Information

2.4.6: Data Dive- Beaver Impacts on Wetlands - Biology


Overview

An ecosystem engineer is any animal that creates, significantly modifies, maintains or destroys a habitat. Figure 2.4.6a below displays some of the results in this study:

Figure(PageIndex{a}): Average number of species observed in sample plots 1-2 years and 10-12 years after beavers were introduced. Graph by Rachel Schleiger (CC-BY-NC) modified from data in Law A, Graywood MJ, Jones KC, Ramsay P, and Willby NJ 2017.

Questions

  1. What is the independent (explanatory) variable and the dependent (response) variable?
  2. What question(s) are the authors trying to answer with this graph?
  3. What trend(s) can be observed in this graph between the 1-2 and 10-12 timetables? Support your answer by referring to appropriate patterns in the graph.
  4. Do you think like the authors are satisfied with the results in the graph? Why?
  5. How can the results of this graph to inform future reintroduction of beavers where wetland restoration is needed?
  6. What information/patterns is not clear from this graph?

Vegetative ecology of natural and constructed wetlands along the Leon River in Comanche County, Texas.

Abstract.--The vegetation found in two riverine wetlands, located along the Leon River in the West Cross Timbers, Comanche County, Texas, was compared for one year. A wetland constructed during 1999 and an adjacent natural reference wetland, established by beaver, were compared using relative coverage, density, frequency, and importance values. Wetland indicator status for plants and diversity were also assessed. Over four sampling periods, the constructed wetland underwent succession to become similar to the adjacent natural wetland. Common cat-tail (Typha latifolia), an obligate wetland species, dominated the inner core of the constructed wetland with marsh-elder (Iva annua) occupying the drier periphery. Ninety-five percent of the species sampled in the reference wetland were hydrophytes and plant succession followed a centrifugal model. The constructed wetland was dominated by marsh-elder and knot grass (Paspalum distichum) with only 67% of species being hydrophytes. Diversity was greater in the constructed wetland because it was in the arrival and establishment phase of wetland succession.

Wetlands are defined by the presence of hydrophytic vegetation, periodic inundation, and hydric soils (United States Army Corps of Engineers (USACOE) 1987 Tiner 1999 van der Valk 2006 Mitsch & Gosselink 2007 Keddy 2010). Texas' wetlands are among its most valuable natural resources. These lands provide many economic and ecological benefits, including flood control, improved water quality, harvestable products, and habitat for abundant fish, shellfish, and wildlife resources (Texas Parks and Wildlife Department (TPWD) 1997 van der Valk 2006 Mitsch & Gosselink 2007 King et al. 2009). Texas contains a diversity of wetland habitat types, both natural and anthropogenic. Although wetlands in Texas comprise less than 5% of the state's total land area (TPWD 1997), Texas is one of 19 states that has exhibited the most significant losses of wetland ecosystems (Tiner 1984) with a 56% loss during the preceding 200 years (TPWD 1997). Texas also has 40 to 60% of its freshwater plant communities at moderate to high risk of being eliminated (Heinz Center 2008). Few wetlands have been studied in the West Cross Timbers region of Texas with one investigation involving only constructed wetlands (Williams & Hudak 2005) and another examining a constructed wetland with comparisons to a beaver-created marsh (Brister 2005), so little is known about wetland plant communities and succession.

The West Cross Timbers area of Texas (Figure 1) lies south of the Red River, west of the Fort Worth Prairie, north of the Lampasas Cut Plain, and east of the Rolling Plains. It encompasses an area of 1,665,686 ha and is considered a mosaic of prairie, woodland, and savannah vegetation characterized by post oak (Quercus stellata) and blackjack oak (Quercus marilandica) as the dominant woody species (Dyksterhuis 1948 Diggs et al. 1999 Hoagland et al. 1999). The Red, Trinity, Brazos, and Colorado River basins all drain a portion of the West Cross Timbers area of Texas, with numerous tributaries scattered throughout the area. Each of these rivers and their associated tributaries contribute a significant amount of acreage to the Texas riparian wetland component. One such tributary is the Leon River.

The headwaters of the Leon begin in Eastland County, near the town of Eastland, and flow southeastward through Comanche, Hamilton, and Coryell Counties to be joined by the Lampasas River in Bell County and the San Gabriel River in Milam County to form the Little River, a major tributary of the Brazos River. Two reservoirs, Lake Leon and Proctor Lake, are located on the upper reaches of the Leon. The Leon River contributes water to a number of riparian wetland complexes as it flows into and out of Proctor Lake.

One complex, approximately 32 ha in size, is located below the dam at Proctor Lake. It consists of two wetland areas located between the Leon River channel and the outflow channel that was built by the USACOE when the lake was constructed. One of the wetland areas, a freshwater marsh, has been created by beaver activity, resulting in dams which trap water from the Leon River channel and from seepage through the earthen portion of the lake dam. The age of the marsh is unknown, but according to lake personnel, it is thought to be as old as the lake, which became operational in 1963. For purposes of this study, the marsh will be referred to as the reference wetland.

The second of the wetland areas, a constructed wetland built by USACOE, TPWD, and Ducks Unlimited to increase waterfowl habitat is located to the southeast (downriver) of the reference wetland. It was completed in 1999 and did not receive any significant hydrologic input until the Winter/Spring of 2001. The constructed wetland was not seeded, nor was any vegetation introduced into the wetland. The constructed wetland lies adjacent to the reference wetland and the two are separated by a berm in order to allow independent control of water levels. Both wetland areas would be considered inland, riverine, freshwater marshes (Mitsch & Gosselink 2007).

Riverine marshes often border riparian forests or occupy pockets within them, sometimes in abandoned oxbows and at the headwaters of rivers. In addition, beavers build wetlands by damming streams, creating small riverine marshes that attract wildlife, especially waterfowl (van der Valk 2006 Mitch & Gosselink 2007). Beavers often play significant roles in wetland development and maintenance. If their dams hold, this creates a stable hydroperiod. If their dams frequently wash out, they cause water level fluctuations, thus increasing plant diversity. Increased plant diversity usually causes increased waterfowl and mammal diversity in and around the wetland. Vegetation in freshwater marshes created by beavers is often characterized by monocots such as Typha, moisture-adapted grasses, and sedges (Mitsch & Gosselink 2007). The additional nutrients added to the water as a result of beaver and other mammals, waterfowl, and organic material from decaying vegetation, increase the biodiversity of the wetland by feeding plankton, invertebrates, crustaceans, amphibians, fish, and reptiles. The beaver help to support the entire ecosystem of organisms, thus becoming a keystone species (Primack 2010). Due to their dam construction regulating water level and stream flow, beavers can convert an otherwise unfavorable area into one that is habitable by wetland biota (Schmidly 2004).

Loss of wetland habitat is a serious environmental problem (Mitsch & Gosselink 2007 Heinz Center 2008). As wetlands continue to decrease, not only in Texas but throughout the world, it is becoming increasingly important that wetland ecosystems be better understood in order to facilitate quality mitigation and construction practices. Since few wetlands in the West Cross Timbers region of Texas have been examined, the ecological data for this investigation will serve as a baseline for future studies of wetlands in the region. The objective of this investigation is to assess changes in vegetational ecology of the reference and constructed wetlands at Proctor Lake Reservoir.

The ecological comparisons were conducted below the dam at Proctor Lake in the two previously mentioned wetland areas located in Comanche County, Texas (Figure 1). Both wetlands were sampled to examine vegetation over four time periods. Sampling dates for vegetation were chosen based on growing season data for the area. The mean length of the frost-free period for Comanche County, Texas is about 230 days. The frost-free period begins in mid-March and lasts until mid-November. Sampling was done from August-September 2001, March-April 2002, June-July 2002, and September-October 2002.

Soil samples from each quadrat where vegetation was sampled were taken with a 38 cm soil sampling probe and compared to a Munsell soil color chart for hydric soil color indicators. Based on this test, all soils sampled in the reference wetland were hydric and two-thirds of the samples from the constructed wetland were classified as hydric.

Based on visual observations of inundation, soil saturation, and watermarks on surrounding vegetation, hydrology between the two sites was similar except in the Fall of 2001, at the start of the investigation, when about one-third of the quadrats in the constructed wetland had no visual signs of inundation, soil saturation, or watermarks on surrounding vegetation. Thus, when vegetation was sampled for the first time in the investigation, hydrology in the constructed wetland was more variable than the hydrology of the reference wetland, but during the three following sample periods, hydrology was constant in quadrats for both wetlands.

The ecological comparison for vegetation was conducted by first establishing a baseline that ran parallel to the Leon River and included both wetland areas. The baseline was approximately 890 m and was intersected by 10 transects that traverse the wetland areas. The baseline was divided into equally spaced segments (107 m apart) in the natural wetland and in the constructed wetland (71 m apart). During each sampling period, transect starting point for each segment was determined using a random numbers table, and the transect was situated perpendicular to the baseline of the wetland. Each transect had two randomly-stratified observation points determined by using a random numbers table where a 0.25 [m.sup.2] quadrat was used to determine the number and percent coverage of plants. Taxonomy for plants followed Diggs et al. (1999) and scientific names with authorities are included in Tables 1-4. Once identified, vegetation was characterized at each observation point along each transect utilizing the following methodology as modified from USACOE (1987) and Brower et al. (1990).

Herbs were sampled in both wetlands. Herbs according to USACOE (1987) include all non-woody plants and woody plants < 1 m in height. In order to measure herbaceous vegetation, a 0.25 [m.sup.2] quadrat was randomly placed on the transect within the 1 [m.sup.2] observation point. Each species found within the quadrat was then identified, counted, and its percent cover estimated. An importance value (IV) was assigned to each herbaceous or woody seedling species found at each observation point, based on the sum of the relative coverage, relative density, and relative frequency of each species, while the percent relative importance value (%RIV) was determined by comparing the IV of each particular species to the total IV of all of the species (Fredrickson 1979 Brower et al. 1990, Smith 1996). A diversity index for all herbaceous and woody seedling species was applied using the Shannon-Weaver index, based on information obtained from the above data (Miller 1985 Brower et al. 1990). In addition, richness and evenness were calculated (Brower et al. 1990).

Those plants that have adapted to growth in water or in soil, which is at least periodically saturated or inundated, are often referred to as hydrophytic plants, or simply hydrophytes (USACOE 1987). Plants have been categorized, based on their life history into five separate groups called wetland indicator categories (USACOE 1987 United States Fish and Wildlife Service (USFWS) 1997 Tiner 1999). These include obligate wetland species (OBL) that occur in wetlands greater than 99% of the time. Facultative wetland plants (FACW-, FACW, and FACW+) occur in wetlands 67-99% of the time. Facultative plants (FAC-, FAC, and FAC+) occur in wetlands 34-66% of the time and facultative upland plants (FACU+, FACU, and FACU-) occur in wetlands 1-33% of the time. Upland plants (UPL) occur in wetlands less than 1% of the time. This classification system allows one to categorize vegetation in sampling units, and if greater than 50% of the dominant vegetation is OBL, FACW+, FACW, FACW-, FAC+, or FAC it is considered hydrophytic vegetation and the criterion for wetland vegetation in that sampling unit has been established (USACOE 1987). Each component of the vegetation data was classified by its wetland indicator status (USFWS 1997) and the 50/20 rule was then applied in order to determine dominance of species. The species were ranked in descending order of dominance and those species that immediately exceeded 50%, plus any additional species comprising 20% or more of the total dominance, were considered dominant. If more than 50% of the dominant plant species sampled within each observation point were obligate wetland (OBL), facultative wetland (FACW+, FACW, FACW-), or facultative (FAC+, FAC), the area covered by the observation point was considered positive for hydrophytic vegetation (USACOE 1987). In addition, a Mann-Whitney test was used to determine if there were significant differences between the constructed and natural wetland for herbaceous and woody seedling species (Brower et al. 1990).

Relative importance values (RIV%), a measure of dominance for herbaceous vegetation in the reference and constructed wetlands, were determined for each sampling period. During the Fall 2001 sampling period, common cat-tail (Typha latifolia 49.8%) and Bermuda grass (Cynodon dactylon 12.7%) were the most important plants identified in the reference wetland, while frogfruit (Lippia nodiflora 21.3%), Johnson grass (Sorghum halepense 19.2%), and barnyard grass (Echinochloa muricata 15.5%) were most important in the constructed wetland (Table 1). During the Spring 2002 sampling period, common cat-tail (77.3%) was found to be the most important species in the reference wetland, while in the constructed wetland, marsh-elder (Iva annua 34.0%) and fiddle dock (Rumex pulcher, 19.6%) were found to be dominant (Table 2). During the Summer 2002 sampling period the reference wetland was dominated by marsh-elder (42.7%) and Bermuda grass (19.6%), whereas the constructed wetland had marsh-elder (31.8%), and (for the first time) common cat-tail (24.4%) and knot grass (Paspalum distichum 23.7%) as dominants (Table 3). Knot grass (32.1%) and marsh-elder (18.7%) were found to be most important in the reference wetland during the Fall 2002 sampling period, and knot grass (20.0%), water-hemp (Amaranthus rudis 12.7%), cocklebur (Xanthium strumarium 12.4%), and marsh-elder (10.4%) all shared co-importance in the constructed wetland (Table 4). Total %RIV for the reference wetland, when the data was analyzed as a combined total, indicated common cat-tail (25.7%) being most important, followed by marsh-elder (25.6%) (Table 5). The constructed wetland was dominated by marsh-elder (17.6%), but had relatively high importance values for knot grass (17.5%), common cat-tail (10.6%), and cocklebur (8.9%) (Table 5).

In the Fall of 2001, the percentage of herbaceous plant species sampled in the reference wetland that were classified as hydrophytes was 70%, whereas 56% of the plant species sampled in the constructed wetland were classified as hydrophytes (Tables 1 and 5). One hundred percent of the plant species were classified as hydrophytes in the reference wetland Spring 2002, whereas 88% were classified as hydrophytes in the constructed wetland (Tables 2 and 5). During the next two sampling periods, the percentage of plant species classified as hydrophytes were about the same, 80 and 82% in the reference wetland vs. 83 and 78%, respectively, in the constructed wetland (Tables 3, 4, and 5). In the analysis of the data as a combined total, the reference wetland had a greater percentage of plant species classified as hydrophytes than did the constructed wetland, 83% and 72% respectively (Table 5).

Thirty-two species were identified from 80 quadrats in the two wetland areas. Of those 32 species, 13 (40.6%) were found to be present in both wetland areas, eight (25%) were found exclusively in the reference wetland and 11 (34.4%) were found exclusively in the constructed wetland (Table 5). Twenty-eight of the 32 species were natives, with four species being introduced. All four introduced species were found in the reference wetland (Bermuda grass, knot grass, Johnson grass, and fiddle dock and three of the four (excluding knot grass) were identified in the constructed wetland. The species identified in the reference wetland consisted of eight annuals and 13 perennials (Table 5), whereas the species identified in the constructed wetland consisted of 10 annuals and 14 perennials (Table 5).

Herbaceous plant species richness for Fall 2001 (Table 6) was slightly greater in the reference wetland (10) than in the constructed wetland (nine). However, during the Spring 2002 sampling period, the constructed wetland surpassed the reference wetland in plant species richness (eight and five, respectively) (Table 6). During the Summer 2002 sampling period, the reference wetland had higher richness with 11 species compared to eight in the constructed wetland (Table 6). During Fall 2002 (Table 6), the constructed wetland surpassed the reference wetland in species richness with 18 species, while the reference wetland had 11. When the data were analyzed as a combined total, the constructed wetland was found to have a slightly greater richness of 27 species, compared to 25 for the reference wetland (Table 6).

Evenness values (Table 6) were greater in the constructed wetland for every sampling period. Reference wetland evenness values ranged from a low of 0.38 in Spring 2002 to a high of 0.67 in the Fall 2001, for a combined total value of 0.48 (Table 6). Constructed wetland values ranged from a low of 0.53 in Summer 2002 to a high of 0.78 in Fall 2001, for a combined total value of 0.64 (Table 6).

Shannon's diversity index was calculated for each sampling period. Like evenness, diversity (Table 6) was found to be greater in the constructed wetland than in the reference wetland during every sampling period. Diversity values in the reference wetland ranged from a low of 0.62 in Spring 2002, to a high of 1.53 during Fall 2001. When the data were analyzed as a combined total, the Shannon's diversity value was 1.56. Diversity values in the constructed wetland ranged from a low of 1.10 during the Summer 2002 sampling period to a high of 1.86 in Fall 2002. When the data were analyzed as a combined total, a Shannon's diversity value of 2.11 was obtained (Table 6).

In the Fall 2001 sampling period, 100% of the quadrats sampled in the reference wetland met hydrophytic plant requirements (greater than 50% of the dominant species were FAC or wetter), while only 60% of those in the constructed wetland met the same requirements. During the Spring 2002 sampling period, 80% of the reference wetlands' quadrats met hydrophytic vegetation requirements compared to 40% of the constructed wetlands' quadrats. During the Summer 2002 sampling period, the percentage of quadrats meeting the hydrophytic vegetation requirements in both the reference and constructed wetlands had increased to 100 and 80% respectively. During the Fall 2002 sampling period, the percentage of quadrats meeting hydrophytic vegetation requirements in the reference wetland remained at 100%, while the constructed wetlands percentage dropped to 70. These results, when analyzed as a whole, show that the reference wetland met its hydrophytic vegetation requirements 95% of the time, whereas the constructed wetland met its hydrophytic vegetation requirements 63% of the time.

In addition to the above analyses, a non-parametric Mann-Whitney statistical test was conducted on richness, evenness, diversity, plant density, and percent plant coverage, but none were found to be significantly different in comparisons between the reference and constructed wetlands.

Using the standards of hydrophytic vegetation, soils, and hydrology (USACOE 1987), the reference area would be considered a wetland throughout this investigation. The constructed wetland had vegetation that was increasingly hydrophytic throughout the investigation, had about two-thirds of its soil samples classified as hydric, and had a constant hydrology, similar to the reference wetland, following the first sampling period.

Based on Relative Importance Values (RIV%) for the Fall 2001 sampling period, the most important species in the reference wetland were common cat-tail and Bermuda grass (Table 1). Common cat-tail can be an invasive species that has the ability to create large monocultures and out-compete other species. This is due to its rapid growth rate and the fact that it asexually reproduces using rhizomes, which spread quickly. Cat-tail invasion can be problematic because dense monocultures reduce the ratio of open water to vegetation, reduce the diversity of vegetation, and therefore reduce the habitat value of the site (Brown and Bedford 1997 Mitsch & Gosselink 2007). In addition, common cat-tail is an obligate wetland species (found in wetlands >99% of the time) and most of the reference wetland was found to be saturated for the majority of the year during the entire study. Therefore, common cat-tail would be well adapted for this environment. Each individual has the ability to release thousands of seeds, which can drift considerable distances in either the air or the water, giving it the ability to easily disperse.

Bermuda grass was also found to be one of the dominants in the reference wetland. However, it is a FACU+ species, which gives it about a 10% likelihood of being found in a wetland. The probability is low, but wetlands are dynamic ecosystems, which are constantly changing and Bermuda grass is an introduced species that can survive in a variety of conditions, often as a weed, in many parts of Texas (Diggs et al. 1999). It occurred along the wetland edge during the drier months of the growing season because the soil surface along the edge of the wetland had a tendency to dry, allowing species not generally adapted to wetlands a chance to compete with some of the wetland species, which may have died or become dormant. Bermuda grass has also been extensively introduced for livestock hay and grazing in fields adjacent to the Leon River.

In the constructed wetland, during Fall 2001, frogfruit, Johnson grass, and barnyard grass were found to be the most important species present (Table 1). Frogfruit has a wetland indicator status of FAC+ (about 60% chance of occurring in a wetland), Johnson grass is FACU (20% chance of occurring in a wetland), and barnyard grass is FACW (80% chance of occurring in a wetland), averaging a little more than 50% chance of occurring in the wetland. These species are often considered to be weedy, inhabiting disturbed areas (Diggs et al. 1999).

During the Spring 2002 sampling season, the reference wetland was dominated by common cat-tail, while the constructed wetland was co-dominated by marsh-elder and fiddle dock (Table 2). They were classified as FAC+ and FACW- (70% chance of occurring in a wetland) respectively. Marsh-elder is native, occurring in low, moist habitats, whereas fiddle dock is introduced and usually occurs in wet, disturbed areas (Diggs et al. 1999). The constructed wetland was being colonized by more wetland species, which would be expected following a fall and winter of flooding by reservoir personnel, in order to create waterfowl habitat. If the wetland hydrology stabilizes over time, the species composition should become more hydrophytic and change into a secondary seral stage or even a climax community.

During the Summer 2002 sampling period, the reference wetland was dominated by marsh-elder, and Bermuda grass (Table 3). The presence of Bermuda grass as a dominant is not necessarily the result of a shift in the reference wetland's vegetation, but a result of random sampling that occurred along the drier margins during that sampling period. Marsh-elder also dominated the outer core of the reference wetland, where the soil was drier, while common cat-tail dominated the inner core of the wetland, which was saturated most of the time.

During this same time period (Summer 2002), the constructed wetland was also dominated by marsh-elder, as well as common cat-tail and knot grass (Table 3), a FACW+ species (chance of occurring in a wetland 90%). Following the Fall/Winter inundation instigated by reservoir personnel, the constructed wetland was colonized with more hydrophytic species.

During the Fall 2002 sampling period, the reference wetland was dominated by knot grass and marsh-elder (Table 4). During this time of the year there was generally little precipitation, allowing the edge of the reference wetland to recede toward the core. This drying effect around the edges of the wetland allowed more facultative species a chance to re-establish themselves.

The constructed wetland was dominated by knot grass, waterhemp (FAC with a 50% chance of being found in a wetland), cocklebur (FAC- with a 40% chance of occurring in a wetland), and marsh-elder (Table 4). Water-hemp and cocklebur were two new co-dominants, both colonizing the edges of the wetland after drawdown of water levels.

Overall, the reference wetland was dominated by common cattail and marsh-elder (Table 5). The common cat-tail was found throughout the inner core of the wetland, while marsh-elder was found outside the core and toward the edges. These results follow the centrifugal organization model (Mitsch & Gosselink 2007 Keddy 2010), which states that the core habitat in wetlands has low disturbance and high fertility and is dominated by species that form dense canopies such as Typha. Peripheral habitats represent different kinds and combinations of stresses (infertility, disturbance) and support distinctive plant associations (Mitsch & Gosselink 2007). In this case, the stressor was the periodic draw down of the wetlands, allowing the core habitat to pulse inward and outward depending on the hydroperiod.

The dominant plant in the constructed wetland, overall, was marsh-elder (Table 5). In a study of excavated depressions near an earthen dam in north-central Texas, the dominant herb was also marsh-elder (Williams & Hudak 2005). However, in the Proctor Lake constructed wetland, relatively high importance values were found for knot grass, common cat-tail, and cocklebur. Knot grass and common cat-tail were not found in excavated depressions and cocklebur was rarely encountered in the excavated depressions (Williams & Hudak 2005). Each of these species are classified as FAC- or wetter. Overall, the constructed wetland was dominated by greater than 50% hydrophytes and like the reference wetland seemed to be developing centrifugal organization (Deberry & Perry 2005).

Species richness, overall, was found to be slightly greater in the constructed wetland (Table 6). This may be attributed to the fact that the constructed wetland was recently uninhabited due to the disturbance from construction. Early colonizing species took advantage of the uninhabited space and colonized the area causing a greater species richness, like that reported for constructed wetlands in Grimes County, Texas (Noon 1996). Successful early colonists have adaptations that enable them to establish quickly on open sites. They are generally small and low growing, have short life cycles, and reproduce annually by seeds or send out new shoots from buds near the ground (van der Valk 1981 Mitsch & Gosselink 2007). They produce large numbers of easily dispersed small seeds, respond quickly to disturbance (especially exposure of mineral soil), and attain dominance quickly by suppressing growth of any later serai stage plants that might exist as seedlings beneath them (Smith 1996 Mitsch & Gosselink 2007). These early colonists are tolerant of fluctuating environmental conditions particularly a wide range of daily temperatures on the soils surface, alternate wetting and drying, and intense light (Smith 1996 Mitsch & Gosselink 2007). These species represent what has been termed the arrival and establishment phase of wetland succession (Noon 1996) and the higher diversity found in the constructed wetland likely is a result of early successional phase. Richness values in restored and natural prairie wetlands that were adjacent to each other in northern Iowa, three years following restoration, had natural wetlands with 46 total species compared to only 27 total species occurring in restored wetlands (Galatowitsch & van der Valk 1996). Noon (1996), using data from eight constructed wetlands in Texas, found that the older the wetlands became, the fewer species occurred. Therefore, it is likely that, once the constructed wetland moves out of the arrival and establishment phase, species numbers will decrease.

Species present during the first few years of vegetation establishment that were not planted or seeded are assumed to be volunteers from offsite sources, and therefore represent the primary seral stage of vegetation succession (Reinhartz & Wame 1993 DeBerry & Perry 2004). In addition, vegetation in adjacent wetlands may be a potential seed source particularly from habitats dominated by herbaceous species that are more likely to colonize a young substrate (DeBerry & Perry 2004). Shared species composition between the wetlands was greater than 40% because the reference wetland likely served as a seed source for colonization of the constructed wetland.

The constructed wetland had greater diversity then the reference wetland during each sampling period (Table 6). Data from eight constructed wetlands in Texas (Noon 1996) indicated that diversity was highest at younger sites. Marshes, when compared to savannah type habitat, exhibited low species diversity in a study of wetland habitats along a portion of the North Fork of the Forked Deer River in West Tennessee. The savannah habitat was seasonally flooded, bringing in new seed sources and nutrients, while the marsh habitat retained water year around (Miller 1985). Ice scouring, infertile sandy soils, flooding by beavers, and open shorelines are among the stresses that shift communities to be less vegetatively productive, albeit possibly more diverse, assemblages (Mitsch & Gosselink 2007). Fluctuating hydroperiod in wetlands can promote dominance by annuals and nonclonal perennials in zones or patches within the wetland (Collins & Wein 1995 Mitsch & Gosselink 2007), and this patchiness and annual duration may increase diversity. Because, with the exception of the first sampling period, hydrology of the two wetlands was similar, differences in diversity are likely due to differences in successional stages between the two wetlands. The reference wetland is older and exhibited hydrophytic vegetation 95% of the time whereas the younger constructed wetland exhibited hydrophytic vegetation only 63% of the time. Assuming hydrology remains constant over time, the constructed wetland should develop more hydrophytic vegetation.

The adjacent riverine forest does not invade the marsh because the beaver dam and constructed wetland hydrology maintains a habitat too wet for the establishment of most tree species. Mitchell & Niering (1993) found that 30 year old beaver wetlands decimated the anchored forest wetland and replaced it with an abundance of graminoids in a northwestern Connecticut bog. The beaver-created marsh functions in a similar fashion and maintains separate marsh and upland forest communities. Because of more consistent hydrology and the older age of the reference wetland, common cattail, an obligate wetland species, dominates the inner core area with marsh-elder, a FAC+ species, occupying the drier periphery. Ninety-five percent of the species sampled in the reference wetland were hydrophytes. These results closely resemble a successional model of centrifugal organization that describes the distribution of species and vegetation types along gradients caused by a combination of environmental constraints (Wisheu & Keddy 1992 Mitsch & Gosselink 2007 Keddy 2010). In eastern North America, the core habitats are often dominated by species that form dense canopies such as Typha. Habitats peripheral to the core depend on different kinds and combinations of stress and disturbance. When beavers create disturbance, the peripheral habitats include species of sedge near the Typha core and wet-adapted forb species near the edges (Wisheu & Keddy 1992 Mitsch & Gosselink 2007 Keddy 2010). A similar plant community was observed in the reference wetland at Proctor Lake.

In contrast, the constructed wetland was dominated by marsh-elder and knot grass, both FAC+ species, with a greater diversity of vegetation of which 67% were hydrophytes. Richness, evenness, and Shannon's diversity were greater in the constructed wetland. This was likely due to its earlier successional stage with plants being a part of the arrival and establishment phase (Noon 1996).

Because common cat-tail was more important in later sampling in the constructed wetland it may develop according to the centrifugal organization with a Typha core (Deberry & Perry 2004). Mitsch et al. (2005) describes a naturally colonizing, riverine wetland in Ohio as being dominated by Typha. Noon (1996) proposed two phases in early constructed wetland primary succession called the arrival and establishment phase, which is characterized by random arrival and establishment of species, and the autogenic dominance phase, which is characterized by biomass production and competition between species. Because of its greater diversity, the constructed wetland is probably beginning to finish the first successional stage and is moving into the second phase whereas the reference wetland is in the autogenic dominance phase. In van der Valk's (1981) model of freshwater wetland vegetation dynamics, two basic types of wetland species are recognized: (1) species with long-lived propagules that are in the wetland seedbank and can grow when suitable conditions occur, and (2) species with short-lived propagules that can only grow in the wetland if they reach it during a period when conditions are suitable for germination. This causes the wetland to function as a sieve, allowing only establishment of certain species at any given time and dependent on whether the wetland is flooded or in a drawdown (van der Valk 1981 Mitsch & Gosselink 2007). At Delta Marsh in Canada and marshes of Eagle Lake in Iowa, flooding the marshes after drawdowns resulted in plant communities that contained or were dominated by a species of Typha (van der Valk 1981). Of herbaceous species identified in the Leon River wetlands, there were equal numbers of perennial species, but numbers of annuals differed (Table 5). The constructed wetland had more annual species than the reference (10 vs. 7) likely due to its early successional stages (Noon 1996) and more variable early hydrology, while the older riverine marsh, created by the beavers, was centrifugally organized (Wisheu & Keddy 1992 Deberry & Perry 2004 Mitsch & Gosselink 2007 Keddy 2010) with a perennial common cat-tail core.

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Jeffrey S. Brister (1) and Allan D. Nelson (2)

(1) Natural Resource Conservation Service, United States Department of Agriculture Service Center, 5040 Loop 340, Waco, Texas 76706

(2) Department of Biological Sciences, Box T-0I00, Tarleton State University, Stephenville, Texas 76402


2 MATERIALS AND METHODS

2.1 Aquatic plants

We selected E. densa (Hydrocharitaceae) as invader. It is a popular aquarium plant in Europe and world-wide the aquarium trade is considered its main introduction pathway (Yarrow et al., 2009 ). Egeria densa disperses mainly vegetatively for which fragments with only two nodes are enough to establish and develop new stands (Yarrow et al., 2009 ). The root system and shoots can break easily allowing plant fragments to be carried through the water to colonize new areas. This species can grow to over 3 m long and form monospecific stands with closed canopies, that can severely alter the structure of the native communities and local environmental conditions (Yarrow et al., 2009 ). It is well adapted to cold climates and can survive freezing winters by storing starch in its leaves and stems (Thiébaut, Gillard, & Deleu, 2016 ). Egeria densa has caused many problems throughout temperate regions including the United States of America and New Zealand, and has also become a nuisance species in its native range (Bini, Thomaz, Murphy, & Camargo, 1999 ).

The three common native submerged species that we used in the experiments are widely distributed in Northwestern Europe and co-occur in temperate shallow lakes (Van De Haterd & Ter Heerdt, 2007 ). Ceratophyllum demersum (Ceratophylaceae) is a free floating submerged species and M. spicatum (Haloragaceae) and P. perfoliatus (Potamogetonaceae) are rooted species. All three species are perennial and capable of clonal growth. All the plants used in this study were acquired from a commercial plant trader (De Zuurstofplantgigant, Hapert, The Netherlands). The acquired plants were pre-cultivated in 200 L cattle tanks (diameter = 66 cm and height = 60 cm, two tanks per species) under controlled greenhouse conditions with a 16/8 hr light/dark cycle at a temperature of 21 ± 3°C during the day and 16 ± 3°C during the night (Figure S1). The tanks were filled with a 3.4 kg (

2 cm) bottom layer of artificial plant pond sediment (Plant soil Moerings—Velda, organic matter = 34.31%), 44.9 kg (

10 cm) of washed sand on top (0.8–1.0 mm grain size, organic matter content = 0.16%) and filled with water from freshwater Lake Terra Nova (52°12′55.2″N, 5°02′25.7″E). Lake Terra Nova is a shallow peat lake located in the centre of the Netherlands where all three native plant species used in the experiment co-occur (Van De Haterd & Ter Heerdt, 2007 ). The lake is characterized by high nutrient concentrations in the water (water used in the experiment: M ± SD, n = 6 water samples, 0.14 ± 0.05 mg/L P-PO4 0.55 ± 0.46 mg/L N-NO3). The plants were cultivated under the following conditions: water temperature 22.3 ± 0.8°C, dissolved oxygen 12.5 ± 1.3 mg/L, conductivity 263 ± 28 µS/cm, pH 9.8 ± 0.3 and alkalinity 2.37 ± 0.52 mEq/L. Plants were pre-cultivated for at least 20 days before the start of the experiment.

2.2 Generalist herbivore

Lymnaea stagnalis (Gastropoda, Pulmonata, Basommatophora), the great pond snail, is a common and widely distributed generalist herbivore native to the Holarctic region. Most freshwater gastropod species consume mainly algae, bacteria and detritus but large species such as L. stagnalis can consume considerable amounts of aquatic plants having a large impact on aquatic plant abundance (Brönmark, 1989 , 1990 Wood et al., 2017 ). Densities of 10–40 L. stagnalis individuals/m 2 are commonly found under natural conditions (Elger et al., 2007 ), where it occurs in slow flowing and stagnant freshwater systems. This species has also been previous commonly used as model species in aquatic settings (Bakker et al., 2013 Elger & Barrat-Segretain, 2002 , 2004 Grutters et al., 2017 Zhang, Liu, Luo, Dong, & Yu, 2018 ).

Adult snails were collected from a pond located at the Netherlands Institute of Ecology (NIOO-KNAW, 51°59′16.8″N, 5°40′24.7″E, Wageningen, The Netherlands). They were acclimated to laboratory conditions for at least 2 weeks in 15 L buckets filled with groundwater at 20°C and constant aeration and exposed to a 16:8 hr day:night cycle, before being experimentally used. The snails were fed butterhead lettuce (Lactuca sativa L.) 6 days a week. Once a week fish food pellets (Velda, Gold Sticks Basic Food) and chalk were provided to ensure enough nutrients and calcium for shell development (following Grutters et al., 2017 ).

2.3 Experimental design and set-up

A greenhouse experiment was established at the Netherlands Institute of Ecology (NIOO-KNAW 51°59′15.3″N and 5°40′14.8″E) during the summer of 2018 (July–October). The experiment was set up as a full factorial randomized block design, with 3 × 3 × 2 treatment combinations of monocultures of three native submerged plant species (C. demersum, M. spicatum and P. perfoliatus), three levels of competition (no native plants, low density and high density) and absence (no snails) or presence of herbivory (with snails). The 18 treatments were replicated six times using a block design, yielding a total of 108 mesocosms (Figure S1). The greenhouse controlled conditions consisted of a 16/8 hr light/dark cycle at a mean temperature of 21 ± 3°C during the day and 16 ± 3°C during the night.

The mesocosms consisted of 13 L glass cylinder aquaria (18.5 cm diameter and 48 cm height) filled with a bottom layer of artificial plant pond sediment (150 g resulting in a layer of

1 cm depth) with a top layer of washed sand (2 kg resulting in a layer of

5 cm). Each aquarium was filled with 8 L lake water (resulting in 27 cm depth), leaving the upper 15 cm free to prevent snails from escaping. The water level was maintained constant during the whole experiment by refilling once a week with lake water to compensate for evapotranspiration. Abiotic parameters were monitored throughout the experiment and the growing conditions were found to be suitable for the plants (M ± SD, n = 1,166, water temperature 23.3 ± 1.0°C, dissolved oxygen 12.9 ± 1.9 mg/L, conductivity 283 ± 30 µS/cm, pH 9.7 ± 0.7 and alkalinity 2.12 ± 0.47 mEq/L).

To establish native plant communities for the competition treatment, we cut 99 non-rooted apical shoots without lateral shoots from the cultivation tanks from each of the native species C. demersum, M. spicatum and P. perfoliatus. We cut 15 cm long apical shoots and washed them in running tap water to remove any material attached. We randomly selected 15 of the 99 shoots of each species, dried these individually to a constant mass at 60°C for at least 48 hr, and weighed them for initial biomass measurements (dry weight, DW). We established the competition levels by pairing the invader E. densa with a single native plant species at different native shoot planting densities. The planting densities of each native plant species versus E. densa were manipulated to be 0:2 shoots (no competition, invader growing alone), 1:2 (low competition) and 6:2 (high competition), corresponding to

37 plants/m 2 (low competition) and

222 plants/m 2 (high competition) respectively before the invader introduction. These shoot densities are within the range observed in natural conditions (Li et al., 2015 ). The plant shoots of the rooted species were planted 5 cm deep in the sediment while the shoots of the non-rooted submerged species C. demersum were dropped in the water.

The native plants were left to establish for 2 weeks (24 July to 6 August) to allow the growth of at least one new shoot. Then, we introduced the invader by planting two E. densa non-rooted apical shoots per aquarium (7 August), which is considered to represent medium propagule pressure (Li et al., 2015 ). We chose shoots with an apical tip because these have a higher ability to regenerate, colonize and grow than shoots without apical tips (Riis, Madsen, & Sennels, 2009 ). To determine the introduced biomass in DW, we randomly selected 15 E. densa shoots, dried these to a constant mass at 60°C for at least 48 hr, and weighed them individually. Egeria densa was allowed to root for 2 days before we added the herbivore treatment, to simulate an early stage of establishment of E. densa in the new temperate native aquatic community.

In the herbivory treatment, we added two L. stagnalis snails per aquarium to half of our experimental units (10 August), representing intermediate snail densities observed in the field (Elger et al., 2007 ). We selected snails of the same size (shell length 30 ± 1 mm, wet weight 2.19 ± 0.27 g, M ± SD, n = 108) and starved the snails for 48 hr before adding them to standardize their appetite as is common practice in feeding trials (following Grutters et al., 2017 ).

2.4 Harvest and data collection

At the end of the experiment (after 8 weeks, on 8 October), we removed the herbivores, harvested the alien and native plants and, as we observed the growth of filamentous green algae Spirogyra sp. in our mesocosms, we harvested its biomass present on the plants and in the water column (Figure S2). We washed all the plants from each aquarium in an individual container to ensure that all the filamentous algae were kept. Then, this remaining water together with the water left in the aquarium after plant removal was filtered over a sieve of 0.106 mm mesh size. The filamentous algae biomass on the sieve was washed and dried to a constant mass at 60° for at least 48 hr, and weighed to determine DW. We measured invader E. densa performance in terms of the following growth parameters: total root and shoot DW, summing values from both introduced propagules and total biomass summing total root and shoot DW. We also determined native plant biomasses. All plants were dried to a constant mass at 60° for at least 48 hr, and weighed to determine DW.

2.5 Feeding trials

Herbivory consumption rates and preferences depend on plant palatability (Grutters et al., 2017 ). To determine plant palatability for the snails, we performed 24 hr no-choice feeding trials following established protocols (Elger & Barrat-Segretain, 2002 , 2004 Grutters et al., 2017 ). Plant material for the trials was collected from the same cultivation tanks that provided plants for the greenhouse experiment, and washed to remove any attached material. Snails of similar size (shell length 28.9 ± 1.8 mm, M ± SD, n = 48) were selected for the feeding trials.

Ninety-six plastic cups (volume of 500 ml) were filled with 375 ml groundwater (20°C, pH 8, conductivity 212 µS/cm). Twenty-four cups were used per plant species, of which each received approximately 0.2 g (wet weight) of non-apical shoots of either C. demersum, E. densa, M. spicatum or newly grown leaves of P. perfoliatus (one species per cup). Half of the cups received one individual of L. stagnalis whereas the other half was kept snail free, to be used as control to correct for autonomous changes in plant biomass due to growth. Snails were starved for 48 hr prior to the trial to standardize their appetite. All cups were covered with a mesh of size 1 mm to prevent snails from escaping. All cups were randomly placed on a rack in laboratory conditions at 20°C and exposed to a 16:8 hr day:night cycle (Figure S3). All snails were removed from their respective cup after 24 hr and euthanized by freezing at −20°C. Their soft body tissue was separated from their shells and dried in the oven at 60°C for at least 48 hr. The dry weights of plant fragments remaining in each cup were determined as described previously (see Section 2.4).

2.6 Data analyses

To disentangle the direct and indirect effects of native plants and herbivores on biotic resistance, we used piecewise Structural Equation Modeling (piecewiseSEM, Lefcheck, 2016 ). SEM has been shown to be an important tool to describe complex natural systems (Grace, Michael Anderson, Han, & Scheiner, 2010 ). For each of the three native plant species (C. demersum, M. spicatum and P. perfoliatus), we fitted models to investigate whether the native plants, herbivores, filamentous algae and their possible second-order interactions affected the invader E. densa performance (measured as total biomass at the end of the experiment). We fitted GLMM with block (the six replicates) as a random factor in all models (Pinheiro, Bates, DebRoy, Sarkar, & Team, 2018 ). We included these models in the SEM and performed model selection based on AICc criteria starting with the full model that included all second-order interactions among herbivory, filamentous algae biomass and native biomass. The best fitting models (lowest AICc) included only all main effects. Normality of model residuals, homoscedasticity and the influence of possible outliers were checked by visually inspecting plots of residual versus fitted values and quantile-quantile plots of model residuals. Native plant competition was evaluated using native plant species biomass as a continuous independent variable. PiecewiseSEM was performed in the software r (R Core Team, 2017 ) using the packages nlme and piecewiseSEM (Lefcheck, 2016 ).


Science and Products

Western Waters Invasive Species and Disease Research Program

Researchers at the Northern Rocky Mountain Science Center's Western Waters Invasive Species and Disease Research Program work extensively with federal, state, tribal, regional, and local partners to deliver science to improve early detection and prevention of invasive species and disease understand complex interactions that promote invasive species and disease, and their impacts (and.


The reproductive biology of male cottonmouths (Agkistrodon piscivorus): Do plasma steroid hormones predict the mating season?

To better understand the proximate causation of the two major types of mating seasons described for North American pitvipers, we conducted a field study of the cottonmouth (Agkistrodon piscivorus) in Georgia from September 2003 to May 2005 that included an extensive observational regime and collection of tissues for behavioral, anatomical, histological, and hormone analysis. Enzyme immunoassays (EIA) of plasma samples and standard histological procedures were conducted on reproductive tissues. Evidence from the annual testosterone (T) and sexual segment of the kidney (SSK) cycle and their relationship to the spermatogenic cycle provide correlative evidence of a unimodal mating pattern in this species of pitviper, as these variables consistently predict the mating season in all snake species previously examined under natural conditions. In most reptiles studied to date, high plasma levels of T and corticosterone (CORT) coincide during the mating period, making the cottonmouth an exception to this trend we suggest two possible explanations for increased CORT during spring (regulation of a spring basking period), and decreased CORT during summer (avoiding reproductive behavioral inhibition), in this species.


Organic Carbon Stocks in all Pools Following Land Cover Change in the Rainforest of Madagascar

Mieja Razafindrakoto , . Herintsitohaina Razakamanarivo , in Soil Management and Climate Change , 2018

Abstract

Land use change , along with the release of carbon (C) as carbon dioxide, constitutes a major source of emissions that contribute to climate change. Consequently, accurate carbon stock estimation is required to both inform and mitigate climate change. This study determined the importance of five C pools, including above-ground biomass (AGB), below-ground biomass (BGB), soil organic C (SOC), deadwood (DW), and litter, as well as the effect of land use change on these five pools for a region in eastern Madagascar. We assessed the importance of each pool, as well as the effect of land use change, on a closed-canopy forest (CC), tree fallow (TF), shrub fallow (SF), and degraded land (DL). Our results show that more C was stored in below-ground pools than in above-ground pools, and that SOC represented the largest (76.49%) contributor to the total C stock (186.64 Mg C ha − 1 ), followed by AGB (13.54%) and BGB (6.64%). DW represented an important pool in CC, representing 6.64% of the total C stock in this land use type. Conversely, the litter pool represented the lowest contribution to total C stock. Among the five pools, only the SOC showed little variation following land use change, while AGB, DW, and BGB were the most affected after deforestation and subsequent land degradation, most notably from CC to TF. The litter showed significant decreases of C stock from CC to TF and SF. These results highlighted the importance of considering all five pools in an accurate estimate of C stock for a better implementation of initiatives, such as “Reducing C Emissions from Deforestation and forest Degradation” (REDD +).


(21) The 4G Ranch Wetlands: Operating for Our Future

Presenter: Allison Lewis, Jacobs, [email protected]

Co-Authors: Rafael Vazquez-Burney, Jacobs Engineering, [email protected]

Abstract: In 2017, the largest groundwater recharge wetland in the world, known as the 4G Ranch Wetlands, was constructed in Pasco County. Groundwater recharge wetlands are constructed wetlands that do not have a surface water outflow and water is applied at the rate of infiltration to the underlying aquifer. The 4G Ranch Wetlands serve as a wet-weather management option for Pasco County’s reuse system and recharge 5 mgd on annual average to the surficial and Upper Floridan Aquifer. Located in an area suffering prolonged drawdown by regional wellfields , the 4G Ranch Wetlands also restore nearby hydrologically-altered lakes and wetlands.

Through a public-private partnership, the 3,000-acre 4G Ranch was identified as a suitable site for the infiltration wetland system. In 2015, the 176-acre groundwater recharge wetland was designed, and construction followed in 2016 and 2017. The 4G Ranch Wetlands are comprised of 15 individual cells that are each operated via water level measurements and flow control valves. Driven by the 4G Ranch’s desire to use the system for recreation, the wetland system includes several ecological design features and a mosaic of wetland habitats with transitional, shallow, and deep-water areas.

The wetlands have been in operation since 2017 and water levels of each wetland cell are adjusted seasonally to achieve healthy wetland hydroperiods and encourage the growth of desirable wetland species. Since operation, the 4G Ranch wetlands have been monitored for the success of the planted wetland vegetation establishment, the rate of infiltration, nitrate reduction, and presence and diversity of wildlife.

This presentation will describe the project, construction methods and lessons learned, and an update on the success of the overall wetland system following approximately two years of operation.

Biography: Allison joined CH2M now Jacobs as a water engineer after receiving her Masters in Environmental Engineering from the University of Florida in 2014. Her studies there focused on ecological engineering and wetlands. While attending UF, she had the opportunity to work with Dr. Bob Knight at Wetland Solutions where she gained experience in the ecological assessment of springs and the permitting and design of treatment wetlands in North Florida. Since joining Jacobs, Allison has supported various natural treatment systems projects including treatment wetland designs and ecological assessments, groundwater recharge wetland model updates, and biochemical reactor pilot studies and designs.


Spatial distribution of road-kills and factors influencing road mortality for mammals in Northern New York State

One of the most obvious impacts of roads on wildlife is vehicle-induced mortality. The aims of this study were to examine the spatial pattern of mammal–vehicle collisions (MVCs), identify and examine factors that contribute to MVCs, and determine whether the factors that increase the odds of MVCs are similar between species. On 103 road surveys that covered 7,094 total km I recorded the location of each MVC along the survey route. I measured landscape and roadway features associated with each MVC and used kernel density and network analysis tools to identify road mortality hotspots and measure spatial clustering of MVCs. I used logistic regression to model the likelihood of MVCs for all mammal data and separately for Porcupine (Erethizon dorsatum), Raccoon (Procyon lotor), Skunk (Mephitis mephitis), Muskrat (Ondatra zibethicus) and Cottontail (Sylvilagus floridanus) data sets. I identified 51 MVC hotspots and found spatial clustering of MVCs for Porcupines, Raccoons and Skunks. Two landscape variables, distance to cover and the presence of an ecotone, as well as one road variable, road width, appeared as broadly important predictors of mammalian road mortality, though there was also species-specific variation in factors that increased the risk of MVCs. Field-measured variables were more important than remotely-measured variables in predicting the odds of MVCs. Conservation implications are that mitigation of landscape features associated with higher risk of vehicle-collisions may reduce the number of MVCs in general, but species-specific research is required to more carefully tailor mitigation efforts for particular species.

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3 RESULTS

3.1 Compositional differences between wetlands

Estimated sample coverage was generally high (mean = 94%) indicating effective sampling of each taxonomic group per wetland type (Table 1). Within each site, a sufficient number of samples were taken to capture the majority of plant species, and however, sampling more beetles would have resulted in a greater number of species being found (Appendix S2). Half the total species pool of plants (156 species) and 45% of the species pool for beetles (66 species) was shared by both wetland types. For both taxonomic groups, a higher proportion of the total species pool was found only within BP (31% and 30% for plants and beetles respectively), compared to OW (19% and 20% respectively). These general differences can be visualized in the unconstrained ordination (Figure 1a,c). Despite the large overlaps in the hulls for each taxon group between wetland types, the mean species composition (as represented by the centre of each ordispider) differed significantly between wetlands for both plants (p < .001) and beetles (p = .034). Total beta diversity was strongly dependent on turnover for both plants and beetles (96% and 94% respectively), rather than nestedness (4% and 6%).

Group Wetland type (no. of plots surveyed) Total species observed Unique to wetland (% of overall total) Estimated sample coverage (%) No of significant indicators Rarity score
Aquatic plants BP (n = 250) 126 48 (30.8%) 98 27 1.46 ± 0.03 (1.00–4.16)
OW (n = 250) 108 30 (19.2%) 99 10 1.40 ± 0.03 (1.00–3.85)
Beetles BP (n = 50) 54 18 (30.0%) 88 2 1.91 ± 0.07 (1.00–3.50)
OW (n = 50) 47 12 (20.0%) 89 0 1.89 ± 0.09 (1.00–3.00)

A total of 37 plant species were significant indicators (p < .05) of a wetland type (Figure 1b), the majority being associated with BP. Only two beetle species were significantly associated with BP (Ilybius ater and Haliplus heydeni), and no indicator beetles were found for OW (Figure 1d).

Rarity scores of plants and beetles did not differ significantly between wetland types (plants: p = .496 beetles: p = .625), and none of the species found were listed as endangered or threatened on the Swedish red list (The Red List, 2015 ). Two non-native plant species were found in OW (Mimulus guttatus and Acorus calamus) and none in BP. However, both species were uncommon where present and occurred in <1% of plots sampled. No non-native beetle species were found.

3.2 Environmental basis for differences between wetlands

When both species assemblages were constrained by local environmental variables (see Appendix S1), the separation of the two wetland types was more distinct (Figure 2). In both cases, the overall constrained models were significant (p < .001 (plants) p = .018 (beetles)). For plants, plots from BP were associated with more woody debris, open and bare ground, while those in OW had greater leaf litter, plant height and plant coverage (Figure 2a). Water depth was the only significant environmental variable that explained beetle assemblages, though was driven by one outlying site (Figure 2b). When this outlier was removed, the overall model was not significant (p = .186). Wetland type accounted for a significant proportion of the compositional differences for plants (p < .001), over and above the effect of other variables, but not for beetles (p = .136). However, only 11% of variance in composition was explained in either model.

3.3 Differences in growth strategies between wetlands

No significant differences were found between wetland types in the representation of the competitor growth strategy in the quadrat-level vegetation (p = .16) (Figure 3a). However, in BP the representation of stress tolerators was significantly lower (p = .01), while ruderals were more common (p = .002) in comparison with OW (Figure 3b,c). Specifically among the subset of indicator species, the mean representation of growth strategies in BP indicator plant species (23.2 ± 3.7, 29.9 ± 5.3, 46.8 ± 5.4% for CSR respectively) contrasted strongly with the OW indicators (51.5 ± 11.1, 39.3 ± 10.8, 9.1 ± 3.5% for CSR respectively), highlighting a strong characterization of BP vegetation by ruderals and OW vegetation by competitors and stress tolerators.


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Watch the video: Beavers In Kwas Bay wetlands (January 2022).