresponding author - whitney.kroschel@gmail.com. Manuscript Editor: Jerry Cook Proceedings of the 6th Big Thicket Science Conference: Watersheds and Waterflow 2016 Southeastern Naturalist 15(Special Issue 9):42–60 Southeastern Naturalist 43 W.A. Kroschel, S.L. King, and R.F. Keim 2016 Vol. 15, Special Issue 9 Nilsson 1994, Hudson et al. 2008, Pinter et al. 2006, Stanturf et al. 2001, Tockner and Stanford 2002). Particularly on large rivers, e.g., the Mississippi and Arkansas rivers, many floodplains have been disconnected from their associated channels by levee establishment and river channelization. This disconnection between channels and floodplains has eliminated regular flooding events which are important for shaping and maintaining the structure and function of these ecosystems (Gore and Shields 1995, Messina and Connor 1998, Tockner and Stanford 2002, Wharton et al. 1982). Flooding acts as a disturbance agent, seed-dispersal vector, soil-moisture source, and generator of geomorphic features (Wharton et al. 1982). Alterations to the natural flooding regime cause changes in associated BLHs, including drier soils, denser understories, and establishment of less flood-tolerant tree species (Gee 2012, Hanberry et al. 2012). Understanding how modifications to BLH systems have affected tree-species composition in these habitats is particularly difficult because the vegetation is strongly controlled by hydrology and geomorphology, which are intricately connected. Moreover, the lag time between cause and effect complicate understanding: it may take decades for the impacts of major events on forest vegetation to manifest (Elderd 2003, Faust 2006, Gee 2012, Jacobson and Faust 2014, King et al. 1998, McCarthy and Evans 2000). Recent research and anecdotal evidence suggest many BLHs are shifting from more flood-tolerant, shade-intolerant communities (e.g., Quercus lyrata Walter [Overcup Oak]–Carya aquatica (Michx. F.) Nutt. [Water Hickory]) to less flood-tolerant, shade-tolerant communities (e.g., Celtis laevigata Willd. [Sugarberry]–Ulmus americana L. [American Elm]–Fraxinus pennsylvanica Marsh. [Green Ash]; Quercus phellos L. [Willow Oak]–Water Oak–American Elm) (Alldredge and Moore 2012, Darst and Light 2008, Gee et al. 2014, Hanberry et al. 2012). These patterns suggest that conditions under which historic BLH stands developed (Shelford 1954) have been altered to the extent that the communities no longer maintain their former species composition (Lockhart et al. 2010). The regeneration process is critical for maintaining species composition, and changes to it may be the ultimate cause for most transitions in species assemblages. Regeneration is defined as the replacement of a mature individual by an individual of the same species (Grubb 1977, Poorter 2007). This process involves multiple stages including seed production, seed dispersal, germination, seedling emergence, seedling establishment, and survival (Fig. 1). Sprouting is another form of regeneration that has proven useful in forest management, yet it is contingent upon the success of prior establishment by the individual producing the sprout (Stanturf and Meadows 1994). Regeneration depends upon internal and external controls, i.e., innate physiological and autecological mechanisms as well as abiotic, environmental factors such as climate, disturbance, and interactions with other organisms (Price et al. 2001). Studies on the importance of regeneration, specifically seedling establishment, while essential, tend to be tedious and time consuming, and the time needed to develop an understanding of the temporally variable processes exceeds that of traditional ecological studies (Clark et al. 1999a). Specifically in BLHs, there is a Southeastern Naturalist W.A. Kroschel, S.L. King, and R.F. Keim 2016 44 Vol. 15, Special Issue 9 general lack of understanding regarding the regeneration process because of the difficulty in studying a highly variable ecosystem characterized by frequent seasonal flooding that varies within and among years (Wharton et al. 1982). Much of what is known about BLH regeneration pertains to physiological and absolute (i.e., how long a seedling can withstand continuous flooding) flood tolerance of seedlings, and their survival from the establishment stage on through maturity (e.g., Battaglia and Sharitz 2006, Battaglia et al. 1999, Denslow and Battaglia 2002, Elderd 2003, Hosner 1958, Jones et al. 1989, McCarthy and Evans 2000, McDermott 1954). Fewer researchers have focused on the initial regeneration stages and their combined effects (Battaglia et al. 2000, Jones et al. 1994, King 1995, Sarneel et al. 2014, Sharitz and Lee 1985, Streng et al. 1989). Here we focus on regeneration in BLHs, where natural and anthropogenic disturbances have synergistically produced a variegated system of floodplains, each under its own suite of environmental influences that have caused a range of dynamic effects. In this study, we review and synthesize existing literature on BLH regeneration by seed to identify the mechanisms by which regeneration processes, as affected by abiotic and biotic variables, may alter BLH species composition. Figure 1. The stages of the regeneration process and major influential variables. Adapted from Price et al. (2001). Southeastern Naturalist 45 W.A. Kroschel, S.L. King, and R.F. Keim 2016 Vol. 15, Special Issue 9 Seed Production Relatively few studies have evaluated seed production in BLHs. Those that did confirmed findings from previous studies in other ecosystems, which showed that heavier-seeded species (e.g., oaks) produce fewer, larger seeds, and light-seeded species (e.g., Platanus occidentalis L. [American Sycamore]) produce relatively many small seeds (Battaglia et al. 2008, Harper 1977, Jones et al. 1994, Streng et al. 1989). Streng et al. (1989) collected seedfall data for 3 fruiting seasons and calculated a much higher between-year variation in heavy-seeded species compared to light-seeded species. Additional factors such as seed viability and/or herbivory can restrict recruitment, regardless of seed abundance. In a study of Taxodium distichum (L.) Rich. (Bald Cypress )–Nyssa aquatica L. (Water Tupelo) seed production, Sharitz and Lee (1985) determined that total seed production was adequate for successful regeneration, but that low seed-viability, as well as insect parasitism and frugivory, were limiting factors in successful species recruitment. Jones et al. (1994) noted substantial disturbance to the forest floor (>25%) from rooting by Sus scrofa L. (Feral Pig). Other research has demonstrated a strong preference of Feral Pigs for both hard and soft mast, as well as for the soft tissue of fresh herbs and seedlings (Wood and Roark 1980). The fruit of heavy-seeded species (e.g., oaks, tupelos, and hickories) are so favored by Feral Pigs that the presence of these animals in a BLH may influence future overstory composition (Siemann et al. 2009). The effect of seed herbivory from other species is less clear, but Odocoileus virginianus Zimmermann (White-tailed Deer ) are reported to browse on BLH seedlings (Castleberry et al. 1999, 2000). More investigation is needed to determine which frugivores, herbivores, and pests can influence certain species’ seed production and the extent of those effects. Seed Dispersal Upon successful production of viable seeds, mature seeds must be dispersed to suitable regeneration sites within the environment. Dispersal depends on adult-tree fecundity, presence of dispersal vectors (animal, wind, water), phenology (late summer, fall, or spring), and seed type (e.g., drupe, acorn, samara) (Boedeltje et al. 2004; Clark et al. 1998; 1999a, 1999b; Grubb 1977; Howe and Smallwood 1982; Matlack 1994; McEuen and Curran 2004; Russo et al. 2006). Several studies have addressed dispersal of various BLH species. In BLH ecosystems, animals, wind, gravity, and water are the major dispersal vectors for seeds (King and Allen 1996, McCarthy and Evans 2000, Reid et al. 2014, Schneider and Sharitz 1988, Sharitz and Lee 1985, Streng et al. 1989). Hard-mast species such as oaks and hickories are often desirable BLH species for wildlife (Stanturf et al. 2000), but they are poorly dispersed relative to wind-dispersed species such as Sycamore (Platanus occidentalis) or American Elm (Ulmus americana) (Battaglia et al. 2008, Cosgriff and Brown 2004, McCarthy and Evans 2000, Streng et al. 1989). Hydrochory, seed dispersal via water, is an important means of dispersal for floodplain species (Reid Southeastern Naturalist W.A. Kroschel, S.L. King, and R.F. Keim 2016 46 Vol. 15, Special Issue 9 et al. 2014, Schneider and Sharitz 1988) and is particularly important for bottomland oaks. For instance, viable acorns of Overcup Oak are capable of floating (McCarthy and Evans 2000). Floodwater currents promote hydrochory functions by rolling or pushing seeds along the floodplain floor (McCarthy and Evans 2000, Schneider and Sharitz 1988). Animals (e.g., Cyanocitta cristata L. [Blue Jay], Sciurus carolinensis Gmelin [Gray Squirrel]) are another important dispersal vector (McCoy et al. 2004), but their impact is difficult to measure (Darley-Hill and Johnson 1981), which would potentially underestimate the dispersal distance of some of the heavier-seeded species that have relatively poor dispersal otherwise (Battaglia et al. 2008, Streng et al. 1989). Dispersal vectors could influence BLH regeneration, even in the presence of a regular flooding regime (Battaglia et al. 1995). In a study of dispersal patterns in a 20-y-old abandoned agricultural field subject to natural seed rain in northeast Louisiana, Battaglia et al. (2008) determined that dispersal was not a limiting factor in the regeneration process for the light-seeded and wind-dispersed American Elm and Ulmus crassifolia Nutt. (Cedar Elm), or the bird-dispersed Sugarberry. In the same study, the heavy-seeded oak and hickory species were severely dispersal limited; propagules were absent from the site despite the presence of mature trees on the bordering levees. In an earlier study at the same site, Battaglia et al. (2002) documented that the 6 dominant tree species were all primarily dispersed by wind or birds, with the exception of Diospyros virginiana L. (Persimmon). It is unclear how the altered hydrology of many floodplains has affected the efficiency of the various dispersal vectors within BLH communities. With the reduction or elimination of flooding in various locations, we would expect dispersal via hydrochory to become obsolete, which in turn, would negatively influence dispersal of heavy-seeded species. Whether or not reduced flooding has influenced regeneration and subsequent species composition in BLHs remains unknown. The dynamics of animal-dispersed seeds with regard to hydrologic modifications is also unclear and requires further investigation. Germination Germination is primarily dependent on seed viability, soil moisture, soil temperature, ambient air temperature, and light availability (e.g., Evans and Etherington 1990, Holl 1999, McLaren and McDonald 2003, Sarneel et al. 2014). Within a forest ecosystem, microsites of varying quality create a mosaic of conditions in which different species may germinate depending of their tolerance to a specific location (Grubb 1977). The concept of “windows of opportunity” is important in considering germination success. These windows may be due to small disturbance events that create generally favorable germination conditions (Eriksson and Froborg 1996) or species-specific germination windows, i.e., phenological differences in emergence timing (Rathcke and Lacey 1985, Sarneel et al. 2014, Streng et al. 1989). Dynamic environmental conditions that occur within a single growing season (e.g., flood frequency) may regulate germination success (Toner and Keddy 1997). Southeastern Naturalist 47 W.A. Kroschel, S.L. King, and R.F. Keim 2016 Vol. 15, Special Issue 9 Seed viability of BLH species varies across the floodplain because of the influence of abiotic factors and natural intrinsic variation within and between species (Goodson et al. 2003). Seed viability can range from 70–80% in Sweetgum to 35– 45% in Willow Oak (Bonner and Gammage 1967). The degree to which flooding affects seed viability in BLHs is uncertain because there has been limited research on this topic. A few studies have examined germination success of select BLH species following various periods of complete inundation. Overcup Oak acorns can survive and even benefit from periods of prolonged flooding (Cosgriff and Brown 2004, Pierce and King 2007), whereas those of other oak species (Quercus michauxii Nutt. [Swamp Chestnut Oak] and Quercus texana Buckley [Nuttall Oak]) are negatively affected by increased flooding (Briscoe 1961, Pierce and King 2007). In a Baldcypress–Water Tupelo swamp, Schneider and Sharitz (1986) reported a high proportion of nonviable Baldcypress and Water Tupelo seeds, which they attributed to naturally low viability and persistent flooding. Interestingly, one study documented Sugarberry seed viability of up to 5 y in floodplain soils (Meadows et al. 2006). More research is needed to discern the effects of inundation (or lack thereof) on seed viability and germination potential of different BLH species. Other factors involved in the germination process, such as soil conditions, temperature, and light, have not been thoroughly evaluated in BLH systems. Battaglia et al. (2000) found that light positively affected emergence of Swamp Chestnut Oak and Sweetgum, except when Sweetgum was in full sunlight. Emergence was consistently lower in shade conditions (Battaglia et al. 2000). Substantial further research to better understand the mechanisms behind germination of BLH species is needed. Establishment Absolute tolerance of flood events by mature trees may not be as important as 1st-year germinants’ tolerance of flood events. In general, the species-specific time of germination influences the probability of successful seedling establishment. Seeds that germinate earlier in the season are more vulnerable to disturbance (e.g. flooding); but if a seedling emerges earlier in the season and survives periods of stress, it has the advantage of a longer growing season (Baskin and Baskin 1972, Gross 1984, Rathcke and Lacey 1985, Streng et al. 1989). Factors such as light availability and microtopography may also be regulators of germination success (Battaglia and Sharitz 2006, Battaglia et al. 2000, King and Allen 1996). BLH species are adapted to an extremely variable environment; thus, it is possible that the timing of regeneration events, such as germination, are less sensitive to seasonal microsite cues and are more strongly controlled by intrinsic phenological patterns which developed over long time-spans and are specific to each species (i.e., germination “windows of opportunity”). Species-specific relationships between germination conditions and germination timing need to be evaluated in more detail if we are to better define the window of opportunity for each sp ecies. Reproductive strategies for tree species in relatively less-hydrologically altered systems include high seed-set (e.g., elms) with high seed- (or seedling-) mortality, Southeastern Naturalist W.A. Kroschel, S.L. King, and R.F. Keim 2016 48 Vol. 15, Special Issue 9 and less frequent and abundant seed production with high seed- (or seedling-) survivorship (e.g., oaks) (Rathcke and Lacey 1985). The high natural diversity in BLHs indicates that both reproductive strategies have succeeded in the presence of regular flooding events. Research by Streng et al. (1989) showed that BLH species emerged at different points during the growing season, the dates of which were significantly correlated with seed mass. Their data illustrated the use of different seasonal windows of opportunity per species, as well as the trend for lighter-seeded species to emerge before heavier-seeded species (Streng et al. 1989). If species are strongly linked to a temporal schedule as Streng et al. (1989) hypothesized, then the establishment stage becomes crucial for determining the success of newly emerged seedlings. Establishment is here defined as the survival of an individual from emergence through the first 3 growing seasons (Cooper et al. 1999). Successful establishment may determine a particular tree’s long-term role in the community, in that individuals with high initial growth after establishment have a better chance of reaching the canopy (Dekker et al. 2009). Similar to germination, this process is dependent on microsite-quality variables such as soil moisture, soil temperature, ambient air temperature, light availability, soil-nutrient content, and species-specific growth rate (Cooper et al. 1999, Grubb 1977). Much of the current literature on tree regeneration has focused on the establishment stage due to the vulnerability and high mortality-rate of seedlings in the first few growing seasons (e.g., Boerner and Brinkman 1996, Gray and Spies 1997, Jones et al. 1994, McDermott 1954, Molofsky and Augspurger 1992, Sack 2004, Streng et al. 1989). Few studies on BLH seedlings have followed individuals long enough to measure successful establishment (Jones et al. 1994, Streng et al. 1989), however, data collected even within the first season of growth could be valuable in helping to clarify limitations in recruitment. McDermott (1954) tested the effects of extended soil saturation on 6 BLH species, including Alnus incana (L.) Moench subsp. rugosa (Du Roi) R.T. Clausen (Speckled Alder), Sycamore, Betula nigra L. (River Birch), American Elm, Ulmus alata Michx. (Winged Elm), and Acer rubrum L. (Red Maple). He found that all of them recovered either rapidly or moderately well after 32 d of continuous soil saturation. Hosner (1958) tested survival rates of 6 BLH species: Populus deltoides W. Bartram ex Marsh. (Eastern Cottonwood), Acer negundo L. (Boxelder), Salix nigra Marsh. (Black Willow), Green Ash, Sweetgum, and Acer saccharinum L. (Silver Maple). With the exception of Silver Maple, all species survived at least 8 d of submersion. Jones et al. (1989) also tested effects of flooding on seedlings of light-seeded species and found little or no reduced growth in waterlogged conditions. In all of the above cases, light-seeded and moderately light-seeded species were able to survive and/or recover after substantial flood stress, despite some signs of reduced growth. One caveat to note is that each of these studies evaluated seedlings with at least 1 fully expanded leaf or that were ≥7.6 cm tall (Hosner 1958, Jones et al. 1989, McDermott 1954). The effects from flooding may be different if tested on newly emerged seedlings. Concerning heavier-seeded species, Battaglia et al. (2000) examined mortality rates of newly emerged seedlings of Swamp Chestnut Oak and determined that Southeastern Naturalist 49 W.A. Kroschel, S.L. King, and R.F. Keim 2016 Vol. 15, Special Issue 9 mortality was affected primarily by water-table levels. Flood stress reduces the photosynthetic capacity of seedlings (Pezeshki and Anderson 1997); completely submerged seedlings tend to have lower survival rates than those that experience little or no flooding (Jones et al. 1989). When compared to light-seeded species in a natural BLH habitat, Streng et al. (1989) found that seedlings of heavier-seeded species had higher survivorship compared to lighter-seeded species. However, this effect was countered by the greater abundance of lighter-seeded species; the high number increased the chances of at least some of the light-seeded species surviving through the first growing season. The field-based studies of natural regeneration conducted by Streng et al. (1989) and Jones et al. (1994) produced different results regarding flood effects on seedling densities. Where the former study found that flooding significantly reduced the density of light-seeded species, the latter determined that flooding had no effect on the density of light-seeded species. In both studies, heavy-seeded species were less affected by flooding, and the probability of seedling survival of all species significantly increased after the first growing season and continued to do so during the first few years after germinat ion. Establishment is also dependent on seed type and its specific location on the floodplain. Regular disruption from flood events may reduce the abundance of light-seeded species, producing a regeneration success rate comparable to that of heavy-seeded species (Streng et al. 1989). However, if regular disturbance events (i.e., flooding) that tend to limit the regeneration of the light-seeded species are eliminated or significantly reduced, the trajectory of succession in forest ecosystems could change in favor of the light-seeded species (Hobbs and Huenneke 1992). Over time, the more abundant, light-seeded species could outcompete heavy-seeded species (Gee et al. 2014). Several studies have indicated an increased dominance by light-seeded species (Gee et al. 2014, Hanberry et al. 2012, Schneider and Sharitz 1986, Streng et al. 1989), and altered hydrologic processes may be at least partially responsible for this trend. Research suggests that light availability may also be a limiting factor on seedling establishment, though not to the same extent as flooding. Like the effects of flooding, the effects of light seem to vary by seed type. Light-seeded species emerge earlier in the growing season (mid-February–mid-April), a time when light availability is high due to the canopy not having fully expanded (Streng et al. 1989). Conversely, heavy-seeded species emerge later, but their survivorship depends primarily on endosperm, which can persist for up to 1 y (Grime and Jeffrey 1965, Sork 1987). This energy reserve in heavier seeds may explain their relatively high survivorship despite the shorter growing season within the first year, but light may become an important factor in subsequent establishment years. Battaglia et al. (2000) found significant interactions between light availability and water-table level, reflected primarily in reduced establishment and survival in low light and high water-table treatments. In contrast, other evidence did not find light to be a limiting factor through the establishment stage of regeneration, though its importance for survival increases beyond this stage (McCarthy and Evans 2000). Southeastern Naturalist W.A. Kroschel, S.L. King, and R.F. Keim 2016 50 Vol. 15, Special Issue 9 Survival The survival stage is the period from successful establishment (>3 y) and thereafter until mortality. In forest systems in general, as an individual tree ages, soil moisture, temperature, light availability, and growth rate remain important, but some variables such as water-table level, precipitation, and competition begin to play a larger role in the survival process (Grime 1977, Grubb 1977) which in turn, may influence trade-off strategies (e.g., Battaglia and Sharitz 2006, Beckage and Clark 2003, Oki et al. 2013, Sterck et al. 2006). Research focus on this life stage has become increasingly prominent with regard to climate change, drought risk, and the potential implications for stand management. Climate change predictions for North America include greater drought severity in the next several decades (Cook et al. 2015), including longer drought duration, higher heat-severity and greater droughtfrequency, which may cause increased tree-mortality rates and rapid die-off events (Allen et al. 2001, Dale et al. 2001, Iverson and Prasad 1998). In BLHs, as seedlings develop and transition into saplings and then mature individuals, the influence of flooding seems to decline as the importance of light increases (Battaglia and Sharitz 2006, Hall and Harcombe 2001, King and Allen 1996, McCarthy and Evans 2000, Oki et al. 2013). Theoretically, every species has an adapted tolerance to a segment of the light spectrum because physiological attributes for one extreme are usually incompatible with those of the other extreme (Vallardares and Niinemets 2008). Thus, for optimal performance, species are generally recognized as shade-tolerant or shade-intolerant, but not both (Hall and Harcombe 1998); however, ontogenetic-niche shifts may contribute to variability within species (Eriksson 2002, Gabler and Siemann 2012, Nakazawa 2015). In a southeast Texas floodplain, shade-tolerant species grew faster than shade-intolerant species in low levels of light, and shade-intolerant species grew faster than shade-tolerant species in higher light-intensity (Lin et al. 2004). These results were consistent across sites that were and were not subject to flooding, indicating that slight or moderate flooding did not limit survival. Severe flooding influenced sapling survival in that shade-tolerant saplings experienced higher mortality (Lin et al. 2004). Another eastern Texas study suggested that canopy gaps, in combination with climatic variability (i.e., drought and flood events), produced the most favorable conditions for species diversity, and flooding tended to have a similar effect on all species at the sapling age (Hall and Harcombe 2001). Battaglia and Sharitz (2006) found no difference among saplings of species with regards to distance to the water table. However, they included light level in their analysis and found that shade-tolerant species generally grew in drier areas, and less shade-tolerant species occurred in more open, wetter areas, thus suggesting a flood/shade tradeoff strategy. When flooding is also removed or reduced as a stressor, light availability may become more of a limiting factor for some BLH species. Following the construction of a ring levee around a BLH site that had formerly been exposed to flooding, Gee et al. (2014) documented a significant increase in the relatively shade-tolerant Sugarberry compared to the less shade-tolerant Overcup Oak. Removal of flooding can also increase stem densities (Hanberry et al. 2012), presumably also reducing light availability. Southeastern Naturalist 51 W.A. Kroschel, S.L. King, and R.F. Keim 2016 Vol. 15, Special Issue 9 Canopy gaps release less-shade–tolerant species, such as oaks (Allen et al. 2001), from overstory competition (King and Allen 1996, Oliver et al. 2005) and increase species diversity (Bergeron 2000, Nagel et al. 2006, Oki et al. 2013); both of these effects would influence the success of young trees. McCarthy and Evans (2000) found greater survivorship of saplings growing under higher-intensity light than those in shaded conditions. Similar to Battaglia et al. (2000), McCarthy and Evans (2000) found that the combined effects of shading and flooding reduced the survival rate of Overcup Oak saplings, and saplings in shallowly flooded areas had a better chance of survival than ones in deeply flooded areas. Thus, sites that are open and wet may provide a better opportunity for bottomland oaks to succeed, whereas drier sites that are shaded or open may allow other competitor species such as Sugarberry to perform better than oaks. King and Antrobus (2005) examined the effect of canopy gaps on BLH composition, and found that small-scale canopy gaps (i.e., gaps created from the loss of a single tree) may not create sufficiently large openings with adequate light to facilitate canopy replacement by shade-intolerant species. Without large-scale disturbances, such as timber harvests or storm damage, floodplain forests will likely be replaced by more-shade–tolerant plant communities (King and Antrobus 2005). Species Composition Evaluating the role of regeneration in BLH systems may yield important insights into the mechanisms behind compositional transitions. Studies such as Streng et al. (1989) and Jones et al. (1994) demonstrate both the complexity and sensitivity of the regeneration process and how easily recruitment can be influenced within the first few growing seasons. This critical period of development is fundamentally important for understanding the current state of BLH systems. BLH systems vary widely in the hydrologic and geomorphic characteristics that affect the associated vegetation. For instance, in west Tennessee, the effects of channelization have altered sediment deposition such that species composition has transitioned to a more disturbance-tolerant community (Oswalt and King 2005). In eastern Texas, soil moisture may be a limiting factor in controlling the western ranges of several BLH species, such that the presence or absence of certain species is more closely associated with the availability of groundwater than with climate or precipitation patterns (Shankman et al. 2012). Several studies have demonstrated a relationship between reduced flooding and a shift from hydric to more mesic communities. When Alldredge and Moore (2012) sampled at a downstream site following the construction of a major dam in eastern Texas, they found that typical bottomland-forest species were being replaced by species more characteristic of upland forests. Research also suggests that the sudden increase of Sugarberry in the understory since the mid-1900s in some BLHs in Louisiana and Arkansas is due to a rapid decline in flood frequency (Gee 2012, Gee et al. 2014). In southwestern Kentucky, both mature and young stands appear to be moving towards more-mesic species composition as a result of human-altered hydrology (Shear et al. 2006); King and Antrobus (2005) documented a similar pattern Southeastern Naturalist W.A. Kroschel, S.L. King, and R.F. Keim 2016 52 Vol. 15, Special Issue 9 in northeastern Arkansas. By tracing some of these compositional changes back to stages in the regeneration process, further research could reveal the fundamental components of regeneration success as well as regeneration limitations. Succession in BLHs may also contribute to changes in species composition, but it is not the principal factor causing the transitions toward less flood-tolerant communities in many hydrologically altered systems. Progression from an Overcup Oak–Water Hickory type to an ash–elm–Sugarberry type requires flood-associated depositional processes to raise the floodplain elevation high enough to support less flood-tolerant species (Hodges 1997, Lockhart 2010). In many areas with observed species-composition changes, there have been no excessive depositional forces present following hydrologic modifications (Alldredge and Moore 2012, Gee 2012, Gee et al. 2014, Shear et al. 2006). With little to no deposition, BLH stands on floodplain flats will move towards an Overcup Oak–Water Hickory type, which may endure for hundreds of years (Hodges 1997). If deposition processes are present, BLH stands can transition toward an elm–ash–Sugarberry type that may also persist for centuries; but eventually the forests progress toward an oak–hickory forest, which is estimated to take a minimum of 600 years to develop (Hodges 1997, Lockhart et al 2010, Shelford 1954). The rate at which hydric BLHs are transitioning to mesic stands, e.g., ash–elm–Sugarberry, exceeds that which is predicted for natural BLH succession (Hodges 1997, Shelford 1954). With the reduction of flooding and the associated reduction in deposition at many BLH sites, it is doubtful that species-composition changes at these sites are unrelated to human-altered floodplain hydrology. As with most ecological processes, multiple variables are responsible for the BLH species assemblages recognized today. Hence, flooding on its own is not entirely responsible for maintaining BLH communities in their natural conditions. Geomorphic processes, such as the types and rates of sedimentation, also influence regeneration processes and forest composition (Hupp and Osterkamp 1996, Oswalt and King 2005, Pierce and King 2007). Flood regimes affect site composition by influencing soil drainage, aeration, and soil redox potential, among other attributes (Kupfer et al. 2010). Drought is expected to increase in the future, and its effects may be intensified due to altered surface and subsurface hydrology, which affect soil moisture and, hence, seedling survival (Markesteijn and Poorter 2009). In areas such as East Texas, drought has become a more prominent event in BLH forests in recent years and may extend its effects eastward with continued climate change (Martinez-Vilalta et al. 2012, Pederson et al. 2012). Although these disturbances are largely outside the scope of this review, a process-based understanding of regeneration will facilitate our broader understanding of the effects of drought and sedimentation processes on future forest composition. Other past and present disturbances such as wind storms, parasitism/herbivory, and fire must be considered too. Light availability, precipitation, soil nutrients, soil texture, and canopy-gap size are also regulators in site colonization and survival of woody species (Robertson and Augspurger 1999). Shade-tolerant species tend to colonize BLH sites with reduced or eliminated flooding ( Hanberry et al. 2012; King and Antrobus 2001, 2005). Without select harvesting, thinning, or other Southeastern Naturalist 53 W.A. Kroschel, S.L. King, and R.F. Keim 2016 Vol. 15, Special Issue 9 regular canopy disturbance, shade-tolerant trees can thrive in high densities within the understory, shading out competitors that require full sunlight (e.g., oaks) (Hanberry et al. 2012, King and Antrobus 2005). Evidence suggests that flooding alone cannot always maintain shade-intolerant tree species; the presence of large gaps is also necessary to promote their regeneration (King and Antrobus 2001, 2005). In much of the range of BLHs, hurricanes cause extensive wind damage that produce large canopy-gaps, which can and shift succession to a different state (Battaglia et al. 1999). Large-scale disturbances such as channel migration can affect forest-edge composition and structure; migration rates on annual–decadal scales support lower tree-density, basal area, and richness in contrast to decadal–centennial migration rates which allow for greater density, basal area, and richness (Meitzen 2009). Collectively, all of the variables that contribute to the structure, composition, and function of BLH stands are important in time and space, but flooding is the primary disturbance that shapes developing BLH ecosystems. Conclusion Within the past century, BLHs have exhibited a wide range of changes in stand development and species composition as a result of altered hydrology in rivers and floodplains. Evaluating the role of regeneration in BLH systems may yield important insight into the mechanisms behind compositional transitions. Research to date has revealed important trends between abiotic and biotic processes that promote and inhibit successful regeneration of BLH species, but further investigation is needed. The influence of herbivory, frugivory, and pests on seed production is not fully understood for many species; more heavily targeted species could become less prevalent in the future canopy. Wind- and bird-dispersed species seem to have an advantage in reaching favorable regeneration sites compared to gravity and waterdispersed species, but how these vectors have been affected in floodplains with altered hydrology is not directly known. We also have yet to identify the mechanisms behind germination windows of opportunity among species and what controls the timing of seedling emergence. Although we have made progress in discovering promoters and inhibitors of seedling establishment, more long-term studies are necessary to detect successful seedling establishment and how it relates to germination timing, seed type, light availability, location on the floodplain, and canopy gaps. The complexity involved in recruitment and seedling vulnerability provide many open avenues for further research in this area. Acknowledgments We thank the Texas Parks and Wildlife Department and the Louisiana Department of Wildlife and Fisheries for their assistance in providing BLH site-visit opportunities; Dr. Jerry Cook, Dr. Loretta Battaglia, and an anonymous reviewer for providing constructive comments for our synthesis; and Jim Neal and Amie Treuer-Kuehn for their assistance in providing BLH data for East Texas. Southeastern Naturalist W.A. Kroschel, S.L. King, and R.F. Keim 2016 54 Vol. 15, Special Issue 9 Literature Cited Alldredge, B., and G. Moore. 2012. 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