July-December, 2015 Vol. 54, Nos. 3-4 THE MICHIGAN BOTANIST A Journal of Great Lakes Botany THE MICHIGAN BOTANIST (ISSN 0026-203X) is published four times per year by the Michigan Botanical Club (www.michbotclub.org) and is available online at http://quod.hb.umich.edu/rn/ mbot/. The subscription rate is $25.00 per year. Periodicals postage paid at Ann Arbor, MI 48103. On ah editorial matters, please contact Michael Huft, 232 Akela Dr., Valparaiso, IN 46385. Phone: (847) 682-5240; email: mhuft@att.net. All articles dealing with botany in the Great Lakes region may be sent to the Editor at the above address. In preparing manuscripts, authors are requested to follow the “Instructions for Authors” on the inside back cover. For all inquiries about back issues and institutional subscriptions, please contact The Michigan Botanist Business Office, 232 Akela Dr., Valparaiso, IN 46385. Phone: (224) 420-0326; email: michbot.business@gmail.com. Editorial Board Michael Huft, Editor L. Alan Prather Anton A. Reznicek J. Dan Skean, Jr. Anna K. Monfils Timothy M. Evans Catherine H. Yansa THE MICHIGAN BOTANICAL CLUB Membership is open to anyone interested in its aims: conservation of ah native plants; education of the public to appreciate and preserve plant life; sponsorship of research and publication on the plant life of the state and the Great Lakes area in general, both in the USA and in Canada; spon¬ sorship of legislation to promote the preservation of Michigan’s native flora; establishment of suitable sanctuaries and natural areas, and cooperation in programs concerned with the wise use and conservation of ah natural resources and scenic features. Dues are modest, but vary slightly among the chapters. To become a chapter member please contact the chapter presidents listed below. Annual dues include a subscription to The Michigan Botanist. Address changes for Chapter Members should go to the Chapter President. President: Judith Kelly, 18863 Lakewood Circle, Lake Ann, MI 49650; hfcckelly@gmail.com Treasurer: Bob Kelly, 18863 Lakewood Circle, Lake Ann, MI 49650; rgkelly49@gmail.com Great Lakes Chapter: Irene Eiselman, 1873 Pierce Road, Chelsea, MI 48118; eisemari@gmail.com Huron Valley Chapter: Anton Reznicek, University of Michigan Herbarium, 3600 Varsity Drive, Ste. 112, An n Arbor, MI 48109; reznicek@umich.edu Southeastern Chapter: Emily A. Metering, 231 Nash Street, Dearborn, MI 48124-1039; knietering@ sbcglobal.net Southwestern Chapter: Tyler Bassett, 2436 Woodward Ave., Kalamazoo, MI 49004-3481; keepitsimple7@yahoo.com White Pine Chapter: Dorothy Sibley, 7951 Walnut Avenue, Newaygo, MI 49337; dsibley@mail. riverview.net Volume 54, Nos. 1-2 was mailed March 29, 2016. 81 THE MICHIGAN BOTANIST Vol. 54 IN THIS ISSUE The preservation of natural ecosystems has necessarily become a topic of in¬ creasing concern in recent decades. The two principal articles in this issue deal with this common concern in two very different ways. The first article takes se¬ riously the threat posed by aggressive invasive plants. Most frequently, an inva¬ sive is a non-native species unrelated to plants in the native habitat. In this case, however, the threat is particularly pernicious, because the invasive element is an introduced genotype of a species already present as a native and posing no threat to the habitat in question, that is, common reed, Phragmites australis. The au¬ thors address the preliminary process of determining the presence of the differ¬ ent genotypes in habitats along the West Coast of Michigan and their distribu¬ tion. They also discuss the factors that might contribute to an increase in the presence of the invasive genotype and propose certain control measures. In an era of massive habitat destruction, the restoration of natural ecosystems has become an important aspect of overall preservation. The second article, based on a Master’s thesis, presents a thorough study of the effect of introducing species with various ecological roles in the restoration of a sand prairie, an ecosystem that was once common in the Upper Great Lakes region, but has now been reduced to vanishing remnants. The goal of the study is to determine whether the introduction of certain mixes of species in varying quantities has an effect on the diversity and floristic quality of the restored prairie while at the same time reducing the impact of non-native invasives. In this issue, we are making a change in the format for “Noteworthy Collec¬ tions” articles, following a suggestion by a contributor. “Noteworthy Collec¬ tions” will be a section of the journal, and each article in the section will have its own title. That way, the articles can more readily be cited and listed in bibliogra¬ phies in a way that makes their content apparent. Although it happens that each of the Noteworthy Collections articles in this issue treats a single species, that does not mean that articles submitted for this section with more than one item need to be divided into separate articles, each with its own title. Multiple-item articles may carry relevant titles, such as “Five New Reports for Michigan,” or whatever might be appropriate. The two articles in this issue confirm the pres¬ ence in western Michigan of American Lotus, not seen there since 1942, and re¬ port the first occurrence in Michigan of a non-native species that has been spreading in North America for some time. This issue closes with a review of two excellent field guides to a substantial portion of the graminoids of Wisconsin that should be useful throughout the western Great Lakes region. Michael Huft 82 THE MICHIGAN BOTANIST Vol. 53 A SURVEY OF NATIVE AND INVASIVE GENOTYPES OF PHRAGMITES AUSTRALIS ALONG MICHIGAN’S WEST COAST Grady H. Zuiderveen 1 Department of Ecosystem Science and Management Pennsylvania State University State College, Pennsylvania 16802 Timothy M. Evans, Thomas J. Schmidt, and Mark R. Luttenton Biology Department Grand Valley State University Allendale, Michigan 49401 ABSTRACT In North America, Phragmites australis (common reed) consists of a native North American group with several genetic forms and a highly invasive group with a single genetic form. Little is known about the environmental factors that affect the relative distribution of native and invasive pop¬ ulations, and it is often difficult to differentiate between the two groups based on physical character¬ istics alone. The western Lower Penin s ula of Michigan provides an excellent opportunity to evaluate the relationship between enviro nm ental gradients and the spread of Phragmites. The Lake Michigan coast presents a strong temperature gradient from the north to the south, whereas a moisture gradient is found from west to east from the Lake Michigan shoreline inland. The goal of this study is to as¬ sess the risk of invasion by the invasive genotype by: 1) determining the genotypes of individuals in Phragmites australis populations in western Michigan; and 2) evaluating the geographic distribution of the representative genotypes throughout the region. Examination of 58 samples of Phragmites australis using chloroplast gene markers from throughout western Lower Peninsula of Michigan yielded only the invasive genotype in the southern regions and a mixture of native, invasive, and Gulf Coast genotypes in the northern regions. Only one of the populations examined contained individu¬ als of both native and invasive genotypes. No trends were observed along the east-west gradient. KEYWORDS: Invasive genotype, Phragmites australis. Relative distribution. INTRODUCTION Wetlands are a critical habitat component of the Great Lakes region and serve several important functions, including wildlife habitat, sediment retention, and water purification (Keddy 2000). Anthropogenic disturbance (draining or filling) has led to substantial declines in the number and size of Michigan’s wetlands. Exotic invasive species, particularly invasive populations of Phragmites australis (common reed), have emerged as a significant threat to both coastal and non¬ coastal wetlands throughout the region. Phragmites australis is a perennial wet¬ land grass that can be found in both brackish and freshwater habitats (Saltonstall, 'Author for correspondence (gjz5033@psu.edu) 2015 THE MICHIGAN BOTANIST 83 2003a). It is found worldwide in marsh communities; along rivers, lakes, and ponds (Saltonstall 2002); and in roadside ditches. Currently, it is considered the most invasive plant species in wetland regions of the east coast of the United States (Rudrappa et al. 2007). Human factors known to aid in the expansion of the range of Phragmites in¬ clude ditch-digging for mosquito control and improved farm drainage (thereby lowering the water table), nutrient enrichment caused by runoff from lawns and farms (Amsberry et al. 2000), and possibly the transport of soil containing veg¬ etation fragments or seeds (e.g., Keller 2000). Historically, native Phragmites comprised only a small component of the vegetation of coastal marshes (Peter and Burdick, 2010), where it was restricted to the higher border regions, but the invasive Phragmites has more recently been noticeably invading the lower, wet¬ ter areas (Amsberry et al. 2000). North American Phragmites currently consists of a native North American group with several genetic forms (based on plastid genotypes) and an invasive group with one genetic form ( sensu Saltonstall 2002) that is extremely aggres¬ sive and that spreads quickly throughout wetlands. For example, Whyte et al. (2008) reported that in 1993, Phragmites australis occupied less than 1% of the lower wetland at the Old Woman Creek National Estuarine Reserve in Ohio and roughly 3% in 2003. By 2005, P australis dominated 22% of the lower wetland and was also a substantial component of the emergent plant community. Phrag¬ mites australis has gone from being a minor component of marshes to being a nuisance species as the native genotype has been replaced by the invasive geno¬ type, resulting in significant ecological impacts (Peter and Burdick 2010). In some wetlands, the spread of invasive Phragmites has eliminated much of the na¬ tive vegetation and converted diverse wetland plant communities to monospe¬ cific stands of common reed. The ability of Phragmites to colonize open areas rapidly is attributed to ef¬ fective methods of natural reproduction by the spread of hardy seeds and rhi¬ zomes (Saltonstall 2003b), through human practices, and by the dispersal of rhi¬ zome fragments. Another reason that Phragmites has been so successful is its clonal architecture of interconnected culms spanning over 10 meters (Burdick and Konisky 2003). Furthermore, some findings indicate that Phragmites exudes gallic acid, which generates high levels of reactive oxygen species in other plants, resulting in increased mortality (Rudrappa et al. 2007). It should be noted, however, that in a more recent study, Weidenhamer et al. (2013), refuted that claim, since they were unable to find significant amounts of gallic acid in or around invasive Phragmites. Phragmites now occupies significant portions of Great Lakes coastal wet¬ lands, including those along the eastern shore of Lake Michigan, and of isolated non-coastal wetlands. Although previous studies have been conducted to charac¬ terize the distribution of Phragmites genotypes across North America (Salton¬ stall 2003a), it appears that little has been reported about Phragmites populations in the upper Great Lakes region. It is therefore unclear which genotypes are pre¬ sent throughout the region. If the invasive genotype has become established, then the native genotypes of P australis, as well as the coastal and non-coastal wet¬ lands subject to invasion by the invasive genotype, are at risk, thereby affecting 84 THE MICHIGAN BOTANIST Vol. 54 native plant and animal species (Whyte et al. 2008). The goal of this study was to assess this risk by: 1) determining the genotype of individuals in Phragmites australis populations in western Michigan; and 2) evaluating the geographic dis¬ tribution of the representative genotypes throughout the study area. METHODS A total of 58 samples from 40 populations of Phragmites along Michigan’s west coast (Figure 1) were collected and georeferenced using a handheld GPS unit. Maps were created in ArcGIwwS 10 (ESRI 2011) using the locations of the georeferenced samples and projected in NAD 1983 State Plane, Michigan south international feet. Multiple samples were collected at 14 of the 40 collection sites where phenotypic (e.g., plant height, growth density) or environmental (e.g., soil moisture, species richness) differences were evident, or where the genet was spread over a large area. When Phragmites were restricted to small areas that appeared homogenous, only a single sample was col¬ lected as a single genet appeared likely. This sampling approach was designed to maximize effi¬ ciency, since the emphasis of this study is to determine a broad-scale pattern as opposed to compre¬ hensively mapping genotype distributions. Samples were collected from coastal and drowned-river mouth wetlands extending from Berrien County in Michigan’s southwest comer to Antr im County in the northwest portion of the Lower Peninsula. Additional collections were made from sites that rep¬ resented an east-west gradient in roadside ditches and wetlands away from the Lake Michigan coast for as far as 65 kilometers (40 miles). Fresh leaf tissue was collected and either frozen at -80°C or desiccated in silica gel. A voucher specimen for each sample was deposited in the Grand Valley State University herbarium (GVSC). DNA was extracted from leaf tissue using the DNeasy Plant Mini Kit made by Qiagen. Chloro- plast genes trnLb and rbcL were amplified using primers described in Saltonstall (2002). As de¬ scribed in Saltonstall (2002), trnLb fragments were digested with Rsal, and rbcL fragments were di¬ gested with Hhal, in each case to distinguish invasive from either native or Gulf Coast genotypes. RESULTS Of the 58 plant samples that were examined, 47 exhibited the invasive geno¬ type (Figure 1, Table 1). The invasive genotype was distributed throughout the entire range of the study, including at our most northern site. It was the only genotype found in the southern third of the Lower Peninsula, and it was the most common genotype in the study. Of the remaining 11 plant samples, six, from four populations, exhibited a na¬ tive genotype. These populations were restricted to the west-central and northern portions of the Lower Peninsula. Five plant samples from four populations ex¬ hibited the Gulf Coast genotype (Figure 1, Table 1). Populations that exhibited the Gulf Coast genotype were located in the west-central region of the Lower Peninsula. Of the 40 sites sampled, only two mixed populations were found. Both native and invasive genotypes were found at one site in northwestern Ottawa County, and another site in the same region of northwest Ottawa County con¬ tained both the invasive and the Gulf Coast genotypes. At the remaining 38 sites, no other mixed populations were found. This may have been due either to true homogenous strands or to cryptic mixing in which no differences in phenotype or environment were visually identifiable, and thus only one sample was collected. 2015 THE MICHIGAN BOTANIST 85 FIGURE 1. Location of sampling sites of Phragmites australis in western Michigan. Each point in¬ dicates a single individual that was sampled. The collection number, latitude, and genotype of each sample are given in Table 1. In addition, the genotype at each site is indicated on the map by the sym¬ bols ■ (native genotype), ▲ (Gulf Coast genotype), and • (invasive genotype). 86 THE MICHIGAN BOTANIST Vol. 54 TABLE 1. .Collection numbers of samples of Phragmites australis, listed in geographic order from south to north, with the latitude (second column) and the genotype (third column) of each collection. Collection No. Latitude Genotype 32 N 42°00.355 / invasive 33 N 42°00.372 / invasive 34 N 42“00.409' invasive 35 N 42°07.139 / invasive 30 N 42°09.505' invasive 31 N 42“09.531' invasive 36 N 42“14.715' Gulf Coast 27 N 42°22.603 / invasive 26 N 42°26.046 / invasive 25 N42°39.0ir invasive 24 N 42*44.311' invasive 21 N 42°47.009' invasive 22 N 42“47.659' invasive 23 N 42“47.777' invasive 9 N 42°49.19L invasive 10 N 42°49.215' invasive 11 N 42“49.216' invasive 8 N 42°49.239 / invasive 12 N 42“49.254' invasive 58 N 42“52.420' invasive 59 N 42°52.420' invasive 6 N 42°53.056 / invasive 7 N 42*53.057' invasive 3 N 42°53.254' invasive 2 N 42“53.255' invasive 1 N 42*53.257' invasive 49 N 42°57.560 / Gulf Coast 50 N 42°57.590 / Gulf Coast 5 N 42“57.844' invasive 4 N 42°57.869 / invasive 28 N 42“57.894' invasive 29 N42“59.410' invasive 13 N 43°04.029 / invasive 14 N 43“04.038' native 52 N43“04.160' Gulf Coast 51 N 43“04.190' invasive 19 N 43*04.291' invasive 16 N 43“04.553' Invasive 17 N 43“04.553' invasive 18 N 43“04.553 / invasive 20 N 43*04.571' invasive 15 N 43“04.604' invasive 48 N 43“08.775' invasive 56 N 43°13.060' invasive 57 N 43“13.060' invasive 39 N 43“15.537' native 40 N 43“15.623' native 53 N 43*21.450' Gulf Coast 54 N 43*21.450' invasive 55 N 43°23.140' invasive 41 N 43“23.217' invasive 42 N 43“45.770' native 43 N 43“45.775' invasive 44 N 44“15.736' invasive 60 N 44“33.360' invasive 46 N 44“52.969' native 45 N 44“52.984' native 47 N 45“10.174' invasive 2015 THE MICHIGAN BOTANIST 87 DISCUSSION The specific arrival time of the invasive genotype of Phragmites to the Mid¬ west is still unknown. Since the morphological characteristics of the invasive genotype are very similar to those of the native genotype, the presence of the in¬ vasive genotype likely went unnoticed for a number of years. The invasive Phragmites genotype dominates both coastal and noncoastal sample locations (including several wetlands) south of Ottawa County, which suggests that the in¬ vasion may have been from south to north in the Lower Peninsula. Given its broad tolerance and aggressive nature, the invasive genotype of Phragmites could eventually dominate most or all of the wetlands in Michigan, regardless of latitude. Wetlands crossed by roads may be most vulnerable to in¬ vasion by invasive Phragmites. Indeed, drainage ditches along county roads and highways have been found to support robust populations in Michigan and may play a large role in their spread by serving as an initial habitat for invasive Phrag¬ mites throughout the region (Jodoin et al. 2008; Lelong et al. 2007). Because most of our sample sites were adjacent to roads, it is unclear whether the same pattern applies to wetlands in more isolated areas. However, isolated wetlands may be equally vulnerable to alternate routes of invasion. The location of the Benzie County sample site along an isolated lakeshore, for example, suggests that more isolated areas are not immune from the invasive genotype. In these cases, invasion may be through natural dispersal or human agency. For example, human factors such as shoreline development may play a large role in the spread of invasive Phragmites. Silliman and Bertness (2004) found that the removal of woody vegetation along the edges of marshes explained 91% of the variation in Phragmites cover among the marshes of Narragansett Bay in Rhode Island. In western Michigan, riparian residents often mow wet areas in an attempt to main¬ tain their view and to control Phragmites. However, since Phragmites appears to recover more quickly than native wetland species after a mowing (Luttenton, per¬ sonal observation), this process appears to be more detrimental to other vegeta¬ tion and could promote the establishment of monospecific stands by Phragmites. Another plausible factor that may have contributed to the increase in invasive Phragmites in the Great Lakes region could be the decrease in water levels over the past three decades to levels that are comparable to historic mean lake levels (Great Lakes Environmental Research Laboratory 2016). The previous high water levels may have killed off the woody species in coastal communities, thereby leaving space for species that thrive in standing water. Although yearly water level fluctuations are common in the Great Lakes, the recent decrease in water levels to historic lows would leave large areas of exposed sediments that are ideal for the opportunistic and aggressive Phragmites (Whyte et al., 2008). Furthermore, it has been proposed that Phragmites may often become estab¬ lished initially in the high marsh and then expand into the lower, less ideal, areas by clonal integration (Amsberry et al. 2000). If this is the case, even relatively small amounts of disturbance, whether caused by humans, by decreased water levels, or by any of a number of other mechanisms, could result in the expansion of already-present Phragmites. Because it has become so common in many parts THE MICHIGAN BOTANIST Vol. 54 of the United States, invasive Phragmites is considered to be an accurate gauge of wetland disturbance (Saltonstall 2002). A surprising finding in this study was the presence of the Gulf Coast geno¬ type at several sites. The results of this study might lead one to concluded that the Gulf Coast genotype is as common in western Michigan as the native geno¬ type. However, the incidence of the native genotype is likely to have been under¬ estimated due its greater prevalence in northern Michigan, which is less exten¬ sively represented in this study. Previous studies had indicated that the Gulf Coast genotype was confined to the Gulf Coast region (Saltonstall et al. 2004), but it now appears that it has spread northward. The presence of the Gulf Coast genotype may be the result of fill transport from construction. Keller (2000) found two distinct populations of Phragmites in New England that were hypoth¬ esized to be the result of transport of construction fill during the 1950s. Like¬ wise, it has been hypothesized that the early spread of Phragmites was facilitated by boating activities (Chambers et al. 2012), for which the Great Lakes region is renowned. The presence of two populations in Ottawa County that contain both native and invasive genotypes in single locations raises the question of whether hy¬ bridization between native and invasive plants has taken place. Saltonstall (2011) reported that the native lineages may maintain a pure genetic composition even in mixed stands. However, Meyerson et al. (2010) found that in a lab setting, hy¬ bridization between native and invasive genotypes of Phragmites resulting in vi¬ able offspring can take place, which could present further complications if it oc¬ curred in the wild. Meyerson et al. (2010) also found that there is considerable overlap in flowering phenology between the native and invasive genotypes, which could facilitate natural hybridization. Earlier, Saltonstall et al. (2004) doc¬ umented a single instance of hybridization in the field. Hybridization has been shown to result in higher seed viability in the invading hybrids, and the resulting progeny can be the most vigorous in species in which hybridization occurs (Ayres et al. 2008). Eurther, increased genetic diversity can nullity deleterious al¬ leles by forming new favorable genetic combinations (Saltonstall 2003a), mak¬ ing invasive species even more threatening to natural habitats. Municipalities and local NGOs have recently instituted programs to control the spread of Phragmites, for example, the Ottawa County Invasive Phragmites Control Group (Great Lakes Phragmites Collaborative 2010) and The Watershed Center Grand Traverse Bay (2016). In Ottawa County, where we have found mixed stands of genotypes, several extensive Phragmites populations have re¬ cently been eradicated by means of aggressive control measures (Luttenton, per¬ sonal observation). It is possible to differentiate the native genotype from the Gulf Coast and in¬ vasive genotype based on botanically technical differences in morphological fea¬ tures including ligule, glume, and lemma length (Saltonstall et al. 2004; Judziewicz et al. 2014). However, using these morphological differences would limit Municipalities and NGOs time of treatment to when reproductive parts are present and would require some botanical expertise. Therefore, given the simi¬ larity between the native and invasive forms, we suspect that eradication efforts 2015 THE MICHIGAN BOTANIST 89 may not discriminate between them. Consequently, native populations, which pose little threat to native wetland communities, may nevertheless be lost. CONCLUSION Invasive Phragmites has successfully colonized Michigan wetlands and pre¬ sents a significant threat to both coastal and non-coastal wetlands. Phragmites is known to be a habitat-modifying organism in that its vigorous growth in lower marshes can result in reduced sediment erosion and increased water uptake and transpiration, resulting in a lowering of the water table (Amsberry et al. 2000). Silliman and Bertness (2004) found that Phragmites dominance in developed wetlands leads to substantial decrease in plant species richness. This endangers both native plants and the wildlife they support, including migratory birds de¬ pendent upon native plants for nesting. Eradication of the invasive genotype has been difficult, and sites in Grand Haven, Michigan, were found in this study in which, although chemical applica¬ tions had recently been applied, young shoots of invasive Phragmites could nev¬ ertheless be observed. Chemical application is a commonly used treatment for eradicating Phragmites. However, a single treatment is ineffective in perma¬ nently removing Phragmites (Warren et al. 2001); therefore, chemical treatment becomes expensive with repeated application, the only known advantage over al¬ ternative treatment methods being that it has no major impact on fish popula¬ tions (Fell et al. 2006). One alternative to a chemical method is biological con¬ trol as described by Blossey (2003). He claims that the vigorous nature of the invasive Phragmites could be due to the lack of herbivory, allowing the plant to reallocate its resources from fighting herbivory toward more rapid and dense growth. Therefore, if a specialized natural herbivore of Phragmites were intro¬ duced into North America, the competitive edge of having no enemies that the invasive species now enjoys might be removed, ultimately allowing wetlands to be restored to their original state. Such a method would have an inherent risk that the newly released herbivore would consume plants other than the desired tar¬ get—including the native genotype. However, since no other species of Phrag¬ mites are found in North America, the likelihood of such nontarget effects is low (Pemberton 2000). Alternatively, methods of preventing the further spread of invasive Phrag¬ mites into the more northern regions of Michigan should be seriously consid¬ ered. One method that has been successful has been to maintain high levels of native plants in regions at risk for the spread of Phragmites. Specialist native competitors provide an ecological barrier to the invading Phragmites by compet¬ ing for essential resources and reducing available niches (Peter and Burdick 2010). Phragmites tends to colonize disturbed upland boarders of wetlands first, then to expand into the central regions (Warren et al. 2001). Reducing human disturbance and maintaining species richness in Michigan wetlands is an option worth seriously considering, since decisions regarding management of invasive Phragmites have long-term repercussions (Ludwig et al. 2003). Future studies 90 THE MICHIGAN BOTANIST Vol. 54 would be needed to further determine the northern extent of the invasive geno¬ type in order to determine where preventive measures might be the most effec¬ tive as, at least for now, the invasive genotype of Phragmites, once established, is here to stay. ACKNOWLEDGMENTS This research was funded by grants from the Garden Club of America, the Waddell-Treanor Na¬ tive Plants Endowment, and the Hanes Trust. LITERATURE CITED Amsberry, L., M. A. Baker, P. J. Ewanchuk, and M. D. Bertness. (2000). Clonal integration and the expansion of Phragmites australis. Ecological Applications 10: 1110-1118. Ayres, D. R., K. Zaremba, C. M. Sloop, and D. R. Strong. (2008). Sexual reproduction of cordgrass hybrids ( Spartina foliosa xalterniflora) invading tidal marshes in San Francisco Bay. Diversity and Distributions 14: 187-195. Blossey, B. (2003). A framework for evaluating potential ecological effects of implementing biolog¬ ical control of Phragmites australis. Estuaries 26: 607-617. Burdick, D. M., and R. A. Konislcy. (2003). Determinants of expansion for Phragmites australis, common reed, in natural and impacted coastal marshes. Estuaries 26: 407-416. Chambers, R. M., L. A. Meyerson, and K. L. Dibble. (2012). Ecology of Phragmites australis and re¬ sponses to tidal restoration. Pp. 81096 in Tidal marsh restoration: A synthesis of science and man¬ agement. C. T. Roman and D. M. Burdock, editors. Island Press. Washington, D.C.. ESRI. (2011). ArcGIS Desktop: Release 10. Environmental Systems Research Institute, Redlands, California. Fell, P. E., R. S. Warren, A. E. Curtis, and E. M. Steiner. (2006). Short-term effects on macroinverte¬ brates and fishes of herbiciding and mowing Phragmites australis- dominated tidal marsh. North¬ eastern Naturalist 13: 191-212. Great Lakes Phragmites Collaborative. (2016). Program: Ottawa County Invasive Phragmites Con¬ trol Group. Available at greatlakesphragmites.net/program-ottawa-county-invasive-phragmites- control-group/. (Accessed March 14, 2016). Great Lakes Environmental Research Laboratory. (2016). Great Lakes Water Level Dashboard. Available at www.glerl.noaa.gov/data/dashboard/GLWLD.html. (Accessed March 14, 2016). Jodoin, Y., C. Lavoie, P. Villeneuve, M. Theriault, J. Beaulieu, and F. Belzile. (2008). Highways as corridors and habitats for the invasive common reed Phragmites australis in Quebec, Canada. Journal of Applied Ecology 45: 459-466. Judziewicz, E. J., R. W. Freckmann, L. G. Clark, and M. R. Black. (2014). Field guide to Wisconsin grasses. University of Wisconsin Press. Madison. Keddy, P. A. (2000). Wetland ecology: Principles and conservation. Cambridge University Press, New York, N.Y. Keller, B. E. M. (2000). Genetic variation among and within populations of Phragmites australis in the Charles River watershed. Aquatic Botany 66: 195-208. Lelong, B., C. Lavoie, Y. Jodoin, and F. Belzile. (2007). Expansion pathways of the exotic common read {Phragmites australis)'. A historical and genetic analysis. Diversity and Distributions 13: 430-437. Ludwig, D. F., R. J. Iannuzzi, and A. N. Esposito. (2003). Phragmites and environmental manage¬ ment: A question of values. Estuaries 26: 624-630. Lynch, E. A., and K. Saltonstall. (2002). Paleoecological and genetic analyses provide evidence for recent colonization of native Phragmites australis populations in a Lake Superior wetland. Wet¬ lands 22:637-646. Meyerson, L. A., D. V Viola and R. N. Brown. (2010). Hybridization of invasive Phragmites australis with a native subspecies in North America. Biological Invasions 12: 103-111. Pemberton, R. W. (2000). Predictable risk to native plants in weed biocontrol. Oecologia 125: 489-494. 2015 THE MICHIGAN BOTANIST 91 Peter, C. R., and D. M. Burdick. (2010). Can plant competition and diversity reduce the growth and survival of exotic Phragmites australis invading a tidal marsh? Estuaries and Coasts 33: 1225-1236. Rudrappa, T., J. Bonsall, J. L. Gallagher, D. M. Seliskar, and H. O. Bais. (2007). Root-secreted alle- lochemical in the noxious weed Phragmites australis deploys a reactive oxygen species response and microtubule assembly disruption to execute rhizotoxicity. Journal of Chemical Ecology 33: 1898-1918. Saltonstall, K. (2002). Cryptic invasion by a non-native genotype of the common reed, Phragmites australis, into North America. Proceedings of the National Academy of Sciences 99: 2445-2449. Saltonstall, K. (2003a). Genetic variation among North American populations of Phragmites aus¬ tralis: Implications for management. Estuaries 26:444-451. Saltonstall, K. (2003b). Microsatellite variation within and among North American lineages of Phragmites australis. Molecular Ecology 12: 1689-1702. Saltonstall, K. (2011). Remnant native Phragmites australis maintains genetic diversity despite mul¬ tiple threats. Conservation Genetics 12:1027-1033. Saltonstall, K., P. M. Peterson, and R. J. Soreng. (2004). Recognition of Phragmites australis subsp. americanus (Poaceae: Arundinoideae) in North America: Evidence from morphological and ge¬ netic analysis. SIDA, Contributions to Botany 21: 683-692. Silliman, B. R., and M. D. Bertness. (2004). Shoreline development drives invasion of Phragmites austalis and the loss of plant diversity on New England salt marshes. Conservation Biology 18: 1424-1434. The Watershed Center Grand Traverse Bay. (2016). Invasive Species. Available at www.gtbay.org/ our-programs/invasive-species/. (Accessed March 14, 2016). Warren, R. S., P. E. Fell, J. L. Grimsby, E. L. Buck, G. C. Rilling, and R. A. Fertik. (2001). Rates, pat¬ terns, and impacts of Phragmites australis expansion and effects of experimental Phragmites con¬ trol on vegetation, macroinvertebrates, and fish within tidelands of the lower Connecticut River. Estuaries 24: 90-107. Weidenhamer, J. D., M. Li, J. Allman, R. G. Bergosh, and M. Posner. (2013). Evidence does not sup¬ port a role for gallic acid in Phragmites australis invasion success. Journal of Chemical Ecology 39: 323-332. Whyte, R. S., D. Trexel-Kroll, D. M. Klarer, R. Shields, and D. A. Francko. (2008). The Invasion and spread of Phragmites australis during a period of low water in a Lake Erie coastal wetland. Jour¬ nal of Coastal Research SI 55: 111-120. 92 THE MICHIGAN BOTANIST Vol. 54 THE INITIAL EFFECTS OF COMMUNITY VARIABLES ON SAND PRAIRIE RESTORATION: SPECIES ESTABLISHMENT AND COMMUNITY RESPONSES Robert C. Roos 6140 Cottonwood Drive, Suite A Fitchburg, Wisconsin 53719 robb.roos@cardno.com Todd A. Aschenbach Grand Valley State University 1 Campus Drive Allendale, Michigan 49401 aschenbt@gvsu. edu ABSTRACT The tallgrass prairie was one of the most wide-ranging and diverse ecosystems in North Am erica. This diverse ecosystem comprises a mosaic of prairie types that, in northern Lower Michigan, was historically dominated by dry sand prairie. As a result of fire suppression, silvicultural and agricul¬ tural activities, and degradation by invasive species, only approximately 4% of the original extent of sand prairie remains intact in the state. Despite the important ecological role and increasing scarcity of sand prairie, restoration and management of this ecosystem has been severely understudied. In order to gain a better understanding of this ecosystem and of potential restoration techniques that might influence its community variables, a sand prairie restoration experiment was established in the pine-oak barrens of northern Lower Michigan to analyze how different seeding treatments affect cer¬ tain community variables, including vegetative cover, species richness, diversity, and floristic quality. The seeding treatments were varied with respect to seeding concentrations (1,000 seeds/m 2 and 10,000 seeds/m 2 ); and the inclusion of grasses and/or forbs with diverse ecological characteristics, such as early flowering, late flowering, and nitrogen fixers (i.e., legumes). Measurements of the com¬ munity variables were taken during each of the first three growing seasons following seeding. In gen¬ eral, treatments that included a high concentration of grass and/or an early season forb component had the greatest overall positive impact on plant community development. These treatments resulted in significantly greater diversity and higher floristic quality, as well as a lower percentage of non-na¬ tive or invasive cover than other treatments. The benefit of high concentrations of grasses and early season forbs may play a critical role in initial species establishment of a sand prairie restoration due to the facilitative and competitive advantages they may provide in these harsh environments. How¬ ever, it remains to be seen if these initially successful communities will have continued success over longer periods of time. KEYWORDS: Sand prairie, Community, Establishment, Restoration. INTRODUCTION The tallgrass prairie was one of the most wide-ranging and diverse ecosys¬ tems in North America. Today, only approximately 0.1% of its original extent re¬ mains, making tallgrass prairie one of North America’s most endangered ecosys¬ tems (Samson and Knopf 1994). The extensive loss and rapid decline of this 2015 THE MICHIGAN BOTANIST 93 ecosystem can be attributed primarily to European settlers. The dark, rich, tree- free, and easily manipulated Mollic soils of the tallgrass prairie made it a prime target for agricultural use (Howe 1994). This ecosystem has also experienced many other assaults as a result of human development. Fire, which was once commonplace in the Midwestern United States, became actively managed and suppressed (Anderson 1990). Habitat fragmentation, the introduction of invasive non-native species, and the increased establishment of invasive native species also promoted habitat degradation (Noss et al. 1995; Cully et al. 2003). The reduction of tallgrass prairie and its associated biodiversity has resulted in a substantial loss of ecosystem function. Higher levels of plant diversity are associated with greater ecosystem productivity (Tilman and Downing 1994; Hector et al. 1999; Tilman 1999), nutrient use efficiency (Risser 1988; Tilman 1997), resistance to invasion by exotic species (Tilman and Downing 1994; Kennedy et al. 2002; Pokorny et al. 2005), and resistance to environmental change (Ives et al. 2000). According to the insurance hypothesis of Yachi and Loreau (1999), high levels of biodiversity insures ecosystems against declines in their functioning, because the presence of many species guarantees that some will remain functioning even if others fail (Ives et al. 2000). Ecosystem functions that are associated with tallgrass prairie include carbon sequestration, water fil¬ tration, erosion control, soil enhancement, and nutrient cycling (Raison 1979; Seastedt and Knapp 1993; Wedin and Tilman 1990). Since diversity is a benefi¬ cial and necessary attribute that drives a fully-functioning, healthy ecosystem, it is useful to examine the factors that impact diversity. The tallgrass prairie ecosystem comprises a mosaic of prairie types that in¬ clude xeric, mesic, and wet components that sometimes transition into barrens and savannah ecosystems. Tallgrass prairie historically extended eastward through Indiana and into areas of Michigan, Ohio, and Kentucky. This region is referred to as the prairie peninsula. These fingers of grassland followed areas where climate fluctuated enough to support a mosaic of prairie community types, including oak-pine barrens and oak savanna (Transeau 1935; Anderson 1990). In northern Lower Michigan, north of the tension zone (approximately 43°N latitude), grassland was predominantly dry sand prairie (McCann 1991; Kost 2004). As a component of these open, upland mosaics, sand prairie was a primary component of roughly 5,000 hectares of northern Lower Michigan’s nat¬ ural landscape in the early to mid-1800s (Comer et al. 1995). This dry grassland is considered the driest ecosystem east of the Mississippi River (Schaetzl and Anderson 2005). Plant species of the sand prairies are similar to those of a xeric tallgrass prairie, but due to water, heat, and nutrient stresses, the sand prairie vegetation is typically shorter in stature and separated by patches of bare ground. The combination of wildfires, dry and well-drained soils, and the harsh frosts that are associated with northern Michigan have historically maintained these ecosystems (Kost et al. 2007). As a result of fire suppression, silvicultural and agricultural activities, and degradation by invasive species, only approximately 4% of the original extent of sand prairie remains intact in the state (Hauser 1953; Albert and Comer 2008). Consequently, sand prairie is considered one of Michigan’s most endangered 94 THE MICHIGAN BOTANIST Vol. 54 ecosystems. Today, less than 200 hectares of high quality sand prairie still exist in Michigan (Kost 2004). The loss and degradation of the sand prairie has had negative consequences for species that are associated with these ecosystems. Over 25 plant and 30 ani¬ mal species are dependent on these dry grasslands for either all or part of their lives (Kost 2004). This includes the federally endangered Lycaeides melissa samuelis (Karner blue butterfly), a species that depends on Lupinus perennis L. Grassland birds have shown a decline that is greater than that of any other group of North American species (Knopf 1994). Dendroica kirtlandii (Kirtland’s war¬ bler) is a species that depends solely on the matrix of sand prairie and pine bar¬ rens comprised of Pinus banksiana Lamb, in northern Lower Michigan for sur¬ vival (Kost et al. 2007). Common northern dry sand prairie species, that nevertheless are state-listed due to the degradation of this community type, in¬ clude Festuca altaica Trin. Ex Ledeb., Cirsium hillii (Canby) Fernald, and Agoseris glauca (Pursh) Raf.. Despite the important ecological role and increasing scarcity of sand prairie, restoration and management of this ecosystem has been severely understudied. Published research on sand prairies within the eastern prairie peninsula, includ¬ ing the portion in Michigan, is sparse. Instead, the majority of sand prairie liter¬ ature has focused on prairie regions to the west of the Great Lakes (Gleason 1910; Plumb-Mentjes and Center 1990; Cole and Taylor 1995; Bowles et al. 2003). Studies that have addressed the Michigan sand prairies have focused on descriptive analyses (Hauser 1953; Albert 1995; Comer et al. 1995; Kost 2004), or on comparative assessments with other community types, such as dunes and jack-pine barrens (Houseman and Anderson 2002; Emery et al. 2013). To date there have been few, if any, studies that have focused on the restoration of Michi¬ gan’s true sand prairies. Restoration projects in general tend to place an emphasis on many aspects of community structure and ecosystem processes. Although certain components of restoration projects have been successful, they have yet to achieve the goal of creating a historic, natural community (Martin et al. 2005). Current methods of restoring prairie communities are based on a weak scientific rationale that is not consistent with the history of how grasslands formed, and in fact may threaten biodiversity (Howe 1994). Although plant community restoration has the poten¬ tial to help re-establish lost diversity and ecosystem function (Foster et al. 2007), many plant community restoration attempts have not fully re-established the di¬ versity and function found in remnant prairie communities (Sluis 2002, Polley et al. 2005). Sub-optimal results of previous restoration attempts have led ecologists to ex¬ amine different theories of plant community succession and species coexistence in order to identify more successful approaches to restoration (Cairns and Heck¬ man 1996). Community assembly theory is one approach that asks how species arriving at a site form an initial community (Belyea and Lancaster 1999). This approach integrates aspects of succession and species coexistence in an effort to examine how species introductions, biotic interactions (e.g., competition), and abiotic conditions (e.g., soil nutrients) influence community development (Lock- wood 1997; Belyea and Lancaster 1999; Young et al. 2001). 2015 THE MICHIGAN BOTANIST 95 The influence of environmental factors on plant communities is complex, and multiple factors influence plant diversity and community development (Grace 1999). It is well established that the environment influences plant community succession (Tilman 1988; Howe 1995), the frequency and intensity of distur¬ bance (Collins et al. 1995; Suding 1999; Collins 2000), competitive intraspecific and interspecific interactions and the plant’s ability to respond (Grime 1974; Goldberg and Barton 1992; Smith et al. 1999), the availability of nutrients (Rai¬ son 1979; Ojima et al. 1994), and the ability of plants to be productive both veg- etatively and reproductively (Zimmerman and Kucera 1977; Gough et al. 1994; Tilman et al. 1996). A better understanding of how all of these factors influence community development will allow for more practical approaches to community restoration. As one approach to understanding that question, we established a restoration experiment on a degraded site previously occupied by sand prairie in northern Lower Michigan in 2009. We introduced various combinations of native plants belonging to four functional groups—legumes, early flowering forbs, late flow¬ ering forbs, and warm-season grasses—at two different seeding concentrations in an attempt to see how these initial seeding treatments affect plant community development over time. Here we present the results obtained after two complete growing seasons and compare how the different seeding treatments have affected vegetative cover, species richness, diversity, and floristic quality in the harsh en¬ vironment of a Michigan sand prairie. METHODS Study Area The study site is located at the historic Chittenden Nursery in the Manistee-Huron National For¬ est in Manistee County, Michigan. The nursery site is approximately 23 hectares and consists of 13 adjacent 1- to 2-hectare open fields. The location chosen is an appropriate study site, because the area was relatively homogenous inthat it was level, tree-free, and dominated by invasive native and non-native species. The rectangular study area covers an approximate 500 m 2 portion of one of these fields. The site location is depicted in Figure 1. The study site was historically part of the oak-pine barrens ecosystem which included pockets of sand prairie (Albert and Comer 2008). In 1934, the area was converted into the Chittenden Nursery, a tree nursery for the United States Forest Service (USFS). The tree nursery was shut down in the 1970s and has since been used rather insignificantly for USFS housing, conferences, and training ac¬ tivities (e.g., wildfire training, prescribed burns, all-terrain vehicle instruction). Climatic factors at the time of the study were consistent with historical averages. Summer tem¬ perature highs during the survey years averaged 23.1°C, and average monthly rainfall was 17.5 cm. Site soils are mapped as Plainfield sands with a very deep water table (United States Department of Agriculture 2012). The weather and soil conditions are consistent with oak-pine barren and sand prairie ecosystems, and the vegetation is characteristic of a degraded ecosystem (Albert and Comer 2008). Experimental Design Prior to the initiation of our experiment, the entire study area was mowed and foliar herbicide containing glyphosate was applied to it in 2007 and again in 2008 in an attempt to reduce the abun¬ dance of invasive species (e.g., Centaurea stoebe F.). In March 2009, a total of 228 1 m 2 treatment plot boundaries were established. All plots were separated by a 0.5 m border along all four sides to avoid edge effects and to prevent trampling while taking measurements. Typical sand prairie species used in local U.S. Forest Service restorations were 96 THE MICHIGAN BOTANIST Vol. 54 □□□□□□□ □□□□□□□ □□□□□□□ □□□□□□□ □□□□□□□ □□□□□□□ □□□□□□□ □□□□□□□ □□□□□□□ □□□□□□□ □□□□□□□ □□□□□□□ □□□□□□□ □□□□□□□ □□□□□□□ □□□□□□□ □□□□□□□ □□□□□□□ □□□□□□□ □ □□□□ □ □□□□ □ □□□□ □ □□□□ □ □□□□ □ □□□□ □ □□□□ □ □□□□ □ □□□□ □ □□□□ □ □□□□ □ □□□□ □ □□□□ □ □ □ □ □ □ □□□□ □ □□□□ □ □□□□ □ □□□□ □ □□□□ FIGURE 1. Chittenden Nursery, Manistee County, Michigan. Location of the approxi¬ mately 500 m 2 rectangular study site within the nursery and a diagram showing the lay¬ out of the 228 1 m 2 study plots separated by 0.5 m buffers. chosen to be planted based on their specific guild and historic presence in Michigan sand prairies. Three species per guild were selected. Selection criteria reflect a balance of conservative versus less- conservative native species. Species were further selected to include a variety of ecological charac¬ teristics, including both late vs. early season flowering species, both C4 and C3 photosynthetic path¬ ways (among the grasses), and, with respect to soil nutrient relationships, both nitrogen fixers (legumes) and nitrogen extractors. The plants selected for seeding were placed into one of five groups—legumes, early season forbs, late season forbs, grasses, and background species. The species in each of these five groups are listed in Table 1. Twelve of the treatments consisted of one of the first three functional groups (legumes, early season forbs, late season forbs), grasses, and background species, for a maximum of 13 species per treatment. Background species were selected and added to the seed mixes in order to mimic a more traditional species rich and diverse seed mixture similar to those used in sand prairie restoration activities. Each species other than the background species was seeded at either a high concentration (10,000 seeds/m 2 ) or a low concentration (1,000 seeds/m 2 ). The background species were seeded at a density of 500 seeds/m 2 . Seeding densities were equally divided for each species so that an equal number of seeds per species were represented in the mix. Seeds were obtained from Michigan Wildflower Farm, Portland, Michigan. All seeds are of local Michigan genotypes. In January 2009, seeds were counted and weighed to determine the number of seeds per gram for each species. The appropriate number of seeds were mixed with clean, moist sand and placed in a freezer for a minimum of two weeks to simulate cold-moist stratification before being sown into the field. In March 2009, all seed mixes, except those containing legumes, were sown evenly by hand into treatment plots. Due to the unavailability of seed in the spring, legume seed treatments (includ¬ ing their grass and background species components) were sown in December 2009. All seeded group comparisons and control plots were randomly assigned. Each treatment was replicated 12 times for a 2015 THE MICHIGAN BOTANIST 97 TABLE 1. Species in each of the five seeding groups and associated coefficient of conservatism (CC) values. Seeding Group Species CC Value Legumes Lupinus perennis L. 7 Desmodium canadense (L.) DC. 3 Lespedeza capitata Michx. 5 Early Season Forbs Penstemon hirsutus (L.) Willd. 5 Asclepias tuberosa L. 5 Anemone virginiana L. 3 Late Season Forbs Symphyotrichum laeve (L.) A. Love & D. Love 5 Solidago nemoralis Aiton 2 Solidago speciosa Nutt. 5 Grasses Andropogon gerardii Vitman 5 Schizachyrium scoparium (Michx.) Nash 5 Sorghastrum nutans (L.) Nash 6 Background Species Coreopsis lanceolata L. 8 Euphorbia corollata L. 4 Liatris aspera Michx. 4 Monarda fistulosa L. 2 Oenothera biennis L. 2 Rudbeckia hirta L. 1 Verbena stricta Vent. 4 total of 228 plots. The composition by functional group of each of the seeding treatments is given in Table 2. Treatments that are listed as having 24 replicates in Table 2 are those that received separate March and December seedings of 12 replicates each. Baseline Data Collection In July 2009, before the growth of any of the seed treatments, visual estimates were made of the percentage cover of vegetation for each plot. Measurements of the percentage of vegetative cover were broken down into seven different groupings in order to further explore the differences in each treatment group. These groupings were: (i) all species encountered (“total”), (ii) all non-native species encountered (“non-native”), (iii) all native species encountered (“native”), (iv) all non-seeded species encountered, both native and non-native (“resident”), (v) all seeded species encountered (“planted”), (vi) all grass species encountered (“grass”), and (vii) all forb species encountered (“forb”). Initial estimates of percentage cover for each plot were averaged across the entire 228 plot study site. Mean species richness and diversity values were derived from the percentage cover data of each plot and were averaged across the entire study site. Baseline measurement data, representative of pre¬ restoration conditions, are provided in Table 3. Floristic quality index (FQI) values of the different seed mixes were calculated (Table 4). These values may provide insight on the extent that restoration plot FQI may be predetermined by seed mix design. Species nomenclature follows The Taxonomic Name Resolution Service (2014), which is an online tool for automated standardization of plant names as described by Boyle et al. (2013). In late August 2009, a floristic quality assessment (FQA) was completed for the one-half hectare area that surrounds the study site. This area was not affected by the site-preparation herbicide treat¬ ments. The surrounding area included a mix of disturbed, developed land, mesic-mixed forest, and xeric open fields. This FQA provides baseline data for species that occur in the immediate area adja¬ cent to the study site and may provide information for future measures (e.g., seed bank germination 98 THE MICHIGAN BOTANIST Vol. 54 TABLE 2. Composition by functional groups of the 15 treatment groups. N indicates the number of replicates for each treatment. In addition to these treatments, there were 12 control plots that received no treatment. Treatment Group N Early Season Flowering Forbs 1. Early Season Forbs (high); Grasses (low); Background Species 12 2. Early Season Forbs (low); Grasses (high); Background Species 12 3. Early Season Forbs (low); Grasses (low); Background Species 12 4. Early Season Forbs (low); Background Species 12 Late Season Flowering Forbs 5. Late Season Forbs (high); Grasses (low); Background Species 12 6. Late Season Forbs (low); Grasses (high); Background Species 12 7. Late Season Forbs (low); Grasses (low); Background Species 12 8. Late Season Forbs (low); Background Species 12 Legumes 9. Legumes (high); Grasses (low); Background Species 12 10. Legumes (low); Grasses (high); Background Species 12 11. Legumes (low); Grasses (low); Background Species 12 12. Legumes (low); Background Species 12 Grasses and Background Species Only 13. Grasses (high); Background Species 24 14. Grasses (low); Background Species 24 Background Species Only 15. Background Species 24 TABLE 3. Mean pre-restoration, baseline data collec¬ tion values for percentage cover, species richness, FQI, and H’ of the entire 228-plot study site in July 2009. Variable Value Cover (%) Native 42.22 Non-Native 57.78 Resident 99.97 Planted 0.03 Grass 0.75 Forb 99.25 Total 100.00 Species Richness Total 6.98 Native 2.08 Non-Native 4.90 Resident 6.92 Planted 0.06 Grass 0.37 Forb 6.61 Diversity Floristic Quality Index (FQI) 0.99 Shannon Diversity Index (H’) 1.11 and colonization). Plants that could not be readily identified were trimmed with scissors at the base and placed into plas¬ tic bags. The bags were labeled with key plant and environmental characteristics (e.g. soil conditions, exposure to shade/sun). The plants were transported on ice back to Grand Valley State Uni¬ versity in Allendale, Michigan, where they could further be identified. The list of species collected is provided in Table 5. It is important to note that only species that were identifiable in the fall were included in the list. Many native and non-native plants, especially early season forbs and grasses, may have been excluded in this initial sampling due to their lack of key reproductive structures. Data Collection Data on the percentage cover of each species and on species richness were collected from the study plots in mid-July each year from 2009 through 2015 THE MICHIGAN BOTANIST 99 TABLE 4. Floristic quality index (FQI) of each of the seed mixes. Seed Mix FQI Early Season Flowering Forbs Early Season Flowering Forbs; Grasses; Background Species Early Season Flowering Forbs; Background Species 14.98 12.02 Late Season Flowering Forbs Fate Season Flowering Forbs; Grasses; Background Species Late Season Flowering Forbs; Background Species 14.70 11.70 Legumes Legumes; Grasses; Background Species Legumes; Background Species 15.53 12.65 Grasses and Background Species Only Grasses; Background Species 12.97 Background Species Only Background Species 9.45 Entire Seeded Site 18.58 2011. These data were used to calculate the floristic quality index (FQI) and the Shannon diversity index (H’) for each plot. If the species identification of any plant was not possible due to the plant’s early life stage, it was flagged for later identification. Plants that were ultimately unidentified were not incorporated into the data analysis of percentage cover. The percentage cover of each species, of bare-ground, and of litter percent were visually esti¬ mated in 2009 and 2010. In order to provide more objective quantitative data, percentage cover was determined in 2011 using the point-intercept method. This method involves dropping a metal pin at 50 points along an evenly spaced grid that covered the entire plot. Any vegetative part of a plant that touched the pin was counted as an occurrence of that plant at that particular point. Any piece of lit¬ ter or bare-ground that touched the pin was also accounted for. The total number of occurrences of an individual species in the plot was then divided by the total number of occurrences of all species in the plot to provide the relative percentage cover of that species. Any species noted visually as present in a plot that was not encountered by the point-intercept method was marked as being a trace amount occurrence (0.0625% cover) of the plot. In 2009, only the 144 non-legume plots were evaluated. In 2010 and 2011, all 228 plots were evaluated. In an effort to maintain consistent data collection ef¬ forts, estimates of percentage cover were calibrated between the same two researchers throughout the course of the multi-year study. This was a necessary approach, as cover estimates are subjective be¬ tween individuals. Species richness was calculated by counting the number of species that occurred in each plot. Those species for which identification was not possible due to the plant’s early life stage were in¬ cluded as part of these calculations if they were identifiable to the extent that they could be deter¬ mined to be different from species otherwise noted in the plot. A list was kept of all species identi¬ fied, and all unknowns were described in detail. Floristic quality index (FQI) measures the overall quality of an area (Swink and Wilhelm 1994) based on coefficient of conservatism (CC) values provided by the Michigan DNR (Hermann et al. 1996). Assigning CC values to individual plant species provides an approach to ranking the quality of plant communities (Swink and Wilhelm 1994). These values range from 0 to 10 and are assigned only to native species. A plant with a low value would represent a species that is common, can per¬ sist in highly degraded areas, and is likely not indicative of a high quality remnant community, for example, Solidago canadensis L., for which CC = 0, and Rudbeckia hirta L., for which CC=1. Con¬ versely, a species with a high value represents a species that would likely be found only in a place in¬ dicative of an intact remnant of a natural ecosystem, for example, Lithospermum canescens (Michx.) 100 THE MICHIGAN BOTANIST Vol. 54 TABLE 5. List of species collected at the study site prior to the growth of any of the seed treatments. Non-native species are indicated by an asterisk (*). The coefficient of conservatism (CC) value for the native species are taken from Hermann et al. (1996). Species CC Value Species CC Value Acer rubrum L. 1 Plantago rugelii Decne. 0 Ambrosia artemisiifolia L. 0 Populus tremuloides Michx. 1 Andropogon gerardii Vitman 5 Potentilla argentea L.* - Asclepias syriaca L. 1 Potentilla recta L.* - Bromus inermis Leyss. * - Pseudognaphalium obtusifolium (L.) Centaurea stoebe L.* - Hilliard & B.L. Burtt 2 Cichorium intybus L. * - Pteridium aquilinum (L.) Kuhn 0 Cirsium vulgare * (Savi) Ten.* - Robinia viscosa Vent.* - Clinopodium vulgare L. 3 Rudbeckia hirta L. 1 Comptonia peregrina (L.) J.M. Coult. 6 Sassafras albidum (Nutt.) Nees 5 Conyza canadensis (L.) Cronquist 0 Schizachyrium scoparium (Michx.) Nash 5 Daucus carota L.* - Solidago canadensis L. 1 Elaeagnus umbellata Thunb.* - Sorghastrum nutans (L.) Nash 6 Elymus repens (L.) Gould.* - Symphyotrichum laeve (L.) A. Love & Erigeron strigosus Muhl. Ex Willd. 4 D. Love 5 Fragaria virginiana Mill. 2 Symphyotrichum ontarionis (Wiegand) Holosteum umbellatum L.* - G.L. Nesom 6 Hypericum perforatum L.* - Tragopogon pratensis L.* - Juniperus virginiana L. 3 Trifolium arvense L.* - Lepidium virginicum L. 0 Triosteum perfoliatum L. 5 Leucanthemum vulgare Lam.* - Verbascum blattaria L.* - Lupinus perennis L. 7 Verbascum thapsus L.* - Melilotus alba (L.) Lam.* - Verbena hastata L. 4 Melilotus officinalis (L.) Lam.* Monarda fistulosa L. Monarda punctata L. Oenothera biennis L. Panicum capillare L. Pinus strobus L. 2 4 2 1 3 Vitis aestivalis Michx. 6 Lehm.), for which CC=10, and Ceanothus americanus L., for which CC = 9 (Swink and Wilhelm 1994, Hermann et al. 1996). FQI is calculated by Equation 1, where C is the average of CC values from native plants species in the sampled community, and N is the total number of native species in the community. Non-native plant species are not included in this calculation. Equation 1: FQI = Ca/N According to Hermann et al. (1996), an FQI greater than 50 indicates an area of high conser¬ vatism, and an area with an FQI greater than 35 is considered to be floristically high quality in Michi¬ gan. Diversity was expressed by the Shannon diversity index ( H ') (Shannon 1948), which equals (-1) times the sum for all native species in the sampled community of the product of the relative cover (pf) of each species and its natural logarithm, as shown in Equation 2. H' = -f jPi \ nPi i=i Statistical Analysis Measurements of the percentage of vegetative cover and of species richness were broken down into the previously mentioned seven different groupings: total, non-native, native, resident, planted, grass, and forb groupings. Treatments were compared by both forb and grass groupings for percentage cover, species rich- 2015 THE MICHIGAN BOTANIST 101 ness, and diversity. Grass groupings include control plots, background species only (treatment 15 in Table 2), forbs, including legumes with no grass (treatments 4, 8, and 12), high grass (treatments 2, 6, 10, 13), and low grass (treatments 1, 3, 5, 7, 9, 11, and 14). Forb groupings include control plots, background forb species only (treatments 13-15), early season flowering forbs (treatments 1-3), and late season flowering forbs (treatments 5-7), legumes (treatments 9-11), and forb treatments con¬ taining either early season flowering forbs, late season flowering forbs, or legumes that contained no grass (treatments 4, 8, and 12). Exploratory analysis for percentage cover of bare-ground and litter values revealed that they were normally distributed. In order to produce normality in the other variables, a variety of transforma¬ tions were used. The percentage cover values for the native, planted, and grass groupings were square root transformed in order to correct for their positive skew. The percentage cover values for t he non- native, resident, and forb groupings were transformed by formula, Normal Value (X) = \Ik - X. (where K is a constant from which each score is subtracted so that the smallest score is 1), to correct for their negative skew (Tabachnick and Fidell 2007). Once the data were normalized, a one-way analysis of variance (ANOVA) was used to test whether there were significant differences in per¬ centage cover between treatment groups in 2011. In order to further compare the differences among multiple comparison groups, Tukey post-hoc tests were run between all treatment groups. Compar¬ isons were considered significantly different at p < 0.05 after taking into account the Bonferroni cor¬ rection. All statistical analyses were performed using PASW 18 (SPSS Inc. 2011). Data are reported as non-transformed values. Exploratory analysis for species richness and diversity values revealed that they were normally distributed. A one-way ANOVA was used to test whether there were significant differences among treatment groups in species richness and diversity in 2011. In order to further compare the differ¬ ences among multiple comparison groups, Tukey post-hoc tests were run between all treatment groups. Comparisons were considered significantly different at p < 0.05 after taking into account the Bonferroni correction. All statistical analyses were performed using PASW 18 (SPSS Inc. 2011). An independent sample T-test was used to determine if there were significant differences in mea¬ sured variables between spring seeded treatments and fall seeded legume treatments. All treatments were statistically similar. RESULTS Baseline Floristic Quality Assessment A total of 50 different plant species were encountered during the baseline floristic quality assessment of the surrounding areas of the study site in 2009, in¬ cluding 31 native and 19 non-native species (Table 5). The average coefficient of conservatism (CC) value was 2.94, resulting in a calculated floristic quality index (FQI) of 16.37, which indicates that the area was not considered a floristi- cally high quality site (Hermann et al. 1996). A list of resident species encoun¬ tered within the study plots by year is provided in Table 6. Baseline vegetative cover data indicate that the most prevalent species were Conyza canadensis (L.) Cronquist (mean = 41%), Centaurea stoebe L. (mean = 23%), and Potentilla argentea L. (mean = 22%). The most prevalent grass species were Bromus inermis Leyss. (mean = 1%) and Agrostis hyemalis (Wal¬ ter) Britton, Sterns & Poggenb. (mean = 1%). Comparison among Grass Treatment Groups after Three Years Percentage Cover The data reported in this section are summarized in Table 7. Significant dif¬ ferences (p<0.050) were found among treatment groups in the percentage cover TABLE 6. Resident species encountered within the study site in each of the years of the study. Non-native species are indicated by an asterisk (*). 2009 2010 2011 Acer rubrum L. Acer rubrum L. Asclepias syriaca L. Agrostis hyemalis (Walter) Brit., Sterns & Pogg. Agrostis hyemalis (Walter) Brit., Stems & Pogg. Acer rubrum L. Arenaria serpyllifolia L.* Arenaria serpyllifolia L.* Agrostis hyemalis (Walter) Brit., Sterns & Pogg. 102 THE MICHIGAN BOTANIST Vol. 54 & 11 f* o I 1 * Uo 2 _• 2 2 a S J s ^ S B 2 * Q . i : J & g § i g S ■ K i !d’-~ h-I a ^ ^ k 1-^'S 5 i 11 Hi faq £ ^5 tt| .Still s & ^ § W g s ^ -S g a .2 3 2 81 ^ a ,2 ||11 ill IS iBjSSfi ^ Oq cq o o O Q S S 3 '3 ’ 111! « s > l S §3 J _} : s S •5 .W s s -s 53 J a . cs ~ ► J S o -2 ^ r^* et i *«5 ^ >0 •+-* i-* .£> S u s s S i eg § c K R & K W W 2 te te g o a g £0 g g s O & !S 2 53 •B# sf B o § ^ M-H 3 8; ■ o Q : . 2 tilth'll C3, S 3 "53 § 2 R-S &• e W s- i Sjf 5 2 g tig §!“*■ ■go ° -» S > s u t o K * S i =q O ^ 53 o u q ,1|S| =1:^11 Eg eg tg t till I stew?* 11 if£ tq aq ^ kj O 2015 THE MICHIGAN BOTANIST 103 of the bare-ground (F 4i22 y=5. 79,/?<0.001), native {F 4>22 7 =ll -20, /><0.001), non¬ native (F 4 j 27 = 11.41, /?<0.001), resident (F 4227 =2 4.09, /?<0.001), planted (F 4 227=23.90, /7 <0.001), grass (F^227=30.39, p=0.0 14), and forb (F 4227 =30.39, p= 0.014) groupings. There were no significant differences (p>0.050) among treatment groups in litter cover (F 4 227 =0.70, p=0.595, mean = 17.86%). Bare-ground cover averaged 36.03% across all grass treatment groups. High grass treatments had significantly lower bare-ground cover (mean = 31.85%, p< 0.050) than all other treatment groups except the control plots (mean = 37.20%), which showed high variability. Native cover averaged 19.34% across all grass treatment groups. High grass treatments had significantly higher native cover (mean = 35.02%,/?<0.050) than all other treatment groups. Low grass treatments had significantly higher native cover (mean = 22.99%, p<0. 050) than all other treatment groups except the high grass treatments. There were no significant differences in native cover among all other treatment groups (mean range: 8.01-16.94%) Comparison of mean native cover in each of these treatment groups is shown in Figure 2. Non-native cover averaged 80.54% across all grass treatment groups. High grass treatments had significantly lower non-native cover (mean = 64.38%, p< 0.050) than all other treatment groups, while low grass treatments had signif¬ icantly lower cover (mean = 77.01%, /?<0.050) than all other treatment groups except the high grass treatment. There were no significant differences in non-na- Only (No Treatments Grass) (No Grass) Native Species Relative % Cover (±1 SE) FIGURE 2. Grass Treatment Comparison: Mean relative percentage cover of species in the native grouping for several treatment groups in 2011. Bars that do not share the same letter are significantly different (p< 05) as determined by Tukey post-hoc analysis. Error bars represent one standard error about the mean. 104 THE MICHIGAN BOTANIST Vol. 54 100 90 80 70 60 50 40 30 20 10 0 a r-4—| ab ab ii b -1- -1- -1- 1 Control Background Forb Treatments High Grass Only (No Grass) (No Grass) Low Grass Non-Native Species Relative % Cover (±1 SE) FIGURE 3. Grass Treatment Comparison: Mean relative percentage cover of species in the Non-Na¬ tive grouping for several treatment groups in 2011. Bars that do not share the same letter are signif¬ icantly different (p<.05) as determined by Tukey post-hoc analysis. Error bars represent one standard error about the mean. tive cover among all other treatment groups (mean range: 83.06-91.99%). Com¬ parison of mean non-native cover in each of these treatment groups is shown in Figure 3. Resident cover averaged 86.45% across all grass treatment groups. Resident cover in control plots (mean = 99.83%) was significantly greater (/?<0.050) than all other treatments (mean range: 68.00-91.12%). High grass treatments had sig¬ nificantly lower resident cover (mean = 68.00%, p <0.050) than all other treat¬ ments. Low grass treatments had significantly lower (mean = 82.71%,/?<0.050) resident cover than the treatment with background only species and no grass treatment and the control plots. Planted cover averaged 13.43% across all grass treatment groups. Control plots had significantly less planted cover (mean = 0.17%,/?<0.050) than all other treatment groups (mean range: 8.88-31.40%). High grass treatments had signif¬ icantly higher planted cover (mean = 31.40%, /? <0.050) than all other treatment groups. Low grass treatments had significantly higher planted cover (mean = 17.29%,/?<0.050) than the background only treatments with no grass treatments (mean = 8.88%), but was not significantly different (/?>0.050) from the forb treatments with no grass (mean = 9.43%). Grass cover averaged 10.35% across all grass treatment groups. High grass treatments had significantly more grass cover (mean = 28.47%, p< 0.050) than all other treatment groups. Low grass treatments had significantly higher grass TABLE 7. Mean percentage cover values of bare-ground and litter and of the native, non-native, resident, planted, grass, and forb groupings in the grass treat¬ ment groups in 2011. Percentage cover in each grouping was tested only against the treatments within the same grouping. Values within a column that do not share a superscript letter are significantly different (p<0.05) as determined by Tukey post-hoc analysis. Treatment Group Bare-ground Litter Native Non-Native Resident Planted Grass Forb 2015 THE MICHIGAN BOTANIST 105 m "O In 3 a O Cm s . — °5 ■X £ s = PQ O o Tf cb h -H -H -H i oog cb f>J -H & ® ON £ ^ g •2 > is cci eg P *? a § I M -■r a> &o g H O .2 ^ Ph v CD ^ rt CO « 1) p 6 o 13 -p p g jD 55 iZ) <0.001), and forb (7^227=5.32, /?<0.001) groupings. There were no significant differences (£>>0.050) in species richness among grass treatment groups in the non-native (F 4 2 27 = Q-2.6, p=0.905, mean = 5.33 species) and resident (F 4 2 27 = 0.2l, p=0.932, mean = 7.27 species) groupings. The total species richness averaged 10.73 species across all grass treatment groups. It averaged significantly lower for the control plots (mean = 7.75 species, /?<0.050) than for all forb treatments with no grass (mean = 10.61 species), high grass treatments (mean = 12.88 species), and low grass treatments (mean = 12.49 species). High grass and low grass treatments exhibited signifi¬ cantly higher total species richness than all other grass treatment groups (p<0.050). Comparison of mean total species richness in each of these treatment groups is shown in Figure 4. Species richness of the native grouping averaged 5.35 species across all grass treatment groups. The control plots had significantly lower native species rich¬ ness (mean = 2.33 species, /?<0.050) than all other treatment groups (mean range: 4.63-7.48 species). High grass and low grass treatments had significantly higher native species richness (mean = 7.48 and 6.96 species, respectively, £><0.050) than all other treatment groups. Comparison of mean native species richness in each of these treatment groups is shown in Figure 5. Species richness of the planted grouping averaged 3.41 species across all grass treatment groups. All treatment groups had significantly higher planted species richness (mean range: 2.79-5.60 species,/?<0.050) than the control plots (mean = 0.25 species). High grass and low grass treatments had significantly higher planted species richness (mean = 5.60 and 5.03 species, respectively, £><0.050) than all other treatment groups. Species richness of the grass grouping averaged 1.73 species across all grass treatment groups. High grass treatments had significantly higher grass species richness (mean = 3.50 species, /?<0.050) than all other treatments. Low grass treatments had significantly higher grass species richness (mean = 2.63 species, £><0.050) than all other treatments. Species richness of the forb groupings averaged 9.01 species across all grass treatment groups. All treatment groups had significantly higher forb species richness (mean range: 9.17-9.86 species, p<0. 050) than the control plots (mean = 6.92 species). 2015 THE MICHIGAN BOTANIST 107 c ft— fr 7^ a Ettt be rri b jij w U\. Control Background Forb Treatments High Grass Low Grass Only (No Grass) (No Grass) Total Species Richness (±1 SE) FIGURE 4. Grass Treatment Comparison: Mean species richness encountered in the Total grouping for several treatment groups in 2011. Bars that do not share the same letter are significantly different (p<.05) as determined by Tukey post-hoc analysis. Error bars represent one standard error about the mean. Only (No Grass) (No Grass) Native Species Richness (±1 SE) FIGURE 5. Grass Treatment Comparison: Mean species richness encountered in the Native grouping for several treatment groups in 2011. Bars that do not share the same letter are significantly different (p<.05) as determined by Tukey post-hoc analysis. Error bars represent one standard error about the mean. 108 THE MICHIGAN BOTANIST Vol. 54 TABLE 9. Comparison of the mean floristic quality index (FQI) and Shannon diversity index (H’) values of the grass treatment groups and selected non-grass treatment groups in 2011. Each different measure was tested independently. Values within a column that do not share a superscript letter are significantly different (p<0.05) as determined by Tukey post-hoc analysis. Treatment Group FQI H' Control Plots 1.37(±0.56) 1.26 b (±0.09) Background Only, No Grass 6.56 a (±0.45) 1.40 ab (±0.06) Forb Treatment, No Grass 6.71 a (±0.39) 1.34 b (±0.06) High Grass 10.37 c (±0.17) 1.61 a (±0.04) Low Grass 9.35 b (±0.18) 1.49 ab (±0.04) Floristic Quality and Diversity There were significant differences among treatment groups in FQI values (F 4 227=87.83, /?<0.001). The mean FQI of grass treatment groups was 6.87. The FQI of the control plots in 2011 was significantly lower (mean = 1.37) than that of all other treatments (mean range: 6.56-10.37). The high grass treatments had a significantly higher FQI (mean = 10.37) than all other treatment groups while low grass treatments had a significantly higher FQI (mean = 9.35) than all other treatments except the high grass treatments (Table 9; Figure 6). 12 10 Control Background Forb Treatments High Grass Low Grass Only (No Grass) (No Grass) Floristic Quality Index (FQI) (±1 SE) FIGURE 6. Grass Treatment Comparison: Mean FQI for several treatment groups in 2011. Bars that do not share the same letter are significantly different (p<.05) as determined by Tukey post-hoc analysis. Error bars represent one standard error about the mean. 2015 THE MICHIGAN BOTANIST 109 1.8 1.6 1.4 1.2 1 0.8 0.6 0.4 0.2 0 ab ab Control Background Forb Treatments High Grass Only (No Grass) (No Grass) Low Grass Shannon Diversity (H’) (±1 SE) FIGURE 7. Grass Treatment Comparison: Mean H’ for several treatment groups in 2011. Bars that do not share the same letter are significantly different (p<.05) as determined by Tukey post-hoc analysis. Error bars represent one standard error about the mean. There were also significant differences among treatment groups in the Shan¬ non diversity index (H’) {F 42 2f = 4-43, p= 0.002) (Table 9). The mean H’ of the grass treatment groups was 1.42. The mean H' of the high grass treatments was significantly higher (mean = 1.61) than for all other treatments except the back¬ ground species only with no grass treatment (mean = 1.40) and the low grass treatment (mean = 1.49) (Table 9; Figure 7). Comparison among Forb Treatment Groups after Three Years Percentage Cover The data reported in this section are summarized in Table 10. Significant dif¬ ferences (p<0.050) were found among treatment groups in the percentage cover of the bare-ground (F 4 2 27=2-99, p=0.020), native (F 4 2 2^S.75, p<0.00l), non¬ native (F 42 27 = 8.33, /?<0.001), resident (F 4t2 27 = ^7 .72, /?<0.001), planted (^,227=18.51, p<0.00\), grass 227=3 .21, jp=0.014), and forb (F 4 j$t= 3.21, p= 0.014) groupings. There were no significant differences (/?>0.050) among treatment groups in litter cover {F 4 227=0-16, p=9.959, mean = 18.04%). Bare-ground cover averaged 35.59% among forb treatment groups. The late season flowering forb treatment had significantly less (/?<0.050) bare-ground cover than the legume treatments (mean = 33.40% and 37.96%, respectively). Native cover averaged 21.24% among forb treatment groups. Treatments of early season flowering forbs and late season flowering forbs showed significantly 110 THE MICHIGAN BOTANIST Vol. 54 35 30 25 20 15 10 5 0 Control Background Legumes Early Season Late Season Only Native Relative % Cover (±1 SE) b c ab j j , ■ 1 ■ a a j:j FIGURE 8. Forb Treatment Comparison: Mean relative percentage cover of species in the Native grouping for several treatment groups in 2011. Bars that do not share the same letter are significantly different (p<05) as determined by Tulcey post-hoc analysis. Error bars represent one standard error about the mean. higher percent native cover (mean = 34.10% and 27.19%, respectively, p <0.050) than the control plots (mean = 8.01%) and legume treatments (mean = 14.68%). Comparison of mean native cover in each of these treatment groups is shown in Figure 8. Non-native cover averaged 78.61% among forb treatment groups. The control plots and legume treatments showed significantly higher non-native cover (mean = 91.99% and 84.57%, respectively, /?<0.050) than early season and late season flowering forb treatments (mean = 65.90%, 72.81%, respectively). Comparison of mean non-native cover in each of these treatment groups is shown in Figure 9. Resident cover averaged 84.39% among forb treatment groups. Resident cover in control plots (mean = 99.83%) was significantly greater (/?<0.050) than in all other treatments (mean range: 71.16%-90.96%). Early season flowering forbs and late season flowering forbs (mean = 71.16% and 76.46%, respectively) had significantly lower (/?<0.050) resident cover than all other treatments. Planted cover averaged 15.46% among forb treatment groups. Control plots had significantly less planted cover (mean = 0.17%, p<0. 050) than all other treatment groups (mean range: 8.29%-28.84%). Early season flowering forb and late season flowering forb treatments (mean = 28.84% and 23.54%, respectively) had significantly more (p <0.050) planted cover than all other treatment groups. Grass cover averaged 11.93% among forb treatment groups. All treatment groups had significantly higher grass relative cover (mean range: 2015 THE MICHIGAN BOTANIST 111 100 90 80 70 60 50 40 30 20 10 0 ab 5 be Control Background Only Legumes Early Season Late Season Non-Native Relative % Cover (±1 SE) FIGURE 9. Forb Treatment Comparison: Mean relative percentage cover of species in the Non-Na¬ tive grouping for several treatment groups in 2011. Bars that do not share the same letter are signif¬ icantly different (p<.05) as determined by Tukey post-hoc analysis. Error bars represent one standard error about the mean. 8.30%-17.48%,/?<0.050) than the control plots (mean = 3.32%). Early season flowering forbs and late season flowering forbs had significantly higher grass cover (mean = 17.48% and 16.18%, respectively, /?<0.050) than all other treat¬ ment groups except the background only treatments (mean = 14.36%. Forb cover averaged 88.07% among forb treatment groups. All treatment groups had significantly lower (p<0.050) forb (mean range: 82.52%-91.70%) relative cover than the control plots (mean = 96.68%). Early season flowering forbs and late season flowering forbs had significantly lower forb cover (mean = 82.52% and 83.82%, respectively) than all other treatment groups except the background only treatments (mean = 85.64%. Species Richness The data reported in this section are summarized in Table 11. There were sig¬ nificant differences (p<0.050) in species among treatment groups in the total (F 422 7=9.99, p<0.00\), native (F 4 j 2 2=15.77, /?<0.001), planted (F 4>2 27 = 26. 28, /?<0.001), grass (F 4 ^27=3.98,/?=0.004), and forb (F 4 ,227=9.96, /?<0.001) group¬ ings. There were no significant differences (p >0.050) in species richness among forb treatment groups in the non-native (F 42 27 = 0. 65, p= 0.629, mean = 5.35 species) and resident (F 4 ,227=0-38, p=0. 825, mean = 7.29 species) groupings. The total species richness averaged 11.20 species among forb treatment groups. It averaged significantly less for the control plots (mean = 7.75 species, TABLE 10. Mean percentage cover values of bare-ground and litter and of the native, non-native, resident, planted, grass, and forb groupings in the grass treat¬ ment groups in 2011. Percentage cover in each grouping was tested only against the treatments within the same grouping. Values within a column that do not share a superscript letter are significantly different (p<0.05) as determined by Tukey post-hoc analysis. 112 THE MICHIGAN BOTANIST Vol. 54 O H O A -H 22* Cm 9 a -5 20 5 « u U pq fe o ^ oo q q m m HH ■ mm z °°; ca jd jd ON O m oo q cn m Tj iLif if a o £ t~- O ^ in Os oo A in fsj oo no q 55 c5 —1 A A bp bo . £ 13 5 o > Tj G 05 ^ £ 50 85 © © A n ^1 o O -H in m r- in A A 22 Cm s __ o U CQ Legumes 11.83 ab (±0.49) 6.46 a (±0.43) 5.35(±0.23) 7.21(±0.29) 4.60 a (±0.32) 2.06 ab (±0.22) 9.77 ab (±0.38) Early Season 13.25 b (±0.44) 8.06 b (±0.37) 5.15(±0.19) 7.29(±0.27) 5.92 b (±0.29) 2.56 b (±0.22) 10.71 a (±0.33) Late Season 11.63 a (±0.35) 6.17 a (±0.30) 5.42(±0.23) 7.13(±0.26) 4.46 a (±0.20) 2.40 b (±0.19) 9.23 b (±0.26) 2015 THE MICHIGAN BOTANIST 113 16 14 12 10 8 6 4 2 Control Background Only Legumes Early Season Late Season Total Species Richness (±1 SE) FIGURE 10. Forb Treatment Comparison: Mean species richness encountered in the Total grouping for several treatment groups in 2011. Bars that do not share the same letter are significantly different (p<.05) as determined by Tukey post-hoc analysis. Error bars represent one standard error about the mean. /?<0.050) than for all other treatments. The early season flowering forb treatment had significantly higher total species richness (mean = 13.25 species, p <0.050) than all other treatments except the legumes (mean = 11.83 species). Compari¬ son of mean total species richness in each of these treatment groups is shown in Figure 10. Species richness of the native grouping averaged 5.79 species among forb treatment groups. The control plots had significantly lower native species rich¬ ness (mean = 2.33 species, /?<0.050) than all other treatment groups. The early season flowering forb treatment group had significantly higher native species richness (mean = 8.06 species, p<0.050) than all other treatment groups (mean range: 5.93-6.46 species). Comparison of mean native species richness in each of these treatment groups is shown in Figure 11. Species richness of the planted grouping averaged 3.85 species among forb treatment groups. All treatment groups had significantly higher (/?<0.050) planted species richness than the control plots (mean = 0.25 species). The early season flowering forbs had significantly greater planted species richness (mean = 5.92 species,/?<0.050) than all other treatment groups (mean range: 4.00^1.60 species). Species richness of the grass grouping averaged 2.06 species among forb treatment groups. All treatments except for legumes (mean = 2.06 species) ex¬ hibited significantly higher grass species richness (mean range: 2.40-2.56 species,/?<0.050) than the control plots (mean = 0.83 species). 114 THE MICHIGAN BOTANIST Vol. 54 9 8 7 6 5 4 3 2 1 0 Control Background Legumes Early Season Late Season Only a a h c ;:;: bu T Native Species Richness (±1 SE) FIGURE 11. Forb Treatment Comparison: Mean species richness encountered in the Native group¬ ing for several treatment groups in 2011. Bars that do not share the same letter are significantly dif¬ ferent (p<.05) as determined by Tukey post-hoc analysis. Error bars represent one standard error about the mean. Species richness of the forb grouping averaged 9.14 species among forb treat¬ ment groups. The control plots had significantly lower forb species ric hn ess (mean = 6.92 species, p<0.050) than all forb treatment groups (mean range: 9.08-10.71 species). The early season flowering forbs had significantly higher (mean = 10.71 species, p<0.050) forb species richness than all other treatment groups except for legumes (mean = 9.77 species). Floristic Quality and Diversity There were significant differences among treatment groups in FQI values (^22T=37.96,/><0.001). The mean FQI of forb treatment groups was 7.43. The FQI of the control plots in 2011 was significantly lower (mean = 1.37, p<0. 050) than that of all forb treatments (mean range: 8.39-9.99). The early season flow¬ ering forb treatments had a significantly higher FQI (mean = 9.99, p<0.050) than all other treatments except the legumes (mean = 8.89) (Table 12; Figure 12). There were also significant differences among treatment groups in the Shan¬ non diversity index (H') (F 42 27=4.86, p=0.001) (Table 12). The mean H' of forb treatment groups was 1.44. The mean H' of the control plots and legume treat¬ ments (mean = 1.26, 1.33, respectively) was significantly lower (p<0.050) than for the early season flowering forb treatments (mean = 1.62) (Table 12; Figure 13). 2015 THE MICHIGAN BOTANIST 115 Only Floristic Quality Index (FQI) (±1 SE) FIGURE 12. Forb Treatment Comparison: Mean FQI for several treatment groups in 2011. Bars that do not share the same letter are significantly different (p<.05) as determined by Tukey post-hoc analysis. Error bars represent one standard error about the mean. TABLE 12. Comparison of the mean floristic quality index (FQI) and Shannon diversity index (H') values of the forb treatment groups in 2011. Each different measure was tested independently. Values within a column that do not share a superscript letter are signifi¬ cantly different (p<0.05) as determined by Tukey post-hoc analysis. Treatment Group FQI H' Control Plots 1.37(±0.56) 1.26 a (±0.09) Background Forbs Only 8.39 b (±0.23) 1.50 ab (±0.04) Legumes 8.89 ab (±0.40) 1.33 a (±0.06) Early Season 9.99 a (±0.31) 1.62 b (±0.05) Late Season 8.51 b (±0.27) 1.50 ab (±0.05) DISCUSSION Comparison Among Grass Treatment Groups The percentage cover of resident species was significantly lower in all grass seeded treatments (i.e., no grass, low grass, and high grass), which was likely due to a change in competitive balance (Connell 1983, Fargione and Tilman 2006). The percentage cover of the planted grouping, the species richness of the native and planted groupings, and the FQI were all significantly higher in all the 116 THE MICHIGAN BOTANIST Vol. 54 1.8 1.6 1.4 1.2 1 0.8 0.6 0.4 0.2 0 ab ab Control Background Only Legumes Early Season Late Season Shannon Diversity (H’) (±1 SE) FIGURE 13. Forb Treatment Comparison: Mean H’ for several treatment groups in 2011. Bars that do not share the same letter are significantly different (p<.05) as determined by Tukey post-hoc analysis. Error bars represent one standard error about the mean. seeded grass plots than in the control plotsm which is due, in part, to an increase in forb species richness in the grass treatments. The baseline data (Table 3) indi¬ cates that the community was heavily forb-dominated in 2009. However, native forb species included in the grass treatments further increased the species rich¬ ness of forbs. Although some grass treatments included forbs, the grass-only treatments also had significantly greater forb species richness than the control plots. Forb species may be benefiting from seeded treatments, because as seeded species grow, they may facilitate the growth of other seeded species around them (Callaway and Walker 1997, Peltzer and Kochy 2001). This may also apply to non-seeded species in the surrounding area. Species currently on or near the site, as well as those species whose seeds are currently dormant in the seed bank, may realize more favorable conditions for germination from the establishment of seeded species, and thereby increase the species richness of forbs. Comparison Among Forb Treatment Groups The percentage cover of resident and forb species was significantly higher in the control plots than in the seeded forb treatments. Conversely, the percentage cover of the planted grouping, the species richness of the total, native, planted, and forb groupings, and the FQI were all significantly higher in all seeded forb treatments than in the control plots. Resident cover may be higher in control plots due to reduced competition for available open niche space resulting from 2015 THE MICHIGAN BOTANIST 117 the lack of seed additions in these plots. In contrast to the grass-only treatments, forb cover was higher in treatments that included the addition of forbs. Similarly, planted cover was higher in seeded plots than in the control plots due to the ad¬ dition of native seed. The species richness of the total, native, and planted group¬ ings and the FQI all increased with the addition of seeded forbs. Community Responses to Grass Seeding and Forb Seeding Efforts Treatments with either high or low concentrations of grasses were not signif¬ icantly different from each other in any measure of species richness other than grass species richness. High grass treatments had significantly higher grass species richness than both low and no grass treatments. Both low and high grass treatments had significantly more total and planted species richness than treat¬ ments with no grass. High grass treatments were more effective in covering more ground and hence had lower bare-ground cover than all other treatments except the highly variable control plots. High grass treatments also had a significantly higher FQI. Diversity (H') did not differ significantly between high and low grass treatments. However, these treatments were significantly more diverse than all other treatments except for the background only species with no grass treat¬ ment. Although the cover was dominated by grasses, the high grass plots displayed significantly higher native and planted cover than any other treatment. This may be attributed to the inclusion of native species within these planting groups. However, high grass only plots consisting of no planted forb species still exhib¬ ited significantly higher native and planted cover than treatments that included forbs and low grass concentrations. High grass treatments also had significantly lower non-native and resident cover than all other treatments. Native and planted covers were significantly greater in the early season forb and late season forb treatments than in all other forb treatments. Conversely, non¬ native and resident covers were significantly less in early season and late season forb treatments than in all other treatments. Early season forb treatments had significantly higher native and planted species richness than all other treatments. Although early season forb treatments had a significantly higher FQI than most other treatments, the FQI of legume treatments were not significantly different. Early season forb treatments also were significantly more diverse (FT) than most other treatments except the back¬ ground only and late season forb treatments. Facilitation These results suggest that seed additions may facilitate plant community de¬ velopment in the relatively hot and dry conditions of a sand prairie during the first few years of community development. The increased growth of native species within the grass treatments—especially within high grass seed treat¬ ments—as compared with treatments that did not contain grass suggests that seeded grasses may provide suitable microclimatic conditions for native seed 118 THE MICHIGAN BOTANIST Vol. 54 growth, such as shade for reduced heat stress and moisture capture (Plumb-Men- tjes 1990; Smith and Huston 1990; Peltzer and Kochy 2001). Similarly, the growth of early season forbs may create greater cover and biomass early in the year which could result in greater shade and moisture capture for late season forb and grass species (Henderson et al. 1988; Brooker et al. 2007; Callaway and Walker 1997). Similar results indicating a facilitative effect from seed additions have been shown in an experiment where bunchgrasses protected rare plant species during dry years (Greenlee and Callaway 1996) and in other studies where adult plants provided a facilitative response during restoration in arid ecosystems (Maestre et al. 2001; Barchuk et al. 2005). Studies have also attempted to model light and moisture capture of species in dry, harsh environments such as sand prairies. Holmgren et al. (1997) and Smith and Huston (1990) studied the role facilitation plays in community development by modeling the positive and negative effects that “nurse plant” or mature plant canopy cover plays on the establishment of new plants in dry communities. This model shows that the increased availability of water due to facilitation outweighs the detrimental effect of shade (decreased light and photosynthesis) in harsh conditions. Therefore, facilitative effects may outweigh competitive effects of neighboring plants in such environments. Competition These results also indicate that competition resulting from seed additions may play a key role in promoting initial community development. Individual resident species respond in different ways to species additions based on their ability to compete for limited resources (Grime 1974, Goldberg and Barton 1992, Smith et al. 1999). However, the overall decrease in resident species growth within the grass treatments—especially within high grass seed treatments—suggests that resident species are at a competitive disadvantage compared to seeded species. Conversely, the overall increased growth of seeded species in the grass treat¬ ments suggests that species that are native to the sand prairie may be better able to compete for the limited resources than other early successional, weedy, or non-native resident species such as Centaurea stoebe L. Similar results indicating a competitive effect from grass seed additions have been shown by Jordan et al. (2008), where they found that native grasses were less affected by invasive species. That study found that grasses are relatively in¬ sensitive to altered soil biota from invasive plants, and in turn, may shift the competitive balance of restoration efforts on an ecosystem. Similar studies sug¬ gest that grasses in prairie restoration enhance overall diversity and reduce exotic species cover (Middleton et al. 2009; Carter and Blair 2012). These studies also indicate that native grasses in restoration may represent an effective management strategy to reduce exotic plant density. Although grass seeding in restorations has been shown to be beneficial, long-term studies have shown that grasses can dom¬ inate and exclude other native species over time (Schramm 1990). These results also indicate that forb additions—particularly a benefit of early season forb treatments and a lack of response from legume treatments— may play a key role in influencing community competitive balance (Howe 2015 THE MICHIGAN BOTANIST 119 1994; Dukes 2001). Additions of early season forb species may provide an ini¬ tial, early-in-the-year increase in species cover and richness that may in turn directly affect the competitive balance of a community by providing greater re¬ sistance or buffer capability to further invasions by invasive species, which in turn may show increased diversity from other native seeded species later in the growing season (Naeem and Wright 2003; Kennedy et al. 2002; Dukes 2001). Contrary to early season forb treatments, legume treatments, although ex¬ pected to be highly competitive in a nutrient-poor sand prairie ecosystem due to their ability to fix nitrogen, had little effect on plant community composi¬ tion. Three competitive reasons may explain why we did not see significant es¬ tablishment of seeded legume species. The dominant and widespread resident legume species Trifolium arvense L. may have adverse impacts on seeded legumes. It may suppress seed additions either via interspecific competition, being temporally competitive due to its annual life cycle and producing large amounts of seeds both in the spring and summer, or possibly because nitrogen fixing niche space is already being consumed by Trifolium arvense L. (Dukes 2001; Fargione et al. 2003). Similar experimental results have suggested that seeding of early season forbs within a restoration may provide positive competitive effects to a community. Martin et al. (2005) suggested that restoration seeding efforts that contain early season forbs may be more diverse because these species are better able to co¬ exist with other, later growing seeded species because they come to occupy an early season niche. Thus, early season forbs will ultimately increase diversity, species richness, and other important community variables. Similarly, a study by Foster and Tilman (2003) suggests that early season forb seeding (as part of a complete restoration seed mix) presents an opportunity for transient coexistence through competition-colonization trade-offs, thus allowing many species to co¬ exist at the community scale. Additionally, Seabloom et al. (2003) proposed that a single seeding of native forbs (including early season forb species), even in the presence of high densities of exotic species, may be sufficient to create viable populations of native species in areas that are currently dominated by exotic species. Similar results showing a lack of legume establishment after seed additions have been found. In a twenty-five year study of prairie establishment following restoration, Schramm (1990) found that legumes did not become a major com¬ ponent of restorations until later phases of restoration succession. However, after legumes became more established in later years of restoration, it was found that they had greater staying power. This could be due to a lack of com¬ petitive ability during their early establishment periods (Schramm 1990). This may also be due to interspecific competition for limiting resources among legumes with neighboring plants belonging to the same functional group (Nemec et al. 2013). Most studies that have shown minimal impacts of legumes (species cover and richness) on restorations have been in response to herbivory. Restoration attempts that were not excluded by fencing and had populations of deer or voles saw high mortality or reduced fruiting in legume species due to the grazing preference of these animals for these species (Diaz et al. 2003; Howe et al. 2002). 120 THE MICHIGAN BOTANIST Vol. 54 CONCLUSION In general, a comparison between seeded treatments and non-seeded control treatments indicates that our efforts have been more successful in the restoration of native sand prairie than would have resulted from succession alone. Restora¬ tion attempts displayed a significant decrease in invasive, resident species rich¬ ness and increased diversity compared to succession. Treatments that included a high concentration of grass and/or an early season forb component had the great¬ est overall positive impact on plant community development. These treatments exhibited significantly greater diversity and higher floristic quality than most other treatments, and they also displayed less non-native or invasive cover than most other treatments. High concentrations of grasses and early season forbs may therefore play a critical role in initial species establishment of a sand prairie restoration due to the facilitative and competitive advantages they may provide in these harsh environments. However, it remains to be seen if these initially suc¬ cessful communities will have continued success over longer periods of time. ACKNOWLEDGMENTS This article is based on the graduate research and associated master’s program thesis completed as part of the first author’s graduate requirements at Grand Valley State University, Allendale, Michi¬ gan. The thesis upon which this article is based can be found at Roos (2014). He thanks his Gradu¬ ate Committee for their support and advice: Todd A. Aschenbach (the second author of this article and chair of the committee), Tim Evans, of Grand Valley State University, and David Warners, of Calvin College, Grand Rapids, Michigan. LITERATURE CITED Albert, D. A. (1995). Regional landscape ecosystems of Michigan, Minnesota, and Wisconsin: A working map and classification. Gen. Tech. Rep. NC-178. St. Paul: USDA, Forest Service, North Central Forest Experiment Station, St. Paul, Minnesota. Available at http://nr.fs.fed.us/pubs/242 (Version 03UN1998). Albert, D. A., and P. J. Comer. (2008). Atlas of early Michigan’s forests, grasslands, and wetlands. Michigan State University Press, East Lansing. Anderson, R. C. (1990). The historic role of fire in the North American grassland. Pp. 8-18 in Fire in North American tallgrass prairies. S. L. Collins and L. L. Wallace, editors. University of Okla¬ homa Press, Norman. Barchuck, A. H., A. Valiente-Banuet, and M. P. Diaz. (2005). Effect of shrubs and seasonal variabil¬ ity of rainfall on the establishment of Aspidosperma quebrachoblanco in two edaphically contrast¬ ing environments. Austral Ecology 30: 695-705. Belyea, L. R., and J. Lancaster. (1999). Assembly rules within a contingent ecology. Oikos 86: 402-416. Bowles, M. L., M. D. Jones, and J. L. McBride. (2003). Twenty-year changes in burned and unburned sand prairie re mn ants in northwestern Illinois and implications for management. The American Midland Naturalist 149: 35^45. Boyle, B., et al. 2013. The Taxonomic Name Resolution Service: An Online Tool for Automated Standardization of Plant Names. BMC Bioinformatics 14:16. Brooker, R. W., F. T. Maestre, R. M. Callaway, C. L. Lortie, L. A. Cavieres, G. Kunstler, P. Liancourt, et al. (2007). Facilitation in plant co mm unities: The past, the present, and the future. Journal of Ecology 96: 18-34. Cairns, J., and J. R. Heckman. (1996). Restoration ecology: The state of an emerging field. Annual Review of Energy and the Environment 21: 167-189. 2015 THE MICHIGAN BOTANIST 121 Callaway, R. M. and L. R. Walker. (1997). Competition and facilitation: A synthetic approach to in¬ teractions in plant communities. Ecology 78: 1958-1965. Carter, D. L. and J. M. Blair. (2012). High richness and dense seeding enhance grassland restoration establishment but have little effect on drought response. Ecological Applications 22: 1308-1319. Cole, K. L., and R. S. Taylor. (1995). Past and current trends of change in a dune prairie/oak savanna reconstructed through a multiple-scale history. Journal of Vegetation Science 6: 399^410. Collins, S. L., S. M. Glenn, and D. J. Gibson. 1995. Experimental analysis of intermediate distur¬ bance and initial floristic composition: Decoupling cause and effect. Ecology 76: 486-492. Collins, S. L. 2000. Disturbance frequency and community stability in native tallgrass prairie. The American Naturalist 155: 311-325. Comer, P. J., D. A. Albert, H. A. Wells, B. L. Hart, J. B. Raab, D. L. Price, D. M. Kashian, R. A. et al. (1995). Michigan’s native landscape as interpreted from the General Land Office surveys 1816-1856. Michigan Natural Features Inventory, Lansing. Report n. 1995-07. Connell, J. H. (1983). On the prevalence and relative importance of interspecific competition: Evi¬ dence from field experiments. The American Naturalist 122: 661-696. Cully, A. C, J. F. Cully, and R. D. Hiebert. (2003). Invasion of exotic plant species in tallgrass prairie fragments. Conservation Biology 17: 990-998. Diaz, S., A. J. Symstad, F. S. Chapin III, D. A. Wardle, and L. F. Huenneke. (2003). Functional di¬ versity revealed by removal experiments. Trends in Ecology and Evolution 18: 140-146. Dukes, J. S. (2001). Biodiversity and invisibility in grassland microcosms. Oecologia 126: 563-568. Emery, S. M., P. J. Doran, J. T. Legge, M. Kleitch, and S. Howard. (2013). Aboveground and below¬ ground impacts following removal of the invasive species Baby’s Breath (Gypsophila paniculata ) on Lake Michigan Sand Dunes. Restoration Ecology 21: 506-514. Fargione, J., C. S. Brown, and D. Tilman. (2003). Community assembly and invasion: An experi¬ mental test of neutral versus niche processes. Proceedings of the National Academy of Sciences 100: 8916-8920. Fargione, J., and D. Tilman. (2006). Plant species traits and capacity for resource reduction predict yield and abundance under competition in nitrogen-limited grassland. Functional Ecology 20: 533-540. Foster, B. L., C. Murphy, K. Keller, T. Aschenbach, E. Questad, and K. Kindscher. (2007). Restora¬ tion of prairie community structure and ecosystem function in an abandoned hayfield: A sowing experiment. Restoration Ecology 15: 652-661. Foster, B. L. and D. Tilman. (2003). Seed limitation and the regulation of community structure in oak savanna grassland. Journal of Ecology 91: 999-1007. Gleason, H. A. (1910). The vegetation of the inland sand deposits of Illinois. Bulletin of the Illinois State Laboratory of Natural History 9: 23-174. Goldberg, D. E. and A. M. Barton. (1992). Patterns and consequences of interspecific competition in natural communities: A review of field experiments with plants. The American Naturalist 139: 771-801. Gough, L., J. B. Grace, and K. L. Taylor. (1994). The relationship between species richness and com¬ munity biomass: The importance of environmental variables. Oikos 70: 271-279. Grace, J. B. (1999). The factors controlling species density in herbaceous plant communities: An as¬ sessment. Perspectives in Plant Ecology, Evolution, and Systematics 2: 1-28. Greenlee, J. T. and R. M. Callaway. (1996). Abiotic stress and the relative importance of interference and facilitation in montane bunchgrass co mm unities in western Montana. The American Natural¬ ist 148: 386-396. Grime, J. P. (1974). Control of species density in herbaceous vegetation. Journal of Environmental Management 1: 151-167. Hauser, R. S. (1953). An ecological analysis of the isolated prairies of Newaygo County, Michigan. Ph.D. Dissertation. Michigan State College. Hector, A., B. Schmid, C. Beierkuhnlein, M. C. Caldeira, M. Diemer, and P. G. Dimitrakopoulos. (1999). Plant diversity and productivity experiments in European grasslands. Science 286: 1123-1127. Henderson, C. B., K. E. Petersen, and R. A. Redak. (1988). Spatial and temporal patterns in the seed bank and vegetation of a desert grassland community. Journal of Ecology 76: 717-728. Hermann, K. D., L. A. Masters, M. R. Penskar, A. A. Reznicek, G. S. Wilhelm, and W. W. Brodow- icz. (1996). Floristic quality assessment with wetland categories and computer application pro- 122 THE MICHIGAN BOTANIST Vol. 54 grams for the state of Michigan. Michigan Department of Natural Resources, Wildlife Division, Natural Heritage Program. Lansing. Holmgren, M., M. Scheffer, and M. A. Huston. (1997). The interplay of facilitation and competition in plant communities. Ecology 78: 1966-1975. Houseman, G. R., and R. C. Anderson. (2002). Effects of jack pine plantation management on bar¬ rens flora and potential Kirtland’s warbler nest habitat. Restoration Ecology 10: 27-36. Howe, H. F. (1994). Managing species diversity in tallgrass prairie: Assumptions and implications. Conservation Biology 8: 691-704. Howe, H. F. (1995). Success and fire season in experimental prairie plantings. Ecology 76: 1917-1925. Howe, H. F., J. S. Brown, and B. Zom-Arnold. (2002). A rodent plague on prairie diversity. Ecology Letters 5: 30-36. Ives, A. R., J. L. Klug, and K. Gross. (2000). Stability and species richness in complex communities. Ecology Letters 3: 399-411. Jordan, N. R., D. L. Larson, and S. C. Huerd. (2008). Soil modification by invasive plants: Effects on native and invasive species of mixed-grass prairies. Biological Invasions 10: 177-190. Kennedy, T. A., S. Naeem, K. M. Howe, J. M. Knops, D. Tilman, and P. Reich. (2002). Biodiversity as a barrier to ecological invasion. Nature 417: 636-638. Kost, M. A. (2004). Natural community abstract for dry sand prairie. Michigan Natural Features In¬ ventory, Lansing. Kost, M. A., D. A. Albert, J. G. Cohen, B. S. Slaughter, R. K. Schillo, C. R. Weber, and K. A. Chap¬ man. (2007). Natural co mm unities of Michigan: Classification and description. Michigan Natural Features Inventory, Report Number 2007-21, Lansing. Lockwood, J. L. (1997). An alternative to succession: Assembly rules offer guide to restoration ef¬ forts. Restoration and Management Notes 15: 45-50. Maestre, F. T., S. Bautista, and J. Cortina. (2001). Potential for using facilitation by grasses to estab¬ lish shrubs on a semiarid degraded steppe. Ecological Applications 11: 1641-1655. Martin, L. M., K. A. Moloney, and B. J. Wilsey. (2005). An assessment of grassland restoration suc¬ cess using species diversity components. Journal of Applied Ecology 42: 327-336. McCann, M. T. (1991). Land, climate, and vegetation of Michigan. Pp. 15-31 in The atlas of breed¬ ing birds of Michigan. R. Brewer, G. A. McPeek, and R. J. Adams Jr., editors. Michigan State Uni¬ versity Press, East Lansing. Middleton, E. L., J. D. Bever, and P. A. Schultz. (2009). The effect of restoration methods on the qual¬ ity of the restoration and resistance to invasion by exotics. Restoration Ecology 18: 181-187. Naeem, S. and J. P. Wright. (2003). Disentangling biodiversity effects on ecosystem functioning: De¬ riving solutions to a seemingly insurmountable problem. Ecology Letters 6: 567-579. Nemec, K. T., C. R. Allen, C. J. Helzer, and D. A. Wedin. (2013). Influence of richness and seeding density on invasion resistance in experimental tallgrass prairie restorations. Ecological Restoration 31: 168-185. Noss, R. F., E. T. LaRoe III, and J. M. Scott. (1995). Endangered ecosystems of the United States: A preliminary assessment of loss and degradation. Biological Report 28. United States Department of Interior, National Biological Service, Washington, D.C. Ojima, D. S., D. S. Chimel, W. J. Parton, and C. E. Owensby. (1994). Long- and short-term effects of fire on nitrogen cycling in tallgrass prairie. Biogeochemistry 24: 67-84. Peltzer A., and M. Kochy. (2001). Effects of grasses and woody plants in mixed-grass prairie. Ecol¬ ogy 89: 519-527. Plumb-Mentjes, M. L., and G. E. Center. (1990). The dynamics of a sand prairie plant community. Pp 123-127 in Proceedings of the Seventh North American Prairie Conference. C. L. Kucera, ed¬ itor. Southwest Missouri State University, Springfield. Pokorny, M. L., R. L. Sheley, C. A. Zabinsky, R. E. Engel, T. J. Svejcar, and J. J. Borkowski. (2005). Plant functional group diversity as a mechanism for invasion resistance. Restoration Ecology 13: 448^159. Polley, H. W., J. D. Derner, and B. J. Wilsey. 2005. Patterns of plant species diversity in remnant and restored tallgrass prairies. Restoration Ecology 13: 480^487. Raison, R. J. (1979). Modification of the soil environment by vegetation fire, with particular refer¬ ence to nitrogen transformations: A review. Plant and Soil 51: 73-108. 2015 THE MICHIGAN BOTANIST 123 Risser, P. G. (1988). Abiotic controls on primary productivity and nutrient cycles in North American grasslands. Ecological Studies: Analysis and Synthesis 67: 115-130. Roos, R. C. (2014). The initial effects of community variables on sand prairie restoration: species es¬ tablishment and community responses. Masters Thesis, Grand Valley State University, Allendale, Michigan. Available at http://scholarworks.gvsu.edu/cgi/viewcontent.cgi?article=1735&context= theses. Samson, F., and F. Knopf. (1994). Prairie conservation in North America. BioScience 4: 418N21. Schaetzl, R. J, and S. Anderson. (2005). Soils: Genesis and geomorphology. Cambridge University Press, New York, N. Y. Schramm, P. (1990). Prairie restoration: A twenty-five year perspective on establishment and man¬ agement. Pp. 169-178 in Proceedings of the Twelfth North American Prairie Conference. D. D. Smith and C. A. Jacobs, editors. University of Northern Iowa, Cedar Falls. Seabloom, E. W., W. S. Harpole, O. J. Reichman, and D. Tilman. (2003). Invasion, competitive dom¬ inance, and resource use by exotic and native California grassland species. Proceedings of the Na¬ tional Academy of Sciences 100:13384-13389. Seastedt, T. R. and A. K. Knapp. (1993). Consequences of nonequilibrium resource availability across multiple t im e scales: The transient maxima hypothesis. American Naturalist 141: 621-633. Shannon, C. E. (1948). A mathematical theory of communication. Bell System Technical Journal 27: 379—423. Sluis, W. J. (2002). Patterns of species richness and composition in re-created grassland. Restoration Ecology 10: 677-684. Smith, M. D., D. C. Hartnett, and G. W. T. Wilson. (1999). Interacting influence of mycorrhizal sym¬ biosis and competition on plant diversity in tallgrass prairie. Oecologia 121: 574-582. Smith, T. and M. Huston. (1990). A theory of spatial and temporal dynamics of plant co mm unities. Vegetatio 83: 49-69. SPSS Inc. 2011. PASW Statistics 18.0. SPSS Inc., Chicago, Illinois. Suding, K. N. 1999. Processes responsible for changes in plant species abundance following distur¬ bance. Ph.D. Dissertation, University of Michigan, Ann Arbor. Swink, F., and G. Wilhelm. (1994). Plants of the Chicago region. Indiana Academy of Science, Indi¬ anapolis. Tabachnick, G. G., and L. S. Fidell. (2007). Experimental designs using ANOVA. Duxbury Press, Belmont, C alifornia. The Taxonomic Name Resolution Service. (2014). The Taxonomic Name Resolution Service. iPlant Collaborative. Version 3.2. Available at http://tnrs.iplantcollaborative.org. (Accessed June 27, 2014). Tilman, D. (1988). Plant strategies and the dynamics and structure of plant communities. Princeton University Press, Princeton, New Jersey. Tilman, D., and J. A. Downing. (1994). Biodiversity and stability in grasslands. Nature 367: 363-365. Tilman, D., D. Wedin, and J. Knops. (1996). Productivity and sustainability influenced by biodiver¬ sity in grassland ecosystems. Nature 379: 718-720. Tilman, D. (1997). Community invasibility, recruitment limitation, and grassland biodiversity. Ecol¬ ogy 78: 81-92. Tilman, D. (1999). The ecological consequences of changes in biodiversity: A search for general principles. Ecology 80: 1455-1474. Transeau, E. N. (1935). The prairie peninsula. Ecology 16: 423^137. United States Department of Agriculture. (2011). Web Soil Survey, Natural Resources Conservation Survey. Available at http://www.websoilsurvey.nrcs.usda.gov. (Accessed March 11, 2011). Wedin, D. A., and D. Tilman. (1990). Species effects on nitrogen cycling: A test with perennial grasses. Oecologia 84: 433—441. Yachi, S., and M. Loreau. (1999). Biodiversity and ecosystem productivity in a fluctuating environ¬ ment: The insurance hypothesis. Proceedings of the National Academy of Sciences 96: 1463-1468. Young, T. P, J. M. Chase, and R. T. Huddleston. (2001). Community assembly and succession: Com¬ paring, contrasting, and combining paradigms in the context of ecological restoration. Ecological Restoration 19: 5-18. Zimmerman, U. D., and C. L. Kucera. (1977). Effects of composition changes on productivity and biomass relationships in tallgrass prairie. American Midland Naturalist 97: 465-469. 124 THE MICHIGAN BOTANIST Vol. 54 NOTEWORTHY COLLECTION NELUMBO LUTEA WILLD. (NELUMBONACEAE): OCCURRENCE OF A RARE PLANT IN WESTERN MICHIGAN CONFIRMED Previous Knowledge. Nelumbo lutea (American Lotus, Lotus-lily, Water Chinquapin) is an aquatic plant of quiet, shallow waters, widespread in the east¬ ern half of North America from Florida, Louisiana, and Texas northward to southern Maine and Massachusetts west to southern Ontario, southern Michigan, Wisconsin, southeastern Minnesota, Iowa, and eastern Nebraska (Crow & Hel- lquist 2000; USDA, NRCS 2014). Nelumbo lutea was documented on the distri¬ bution map in Voss and Reznicek (2012) only for Wayne and Monroe Counties, which are adjacent to the well-established and extensive populations in Ohio along the southern shore of Lake Erie from Erie, Sandusky, Ottawa, and Lucas Counties (personal observation), as well as a somewhat disjunct site in south¬ western Michigan in Kalamazoo County. According to information from the Michigan Natural Features Inventory (Slaughter, pers. comm.), numerous records from Monroe County are mostly from the Lake Erie marshes. Wayne County records document a population of N. lutea adjacent to Lake Erie, as well as from historic sites on the Huron and Rouge Rivers. The single historical report for Kalamazoo County of a population from Sun¬ set Lake in the town of Vicksburg (Hanes and Hanes 1947) has had its native sta¬ tus questioned (Voss 1985). It had reputedly occurred in the marsh at Sunset Lake in Vicksburg since at least the middle of the 19 th century, according to Voss (1985), based on Beal’s (1878), citation of a report from a Frank Tuthill of Kala¬ mazoo that the American Lotus must have been “introduced after the country was settled,” as the pond in which the plant was found was a mill pond. Beal also mentions a Mrs. Adams, who had lived nearby, who thought Nelumbo lutea had grown in the natural pond prior to the establishment of the mill in 1829 (Beal 1878). Beal later stated under his account of this species for all of Michigan, “Perhaps introduced by the Indians” (Beal & Wheeler 1892, Beal 1904). In his article updating the “Flora and Vegetation of Kalamazoo County, Michigan,” McKenna (2005, p. 226) stated: “To my knowledge, [the Vicksburg population was] last collected by the Haneses on 11 July 1934.” He also reported that a “short, typed report on the flora of the county written by C. R. Hanes (date of preparation unknown, but certainly post 1934) reads ‘The Lotus, which formerly grew at Vicksburg ....’” McKenna (2005) regarded it as extirpated and proba¬ bly not native. Voss (1985) indicated that various attempts had been made in the 1800s to in¬ troduce new populations in the Detroit River, with none persisting. I once tried to establish a population of Nelumbo lutea in the old reservoir on the campus of the University of New Hampshire, Durham, New Hampshire, with plants taken from a lake in Wisconsin. After I transplanted three tubers, the plants flowered 2015 THE MICHIGAN BOTANIST 125 nicely the following summer, thereafter producing only a few floating leaves for a few years, and then disappearing from the transplanted site. So, if the Vicks¬ burg population was sustained for nearly a century, does it not stand to reason that the occurrence might have been natural? Furthermore, Michigan Flora On¬ line (2011) notes that while Nelumbo lutea may have been regarded as perhaps introduced . . by Native Americans as a food-plant, . . . there is no documen¬ tation one way or the other. Some of our colonies, at least, are assumed to be nat¬ ural, at the northern edge of the range for the species.” Discussion. The Michigan Natural Features Inventory (2007) lists Nelumbo lutea as a rare plant with Threatened state status and S2 (Imperiled) state rank, although it is considered globally Apparently Secure (G4). This documentation of two populations in Ottawa County by herbarium specimens reconfirms the presence of this rare plant in western Michigan, which was first collected by K. Karsten on August 25, 1942; the occurrence was unknown to Voss (1985) in his extensive work on his Michigan Flora and was still unknown to Voss and Reznicek (2012) at the time of publication of Field Manual of Michigan Flora. The western Michigan population came to my attention as I was working in the Calvin College Herbarium (CALVIN) in January 2014 on a project to image all of the specimens in the collection. I ran across two sheets of a collection that was made from a population with the label data: “Grand River at Stearns Bayou, Ot¬ tawa Co., Coll. M. Karsten la & 7B , 7-25- ’42” (Figure 1). Since there was no dot recorded for Ottawa County on the distribution map for this species in Voss and Reznicek (2012), I sent images of Professor Karsten’s collection to A. A. Reznicek so that the Michigan Flora Online (2011) website could be updated; Ottawa County has now been added to the list of counties and a dot added to the map. In August 2014 I set out to relocate the 1942 Ottawa County population doc¬ umented by Professor Karsten. Using Google Earth I was able to locate a bridge crossing Stearns Bayou on Green Street, Robinson Township. A large population was hiding in plain sight adjacent to an old store/boat livery and launching site called Felix’s Landing, with a small sign, “Stearns Bayou,” on the side of the building. This large population, in full bloom (Figure 2), was adjacent to a large, dense cattail marsh; the opposite side of the waterway was dominated by Nymphaea odorata, but no Nelumbo lutea was observed on that side near the bridge. A second visit to the area in August 2015, by canoe, revealed that this population was extremely large, extending the whole length of the channel (both sides) connecting Stearns Bayou to the main flow of the Grand River. Specimens collected now provide present-day documentation of that population, which had been visited by Professor Karsten 72 years earlier, at Stearns Bayou. Approximately one mile westward, I found a second population along the west shore of Millhouse Bayou, in Grand Haven Township. Here the population was small, consisting mostly of scattered plants at the Township picnic area wa¬ terfront and at an adjacent neighbor’s dock, growing in water ca. 1 m deep. Spec¬ imens were collected to document this additional locality. In searching for additional information on Nelumbo lutea in Michigan, I was surprised to learn that the Michigan Natural Features Inventory (2007) (MNFI) also had internal reports on two other populations in the Grand River system in Ottawa County (Slaughter, pers. comm.): (1) a large population in Grand Haven 126 THE MICHIGAN BOTANIST Vol. 54 C°h»ir Colton HartuMum iiiiiiiiiSiiii TTHchigan Ho, t z l //elMtnbo lutes l (..VUt&tj £f-r&ru( Rive*- is &«.+*>* &rtjL*A.C 4 , FIGURE 1. Herbarium specimen of Nelumbo lutea collected by K. Karsten in 1942 from Stearns Bayou, Ottawa County, Michi¬ gan (CALVIN). Township at the east end of Martinique Island, approximately 3 miles down¬ stream from my Grand Haven Township site; and (2) a large population at the mouth of the Stearns Bayou, about 0.5 mile from my Green Street Bridge site. Furthermore, the MNFI website has posted a single site in each of Barry, Berrien, and Genesee Counties—all unmapped on Michigan Flora Online (2011). There is clearly a need for documentation of populations of rare plants by depositing voucher specimens in our major state herbaria, thereby assuring that there will be a permanent record. Information from internal records at MNFI (Slaughter, pers. comm.) for Genesee County indicates a healthy population on Fenton Lake in 2005; in southwest Michigan, the Barry County population at Al¬ gonquin Lake was surveyed in 2004; and a report for Boyle Lake, Berrien County, first reported in 1934, indicates that N. lutea was “Apparently observed in 1981 . . . with no other data provided.” Diagnostic Characters: Nelumbo lutea is readily recognized, being a large aquatic plant producing large circular leaves with the petiole attached in the cen¬ ter of the blade (peltate), the leaves standing erect and emergent, and as floating leaves (especially in deeper water or earlier in the growing season). Populations are conspicuous in late summer (Figures 2 and 3), as the flowers are pale-yellow 2015 THE MICHIGAN BOTANIST 127 FIGURE 2. View of present-day population of Nelumbo lutea at Green Street Bridge, Steams Bayou, Robinson Township, Ottawa Co., Michigan. Photo by Garrett Crow. FIGURE 3. Nelumbo lutea population at mouth of Steams Bayou. Photo by Garrett Crow. to sulfur yellow, with flowering often peaking in mid-August; the petals and sta¬ mens are numerous, and the numerous carpels are separate and imbedded in a broad, flat-toped receptacle (Figure 4). Large populations can readily be spotted from a distance. The plants readily spread by rhizomes, often forming large pop¬ ulations that are near monocultures. Large tubers are produced in late summer, and were eaten by Native Americans. The fruits are round nutlets with a hard 128 THE MICHIGAN BOTANIST Vol. 54 FIGURE 4. Flower of Nelumbo lutea. Photo by Garrett Crow. FIGURE 5. Maturing fruiting receptacle (torus) of Nelumbo lutea with developing nutlets in cavities. Photo by Garrett Crow. 2015 THE MICHIGAN BOTANIST 129 outer wall, remaining in or shed from cavities in the mature woody receptacle (Figure 5). Nelumbo lutea is the only species of the Nelumbonaceae native to North America and is closely related to the well known pink- or white-flowered N. nu- cifera (Sacred Lotus or Oriental Lotus) native to eastern Asia and northern Aus¬ tralia and commonly cultivated in water gardens. It occurs very sporadically as naturalized populations in the southeastern United States (Wiersema 1997). Nelumbo lutea and N. nucifera, the only two taxa in the Nelumbonaceae, are very closely related, and it has even been proposed that the two be regarded as a sin¬ gle species with two widely disjunct subspecies; the name for our taxon would then be N. nucifera subsp. lutea (Willd.) Borsch & Barthlott. But I am retaining the two species viewpoint, following the nomenclature of the taxonomic treat¬ ment of Flora of North America North of Mexico (Wiersema 1997), Field Man¬ ual of Michigan Flora (Voss and Reznicek 2012), and Aquatic and Wetland Plants of Northeastern North America (Crow and Hellquist 2002). Specimen Citations. Michigan, Ottawa Co., Grand River at Stearns Bayou. July 25, 1942. M. Karsten 7a & 7B [CALVIN002319 & CALVIN002321]; Robinson Township, Stearns Bayou at Green Street Bridge, Felix’s Landing; backwater off Grand River; large population adjacent to boat launch on northeast side of bayou. August 6, 2014, Garrett E. Crow 10831, and 17 August 2015, Garrett E. Crow 10978 (CALVIN, MICH, MSC); Grand Haven Township, Mill- house Bayou at township picnic area on Bignell Road; backwater off of Grand River; small population with scattered plants in shallow areas near shore. August 6, 2014. Garrett E. Crow 10836 (CALVIN, MICH, MSC). ACKNOWLEDGMENTS Special thanks are extended to Brad Slaughter for providing information from the Michigan Nat¬ ural Features Inventory on Nelumbo lutea, as well as for providing suggestions on improving the manuscript. Anonymous reviewers also provided constructive comments. Th a nk s are extended to David and Charylene Powers for providing the opportunity to examine the population by canoe, as well as to my field assistant, Charlyn Crow. LITERATURE CITED Beal, W. J. (1878). Nelumbium luteum in Michigan. Botanical Gazette 3: 13. Beal, W. J. (1904). Michigan flora, fern and seed plants growing without cultivation. Annual Report of the Michigan Academy of Science. 5: 1-147. [reprinted in 1905 as Michigan flora: a list of the fern and seed plants growing without cultivation. The State Board of Agriculture, Agricultural College, Mich., Lansing, Michigan.] Beal, W. J. and C. F. Wheeler. (1892). Michigan Flora. Robert Smith & Co., Lansing, Michigan. Crow, G. E., and C. B. Hellquist. (2000). Aquatic and wetland plants of northeastern North Am erica: Vol. 1. Pteridophytes, gymnosperms and angiosperms: Dicotyledons. (Corrected paperback edi¬ tion, 2006). University of Wisconsin Press, Madison. Hanes, C. R., and F. N. Hanes. (1947). Flora of Kalamazoo County, Michigan: Vascular plants. Pri¬ vately published, Schoolcraft, Michigan. McKenna, D. D. (2005). Flora and vegetation of Kalamazoo County, Michigan. The Michigan Botanist 43: 137-359. MICHIGAN FLORA ONLINE. A. A. Reznicek, E. G. Voss, & B. S. Walters. (2011). University of Michigan. Available at http://michiganflora.net/home.aspx (Accessed Oct 5, 2015). 130 THE MICHIGAN BOTANIST Vol. 54 Michigan Natural Features Inventory. (2007). Rare species explorer. Available at http://mnfi.anr.msu.edu/explorer (Accessed Dec 18, 2014) USDA, NRCS. (2014). Nelumbo lutea. The PLANTS Database. National Plant Data Team, Greens¬ boro, North Carolina. Available at http://plants.usda.gov/core/profile7symboHNELU (Accessed December 18, 2014). Voss, E. G. (1985). Michigan Flora. Part II. Dicots (Saururaceae-Cornaceae). Cranbrook Institute of Science Bulletin 59 and University of Michigan Herbarium, Ann Arbor. Voss, E. G., and A. A. Reznicek. (2012). Field manual of Michigan flora. University of Michigan Press, An n Arbor. Wiersema, J. H. (1997). Nelumbonaceae. Pp. 64-65 in Flora of North America, Volume 3: Magno- liophyta: Magnoliidae and Hamamelidae. Flora of North America Editorial Committee, editors. Oxford University Press, New York, N.Y. -Garrett E. Crow Professor Emeritus, University of New Hampshire Visiting Scholar, Calvin College Visiting Research Botanist, Michigan State University Herbarium garrett. crow@unh.edu 2015 THE MICHIGAN BOTANIST 131 NOTEWORTHY COLLECTION THE FIRST RECORD FOR MICHIGAN OF GALIUM PEDEMONTANUM (RUBIACEAE) Significance of the Report. The first recorded occurrence in Michigan of this widespread European annual. Previous Knowledge. The earliest reported collection of Galium pedemon- tanum (Bellardi) All. (Foothills bedstraw) in the United States is from Kentucky in 1933 (Sanders 1976). By 1976, it had already spread to Arkansas, Illinois, Missouri, Ohio, Oklahoma, Tennessee, Virginia, and West Virginia (Sanders 1976). Currently, its distribution includes the additional states of New York, Pennsylvania, North Carolina, South Carolina. Georgia, Mississippi, Louisiana, Texas, Indiana, Nebraska, and the Pacific Northwest states of Washington, Ore¬ gon, Idaho, and Montana (USDA, NRCS 2015). The nearest location to the one reported here for Michigan is about 23 km south-southeast in St. Joseph County, Indiana (Swink and Wilhelm 1994). Discussion. Galium pedemontanum, a slender, sprawling, annual plant with retrorsely-barbed stems, was probably an accidental introduction into the flora of the United States. Using collection data from herbarium specimens, Sanders (1976) traced the spread of the species from its first reported locality, showing that it had been transported in all directions from Lexington, Kentucky. Going a step further, Sanders also predicted the future distribution of this species by iden¬ tifying and using the soil and climate characteristics of its European habitats to delineate similar sites in the USA. In Europe, G. pedemontanum is found where the mean annual temperature is 8-18°C, the mean annual rainfall is 40-150 cm., and the soils, which have developed under forests, are acidic (Sanders 1976). With a mean annual temperature of 7-9°C, a mean annual precipitation of 76-81 cm, and soils mapped as fine sands which are acidic to neutral (USDA, NRCS 2013), the climate and soils of southwestern Michigan are similar to Sanders’ European climate and soils. Interestingly, his map shows southwestern Michigan at one of the northern edges of the potential USA distribution of G. pedemon¬ tanum. I spotted the population of Galium pedemontanum, the dozen plants of which were scattered over an area of approximately 2 m 2 , because the elongated, lax, pale stems stood out against the dark green of the lawn grasses in the middle of summer 2015. In early May 2016, numerous plants, now verdant and semi-erect, were growing in a 25 m 2 area at the same location, but now associated with other spring-flowering introduced species (e. g., Scleranthus annuus L., Veronica ar- vensis L., Arenaria serpyllifolia L., Cardamine hirsuta L., Lamium purpureum L., Oxalis stricta Jacq., and Myosotis stricta Roem. & Schult.). Because this lo¬ cation is close to a road, I suspect that the seeds of G. pedemontanum may have been carried in by vehicle traffic. Recent studies (e.g., Ehrendorfer and Barfuss 2014) suggest that Galium 132 THE MICHIGAN BOTANIST Vol. 54 FIGURE 1. Galium pedemontanum at the collection site. pedemontanum, along with several other Eurasian species, should be segregated as the genus Cruciata. In that case, our species would be known as Cruciata pedemontana (Bellardi) Ehrend. Diagnostic Characters. Galium pedemontanum is one of only two annual species among the 21 species of Galium now known from Michigan (Voss and 2015 THE MICHIGAN BOTANIST 133 Reznicek 2012). It is a weak-stemmed, scrambling plant with noticeably scabrous stems. It may be distinguished from the other annual species in Michi¬ gan, Galium aparine L., by its greenish-yellow flowers with glabrous ovaries and its four small ovate leaves per node, and from Galium verum L., the other yel¬ low-flowered Michigan bedstraw, by its shorter leaves. Below are the key char¬ acters that can be used to separate Galium pedemontanum from Galium aparine and from Galium verum. Galium aparine —leaves 6-8 at each node, 1-8 cm long, flowers greenish white, fruits hispid. Galium pedemontanum —leaves 4 at each node, 3-11 mm long, flowers dull yel¬ low, fruits glabrous Galium verum —perennial, leaves 4 at each node, 1-5.4 cm long, flowers yellow, fruits glabrous. Specimen Citation. Michigan, Berrien County, Sawyer (N41° 53.681'; W86° 36.585'). In sandy soil of open lawn with Poa pratensis, Medicago lupulina and Oxalis stricta. July 27, 2015, Tatina s.n. (MICH). LITERATURE CITED Ehrendorfer, F., and M. H. J. Barfuss. (2014). Paraphyly and polyphyly in the worldwide tribe Ru- bieae (Rubiaceae): Challenges for generic delimitation. Annals of the Missouri Botanical Gar¬ den 100: 79-88. Sanders, R. W. (1976). Distributional history and probable ultimate range of Galium pedemontanum (Rubiaceae) in North America. Castanea41: 73-80. Swink, F., and G. Wilhelm. (1994). Plants of the Chicago region, 4 th edition. Indiana Academy of Science. Indianapolis. USDA, NRCS. (2103). Web Soil Survey. Available at www.websoilsurvey.nrcs.usda.gov/. (Accessed May 12, 2016). USDA, NRCS. (2015). The PLANTS database. National Plant Data Team, Greensboro, North Car¬ olina. Available at plants.usda.gov/core/profile?symbol=CRPE10 (Accessed July 28, 2015). Voss, E. G., and A. A. Reznicek. (2012). Field manual of Michigan flora. The University of Michi¬ gan Press. Ann Arbor. -Robert Tatina Dakota Wesleyan University Mitchell, South Dakota 57301 rotatina@dwu. edu 134 THE MICHIGAN BOTANIST Vol. 54 BOOK REVIEW Emmet J. Judziewicz, Robert W. Freckman, Lynn G. Clark, and Merel R. Black. 2014. Field Guide to Wisconsin Grasses. The University of Wisconsin Press, Madison, ix + 396 pp. ISBN 978-0-299-30134-7. Paperback. $29.95 (also available as an e-book for $24.95). Andrew L. Hipp. 2008. Field Guide to Wisconsin Sedges: An Introduction to the Genus Carex (Cyperaceae) . The University of Wisconsin press, Madison, xi + 265 pp. ISBN 978-0-299-22594-1. Paperback. $29.95 (also available as an e-book for $19.95) (out of print in hardcover). A significant component of nearly all ecosystems consists of the graminoids—grasses (family Poaceae), sedges (family Cyperaceae), and rushes (family Juncaceae)—and in ecosystems such as prairies and wetlands, where they are particularly numerous in species, they also form the bulk of the biomass. Accurate identification of graminoids is therefore essential for any floristic study, ecological survey, restoration project, or similar endeavor. Unfortunately, these ecologically and floristically important plant groups are rarely included at all in wildflower guides and other popular references. In many places, one must depend on standard floras that, although including graminoids, often do little to make identification easy, relying primarily on keys, only sometimes containing descriptions (often short), and often lacking any illustrative materials. Such flo¬ ras generally give short shrift to explanations of the unique structural features of these plants and the specialized terminology necessary for understanding and identifying them. For an expert, this may be no great barrier, but for the novice, it creates a steep learning curve indeed. Wisconsin is fortunate in having two up-to-date guides that cover all species of two major groups of graminoids. The first book under review covers all of the grass species in Wisconsin, and its companion covers all of the species of Carex, which, with 157 species, is the largest genus of sedges in Wisconsin. Both of these volumes go to great lengths to make life easy for the novice and expert alike, while still being complete and technically accurate. Let’s look at the Field Guide to Wisconsin Grasses first. The book starts with a detailed introduction to grass structure, replete with photographs of such things as culms (stems), leaves (including sheaths, auricles, and ligules), all the details and variations of inflorescences—spikelets and their various arrangements, glumes, and lemmas—and finally the tiny flowers. There is a key to genera, ac¬ companied by detailed line drawing of the relevant characters in the margin. The key is preceded by several very useful features: (i) a discussion of what to look for in identifying an unknown grass, with reference to several critical characters, (ii) lists of grasses with unusual characters, and (iii) a list and short descriptions of the ten commonest grass species in the state. The main part of the book is devoted to a species-by-species account, 2015 THE MICHIGAN BOTANIST 135 arranged alphabetically by genus. Each genus has a short description and, if there are more than one species in Wisconsin, a key to species. The individual species treatments vary somewhat, depending on what is felt to be needed in each case. There may be a short description; more commonly, however, a brief statement of characters contrasting the species with related, or similar, species, is given. Habi¬ tat and distribution data is generally included, as well as a county distribution map. Most usefully, most accounts contain both photographs and line drawing of important characteristics. Field Guide to Wisconsin Sedges takes a somewhat different approach. Al¬ though there is a brief discussion of structure and terminology, the principal early portion of the book is a breakdown of the genus Carex into two subgenera and then into artificial groupings within each subgenus. Several keys are pro¬ vided, ultimately reaching all species of Carex in Wisconsin. These are followed by a lengthy treatment of the several sections of the genus, each with a descrip¬ tion, and a discussion of the species within the section. All of this preliminary material is based on the sound idea that learning sedges is made much easier by grasping the subgenera and sections before trying to tackle identification of in¬ dividual species. The individual species treatments, which constitute about half of the book, treat 55 common species in detail. Each treatment contains a brief diagnostic statement, a detailed description of the plant, a description of the habitat and range within the state, and, significantly, a detailed discussion of sim¬ ilar species and discussions between them. Each treatment is also accompanied by an illustration in ink and watercolor by the artist Rachel D. Davis containing both a view of the habit of the plant and a detailed view of the inflorescence or individual perigynia (the closed sac-like structures that enclose the female flow¬ ers). Distribution maps of all species of Carex in the state, prepared by Merel R. Black and Theodore S. Cochrane, are collected in an appendix. A particularly valuable feature of both volumes under review is a discussion of the plant communities in which species of grasses or sedges occur, along with lists of species for each habitat. For the novice in particular, this is a very useful aid in narrowing down the process of identifying an unknown. This portion of the sedge guide was prepared by Theodore S. Cochrane. Because a large percentage of the species treated in both of the field guides tend to occur in suitable habitats throughout the western Great Lakes region, these volumes will be valuable resources for students and investigators alike throughout this region. Michael Huft 136 THE MICHIGAN BOTANIST Vol. 54 ANNOUNCEMENTS ISOBEL DICKINSON MEMORIAL AWARD RECIPIENTS Congratulations to Emily Huizenga and Christopher Bouma, who were the re¬ cipients of the Isobel Dickinson Memorial Award for best student-authored paper published in The Michigan Botanist, Volume 52. The selected paper was entitled “Assessing a Reconciliation Ecology Approach to Suburban Landscap¬ ing: Biodiversity on a College Campus” by Christopher Bouma, Emily Huizenga, and David Warners, The Michigan Botanist 52: 93-104. We acknowl¬ edge the Michigan Botanical Club-Dickinson Award Committee for evaluation of the student papers and the Michigan Botanical Foundation for funding this award. REVIEWERS FOR VOLUMES 53 AND 54 I wish to thank the following people who reviewed manuscripts for Volumes 53 and 54 of The Michigan Botanist. Their comments were important, both to the authors and to the editor, and their efforts, which are essential to maintaining the high quality of the journal, are greatly appreciated. Dennis Albert Garrett Crow Bob Grese Virginia Hayes Emmet Judziewicz Sheila Lyons-Sobaski Scott Namestnik Michael Penskar Jeff Plakke Anton A. Reznicek Dennis Riege Paul Rothrock Michael Rotter Lisa Schulte Gordon Tucker Dennis Woodland John Wiersema INSTRUCTIONS TO AUTHORS Refer to http://quod.lib.umich.edU/m/mbot/submit for more detailed instructions, especially for formatting, style conventions, literature cited, and voucher specimen requirements. Please contact the editor with any questions. 1. Create text in 12-point Times New Roman font and double space paragraphs throughout. Research articles should be organized as follows: Title, Author(s) and address(es), Abstract with up to 5 keywords, Introduction, Materials and Methods, Results, Discussion, Acknowledgements, Literature Cited, Tables, Figure Legends, and Figures. Sections may be omitted if not relevant. All pages should be numbered. 2. For noteworthy collections, manuscripts should be formatted as follows. The title, “Noteworthy Collections,” should begin each submitted manuscript, followed on the next line by the State or Province for the species reported. The next line should list the taxon of interest using the fol¬ lowing format: Species Author(s) (Family). Common name. The rest of the manuscript should include the following named sections: (i) Significance of the Report, (ii) Previous Knowledge, (iii) Discussion, (iv) Diagnostic Characters (if desired), (v) Specimen Citations, (vi) Acknowledgements (if desired), and (vii) Literature Cited. Each of these sections is largely self- explanatory; however, the “Significance of the Report” section should be limited to a brief sen¬ tence or phrase indicating the significance of the collection(s), and this may be expanded upon in the “Discussion” section; the “Specimen Citations” section should include the relevant label data from the voucher specimen(s) including location data, collector(s), collection number, etc., as well as the Index Herbariorum acronym(s) of the herbarium or herbaria where the specimen(s) are deposited. The manuscript should end with the name and address of the author(s). 3. Non-research articles, such as book reviews, letters to the editor, notices, biographies and other general interest articles can be formatted as general text without the specific sections listed above. However, literature cited and any tables or figures should be formatted as described below. 4. Create tables either as an MS Word table or using a tab-delimited format. Each table is to be sub¬ mitted as a separate file. Table captions should be placed at the top of the table. Any footnotes should appear at the bottom of the table. Please do not insert tables within the body of the text. 5. Send each figure as a separate file in a high-resolution format—eps, jpg, or tif. Figures like bar graphs that gain their meaning with color won’t work—use coarse-grained cross-hatching, etc. Create figure legends as a separate text file, and the typesetter will insert them as appropriate. Please do not insert the figure in the body of the text file. 6. Citations: Please verify that all references cited in the text are present in the literature cited sec¬ tion and vice versa. Citations within the text should list the author’s last name and publication year (e. g. Smith 1990). For works with more than 2 authors, use “et al.”, and separate multiple citations with a semicolon. 7. Literature Cited: List citations alphabetically by author’s last name. The first author’s name is to be listed with surname first, followed by initials (e.g. Smith, E. B.), and subsequent authors are to be listed with initials first. Separate author’s initials with a single space. The year of publica¬ tion should appear in parentheses immediately before the title of the citation. The entire journal name or book title should be spelled out. Please put a space after the colon when citing volume number and page numbers. 8. Italicize all scientific names. Voucher specimens must be cited in floristic works and in any other study whose results depend on the identity of the plant(s). Papers citing plant records without documenting vouchers are generally not acceptable. 9. Manuscripts must be submitted electronically to the email address of the editor. All manuscripts will be reviewed by at least two referees. CONTENTS In This Issue Michael Huft 81 Articles A Survey of Native and Invasive Genotypes of Phragmites australis along Michigan’s West Coast Grady H. Zuiderveen, Timothy M. Evans, Thomas J. Schmidt, and Mark R. Luttenton 82 The Initial Effects of Community Variables on Sand Prairie Restorations: Species Establishment and Community Responses Robert C. Roos and Todd A. Aschenbach 92 Noteworthy Collections Nelumbo lutea Willd. (Nelumbonaceae): Occurrence of a Rare Plant in Western Michigan Confirmed Garrett E. Crow 124 The First Record for Michigan of Galium pedemontanum (Rubiaceae) Robert Tatina 131 Book Reviews Field Guide to Wisconsin Grasses by Emmet J. Judziewicz, Robert W. Freckman, Lynn G. Clark, and Merel R. Black; and Field Guide to Wisconsin Sedges by Andrew L. Hipp Michael Huft 134 Announcements 136 On the cover: Nelumbo lutea, Beulah Beach, Ohio. Shallow pond behind the beach at mouth of creek entering Lake Erie. Crow 10830 (CALVIN, MICH, MSC). Photo by Garrett E. Crow.