Algal blooms often release toxins into the water column which may cause irritation of the skin and eyes, respiratory problems and even paralyze or kill fish and deplete oxygen levels. In the current study we created small patches of partially restored native plant communities along several small tributaries in the lower St. Johns River basin to investigate their effects on local water quality and biodiversity. Biodiversity was assessed across multiple taxa including algae, plants, as well as both aquatic and terrestrial invertebrates. Lastly, the ability of these semi-restored patches of native plant communities to ameliorate nutrient concentrations in both the soil and water was compared to non-manipulated urban tributaries. We found that even modestly restored riparian plant communities resulted in substantial increases in both terrestrial and aquatic biodiversity while decreasing nutrient loading for small tributaries of the lower St. Johns River. Increased plant biodiversity and the presence of rare species may have resulted from facilitation; this effect may have been enhanced because we purposely utilized plants that varied greatly in size, structure and habitat requirements. For instance, pickerelweed, which is an aquatic emergent species, likely aids in stabilization of the edges of the streambed and increases substrate accumulation around its stems while providing refugia for aquatic invertebrates and juvenile fish. Complementarily, shrub and tree species such as Virginia willow and river birch respectively, which are structurally much more complex and located at higher elevation along the stream banks, likely create microhabitats that have higher degrees of shading, lower temperatures and higher soil moisture.
The increase in structural complexity of partially restored riparian communities, results in increased spatial heterogeneity and creates more niches within the overall landscape. Although within-trophic level community composition is a composite of positive and negative interactions; facilitation has been hypothesized to be more important in stressed or highly disturbed conditions. Indeed, studies of plant communities have found that colonization of new sites, especially in stress habitats with high levels of light and elevated temperatures, ebb flow tray is restricted to shady microhabitats created by the canopy shade of “nurse plants” or, in the case of stream habitat, emergent species provide stabilization of lotic substrate which reduces mortality for other plants. Our study supports the general contention by Bertness & Callaway that plant communities in open, sunny, highly disturbed sites are more likely to be influenced by facilitation rather than shaded later successional stages. However, aquatic emergents and shoreline specialists such as pickerelweed and golden Canna respectively are also unlikely to compete strongly with plants at higher elevations along the stream banks .Because of the importance of the riparian zone to the stream and entire watershed, assessments of anthropogenic effects on lotic freshwater systems should include the banks as well. This is especially true if restoration of native vegetation is to be used to ameliorate damaged or disturbed stream ecosystems because the success of this technique depends upon long-term sustainable uptake of nutrients or pollutants while maintaining terrestrial biodiversity and ecosystem function. It has been shown that soil nutrient levels can have strong effects on plant and arthropod community composition, but these effects can be positive or negative. Therefore, terrestrial plant and arthropod community assessments should be conducted in order to determine whether or not native riparian buffer zones would be a viable method for reducing nutrient loading in a stream.
In the present study, experimental gardens absorbed more nutrients than did non-manipulated riparian zones, resulting in an increase in plant and arthropod diversity . The increase in arthropod richness resulted primarily from an increase in phytophagous Coleoptera and Lepidoptera . This was likely attributable to the concomitant increase in plant species diversity which has been shown in previous studies to affect the diversity of phytophagous insects and is supported by the strong positive correlation between plant and arthropod diversity measures reported in this study. Interestingly, however, there were also more species of insects with aquatic larval stages, primarily Odonata and Ephemeroptera, at sites with experimental gardens, suggesting increased recruitment of adults for mating and oviposition. It has been suggested that terrestrial arthropods are good indicators of a river system’s health. Therefore, the increase in terrestrial arthropod diversity observed in this study indicates that restoration of native riparian zones not only improves stream ecosystem health, but also improves riparian ecosystem health and may be a viable and sustainable restoration option. Algae are considered excellent indicators of the health of aquatic ecosystems and algal biodiversity, especially that of diatoms, has been correlated with nutrient levels, the presence of contaminants and temporal variability. In our study, Shannon algal biodiversity within the garden sites was significantly greater at the start and end of the monitoring period, but not significantly different during the most active growing season . While some researchers have also noted that nutrient additions do not necessarily elicit a significant change in algal biodiversity, others have reported that confounding factors may mask significant effects of nutrient additions. In the St. Johns River, changes in lotic algal diversity have been correlated with varying nutrient levels and seasonality over a two-year period. Our study, which examined similar factors in smaller tributary streams showed similar trends, but utilized much shallower tributaries and, therefore, was subject to different ecological parameters than previous studies.
The percent coverage of macroalgae was also more variable both spatially and temporally in garden sites, while non-garden sites exhibited significantly less visible cover and variability. The majority of the taxa composing macroalgal mats were filamentous greens , which is common in low-order, oligo-mesotrophic streams. Macroalgal mats support a diverse epiphytic algal and invertebrate community, which may account for the aquatic insect community we observed. An introduced species may behave invasively in the new range because it already possesses traits that confer invasiveness , or alternatively, it may evolve invasiveness in situ in the new range. Moreover, the introduction history of a species can affect the likelihood that it will become invasive. While many plant species suffer a genetic bottleneck when introduced in the new range, multiple introductions of a plant species may make it more likely that a species becomes invasive in part because multiple introductions can inflate genetic diversity. In addition, multiple introductions may result in the admixture of genomes that have never come into contact with each other creating novel, invasive genotypes that may express different traits and enhanced fitness. A successful introduction can also depend upon the relationship between the introduced individuals and the new environment. Introduced individuals may have different climatic tolerances than their native counterparts and/or wider climatic tolerances and this may contribute to their success. Testing whether invasive genotypes have a wider climatic tolerance than native genotypes requires planting clones of both known native and invasive genotypes in their home climate and in a different climate to test whether invasive genotypes have greater ability to survive and grow under a new climatic condition. The increased growth of invasive genotypes found under controlled conditions may not be observed under field conditions. Disentangling whether the reason is due to the traits of the introduced individuals or environmental factors in the field is difficult because traits favored in the new range may incur a cost in the native range. For example, reduced herbivore loads in the new range may have selected for genotypes that have reallocated resources from defense to growth resulting in invasive genotypes having faster growth than native ones. But, if these genotypes are transported back to the native region in which herbivores are more abundant, flood and drain tray they may experience higher levels of herbivore damage than their native counterparts and this can negate any increases in growth.
Yet, if invasive genotypes are superior, then invasive genotypes may still experience greater growth even if they are preferentially preyed upon. Documenting higher growth following herbivore damage is difficult in a common garden because of variability in herbivore prevalence and local conditions. To document such a pattern requires simulating herbivore damage experimentally. In this paper, we examine the performance of invasive genotypes when compared to that of native genotypes in the native’s own range in the invasive wetland plant, Phalaris arundinacea. The invasive grass Phalaris arundinacea is a good model system to address the issues of the emergence of novel and superior invasive genotypes. Invasive genotypes of Phalaris arundinacea have been shown to be the product of multiple introductions and subsequent admixture and in common greenhouse conditions, invasive genotypes were shown to have faster growth rate, to be taller, have more tillers and greater final biomass. Additionally, invasive genotypes were also shown to have a smaller genome size and had no consistent differences in genetic architecture. However, in order to test whether these invasive genotypes are superior requires that native and invasive genotypes be compared in the native range to determine if the invasive genotypes created in the new range are superior to native genotypes even under conditions in which native genotypes have evolved. In this way, we can determine if the superior performance of invasive genotypes results in overall greater performance under conditions in which the native genotypes should be favored. In addition, by planting genotypes in a field common garden, we can assess in situ herbivore damage to determine if invasive genotypes suffer more herbivore damage than native ones. However, greater herbivore damage may not necessarily result in reduced growth and/or biomass as greater herbivore damage may be compensated for by a faster growth rate. We test this idea in an experimental common garden in which the same genotypes are subject to biweekly biomass removal to simulate grazing by large herbivores. Finally, by planting genotypes collected from northern and southern populations into both northern and southern gardens, we can determine if differences in performance are due to genotypes having adapted to a similar climate rather than native and invasive differences.
e selected 18 invasive genotypes and 18 native genotypes . Fewer French genotypes were chosen because one French population contained only hexaploids, while all other populations contained tetraploids; we thus, eliminated hexaploid individuals from our study. Chosen genotypes were transplanted into pots in the University of Vermont greenhouse, where they were maintained, and then sequentially propagated prior to experimentation to remove any maternal environmental effects. Selected tillers of the chosen genotypes were placed into greenhouse flats, placed on their side and allowed to produce replicate tillers. In this way, we created identical copies of all genotypes used in the experiments. We chose garden sites in the Czech Republic and France . Because we were interested in whether introduced genotypes could outperform native genotypes under conditions where natives should be favored, we planted the native and invasive genotypes back into gardens near their original sampling location ensuring that climatic conditions in the garden were most similar and therefore favorable for the native genotypes. Our garden sites were within 50 km of the original collection of genotypes, ensuring that climatic conditions would be similar to the genotypes collection location. Each garden site was a wet meadow and characterized by a mix of herbaceaous perennial plants and also had large stands of Phalaris arundinacea. Specific site measurements were not collected at either site. However, at each site we planted nine replicate blocks designed to span the range of conditions found in the wet meadow. Our only criteria for eliminating a potential location for a plot was that it could not contain native Phalaris sincethe presence of established Phalaris might have made it difficult to locate our transplants. In our statistical models, we treated the nine blocks as a random factor to incorporate local unexplained environmental variation. After planting, the plots were left to regrow and we did not remove any of the native vegetation. Plant quarantine rules prevented us from bringing tillers into Europe. Therefore, all plants were transported in moist paper towels as rhizome pieces without any soil present. In both Trebon and Moussac, rhizome pieces were transplanted into the common gardens approximately one week after they were prepared. Some early mortality may have resulted because of the lag between plant preparation and planting into the gardens.