Impacts of a Flood Pulsing Hydrology on Plants and Invertebrates in Riparian Wetlands

Home , Deposition (geology), Flood pulse concept, Hydrograph, Spring (hydrology)

IMPACTS OF A FLOOD PULSING HYDROLOGY ON PLANTS AND INVERTEBRATES IN RIPARIAN WETLANDS

A dissertation submitted to Kent State University in partial fulfillment of the requirements for the degree of Doctor of Philosophy

Maureen K. Drinkard

August 2012

Dissertation written by Maureen K. Drinkard B.S., Kent State University, 2003 Ph.D., Kent State University, 2012

Approved by

___Ferenc de Szalay_, Chair, Doctoral Dissertation Committee

___Mark Kershner______, Members, Doctoral Dissertation Committee

_____Oscar Rocha______,

____Mandy Munro-Stasiuk_,

Accepted by

_____James Blank______, Chair, Department of Biological Sciences

______Raymond Craig___, Dean, College of Arts and Sciences

TABLE OF CONTENTS

LIST OF FIGURES ...... vi

LIST OF TABLES ...... vii

ACKNOWLEDGEMENTS ...... x

CHAPTER

I. INTRODUCTION ...... 1 Dissertation Goals ...... 1 Definition of the Flood Pulse Concept ...... 2 Ecological and economic importance ...... 3 Impacts of environmental characteristics in riparian wetlands ...... 5 Dissolved oxygen and desiccation ...... 5 Chemical conditions ...... 6 Sedimentation, erosion and deposition ...... 6 Effects of flood pulsing on biota ...... 7 Plant adaptations and responses to abiotic stresses ...... 7 Flooding, dissolved oxygen and desiccation ...... 7 Erosion and deposition ...... 9 Invertebrate responses to abiotic stresses ...... 10 Flooding, dissolved oxygen and desiccation ...... 10 Erosion and deposition ...... 17

II. USING MESOCOSM EXPERIMENTS TO TEST RESPONSES OF PLANT AND INVERTEBRATES TO FLOOD PULSING Abstract ...... 19 Introduction ...... 20 Methods ...... 24 Study site ...... 24 Physicochemical measurements ...... 28 Plant and algae communities ...... 29 Aquatic invertebrates ...... 31 Statistical analyses ...... 32 Results ...... 33 Abiotic conditions ...... 33 Aquatic invertebrates in the permanent pool ...... 39 Plant communities in the permanent pool ...... 43 Plant communities in the Intermittently Flooded Zone ...... 43

iii

Discussion ...... 54 Impacts of flooding in the IFZ...... 54 Impacts of flooding in the permanent pool ...... 57 Management implications ...... 59 Acknowledgements ...... 61 Appendix A ...... 62

III. ZONATION OF PLANT COMMUNITIES CAUSED BY HYDROLOGICAL STRESSES IN HEADWATER RIPARIAN WETLANDS Abstract ...... 69 Introduction ...... 71 Methods ...... 74 Study site ...... 74 Design of hydrology treatments and plant sampling ...... 75 Statistical analysis ...... 77 Results ...... 79 Abiotic conditions ...... 79 Plant community responses ...... 84 Discussion ...... 119 Zonation responses in the plant community ...... 119 Management implications ...... 124 Acknowledgements ...... 127

IV. USING MESOCOSMS TO TEST IF FLOOD PULSING AFFECTS EMERGENCE OF AQUATIC INSECTS IN HEADWATER WETLANDS Abstract ...... 128 Introduction ...... 129 Methods ...... 132 Study site ...... 132 Experimental design ...... 134 Hydrology ...... 134 Adult insect emergence ...... 135 Statistical analysis ...... 136 Results ...... 137 Hydrology ...... 137 Invertebrate communities ...... 140 Multivariate community analysis ...... 145 Discussion ...... 156 Overall implications ...... 163 Acknowledgements ...... 165

V. THE IMPORTANCE OF COARSE WOODY DEBRIS AS INVERTEBRATE MICROHABITAT IN HEADWATER FLOODPLAINS Abstract ...... 166 Introduction ...... 168 Study objectives and hypotheses...... 172 Methods ...... 173 Study site ...... 173 Experimental design ...... 174 Statistical analysis ...... 176 Results ...... 177 Soil environmental conditions ...... 177 Plant community responses ...... 178 Invertebrate community analysis ...... 181 Discussion ...... 192 Acknowledgements ...... 198

VI. SYNTHESIS AND DISCUSSION Synthesis ...... 199 Impact of flood pulsing on biota in headwater wetlands ...... 201 Positive impacts of flood pulsing ...... 201 Negative impacts of flood pulsing...... 206 Using mesocosms to test large-scale ecological paradigms ...... 208 Overall conclusions ...... 212

LIST OF FIGURES

CHAPTER 1. INTRODUCTION

CHAPTER 2: USING MESOCOSM EXPERIMENTS TO TEST RESPONSES OF PLANT AND INVERTEBRATES TO FLOOD PULSING

Figure 1. Cross-section along the longitudinal axis of a Static Wetland and a Flood Pulse Wetland...... 26

Figure 2. Hydrograph showing mean water levels in all mesocosms at HAERF from 2003-2005 ...... 34

Figure 3. NMDS ordinations of plant communities in Flood Pulse and Static Wetlands with MRPP statistics...... 45

Figure 4. Plant community in Flood Pulse and Static Wetlands 2003-2005 . 48.

CHAPTER 3: ZONATION OF PLANT COMMUNITIES CAUSED BY HYDROLOGICAL STRESSES IN HEADWATER RIPARIAN WETLANDS

Figure 5. Mean relative water levels in Flood Pulse and Static Wetlands in 2005 and 2006 ...... 80

Figure 6. Mean total biomass in the treatment/elevation zones in 2005 .... 98

Figure 7. Mean percent bare ground in the treatment/elevation zones in 2005 and 2006 ...... 100

Figure 8. Mean species richness in the treatment/elevation zones in 2005 and 2006 ...... 102

Figure 9. Plant community type in the treatment/elevation zones in 2005 and 2006 ...... 104

Figure 10 NMDS ordinations of plant communities in the treatment / elevation zone habitat types ...... 106

CHAPTER 4: USING MESOCOSMS TO TEST IF FLOOD-PULSING AFFECTS EMERGENCE OF AQUATICINSECT IN HEADWATER WETLANDS

Figure 11. Water levels in 2005 and 2006 at Mud Brook Preserve and at the Herrick Aquatic Ecology Research Facility ...... 138

Figure 12. Invertebrate community numbers are HAERF and MBP ...... 141

Figure 13. Results of RM-ANOVAS comparing total abundance and species richness at HAERF in 2005 and in HAERF and MBP in 2006...... 146

CHAPTER 5: THE IMPORTANCE OF COARSE WOODY DEBRIS AS INVERTEBRATE MICROHABITAT IN HEADWATER FLOODPLAINS

Figure 14. Soil environmental conditions below logs and in control areas at Mud Brook Preserve in 2006 and 2007 ...... 179

Figure 15. Invertebrate communities collected in pitfall traps at Mud Brook Preserve. A. Species richness per trap (Mean ± 1 SE) ...... 182

Figure 16. NMDS analysis of invertebrate communities in September 2006, May 2007, and September 2007 ...... 184

CHAPTER 6: SYNTHESIS AND DISCUSSION

vii

LIST OF TABLES

CHAPTER 1. INTRODUCTION

CHAPTER 2: USING MESOCOSM EXPERIMENTS TO TEST RESPONSES OF PLANT AND INVERTEBRATES TO FLOOD PULSING

Table 1. Percent of time mesocosms were at different water levels in each treatment. . 37

Table 2. Comparisons of aquatic invertebrate communities in Flood pulse (FP) and Static (ST) wetlands using Repeated Measures ANOVA ...... 40

Table 3. Life history traits for the emergent plant community in the IFZ of Flood pulse (FP) and Static (ST) wetlands using Repeated Measures ANOVA ...... 50

CHAPTER 3: ZONATION OF PLANT COMMUNITIES CAUSED BY HYDROLOGICAL STRESSES IN HEADWATER RIPARIAN WETLANDS

Table 4. Student’s t-test results showing mean [SE] percent of time the elevation zones in each treatment ...... 82

Table 5. Mean [SE] percent cover of the common taxa in 2005 and 2006 in each treatment / elevation zone habitat type...... 85

Table 6. 2-way ANOVAs comparing communities characteristics in treatment / elevation zone habitat types ...... 108

Table 7. Results of pairwise comparisons of plant communities with MRPP tests in 2005 and 2006...... 112

Table 8. . Indicator species in each treatment / elevation zone habitat type...... 115

CHAPTER 4: USING MESOCOSMS TO TEST IF FLOOD-PULSING AFFECTS EMERGENCE OF AQUATICINSECT IN HEADWATER WETLANDS

Table 9. Results of RM-ANOVAS comparing total abundance and species richness at HAERF in 2005 and in HAERF and MBP in 2006 ...... 143

Table 10. Abundance of dominant taxa at Herrick Aquatic Ecology Research Facility and Mud Brook Preserve in 2005 and 2006...... 149

viii

Table 11. P-values of MRPP pair-wise comparisons between habitats at HAERF and MBP...... 152

Table 12. Indicator species of the habitat types at HAERF and MBP in 2006 ...... 154

CHAPTER 5: THE IMPORTANCE OF COARSE WOODY DEBRIS AS INVERTEBRATE MICROHABITAT IN HEADWATER FLOODPLAINS

Table 13. Comparisons of invertebrate community at the four habitat types at HAERF and Mud Brook, Ohio ...... 186

Table 14. Indicator Species Analysis for the four habitat types at HAERF and Mud Brook in 2006 and 2007 ...... 189

CHAPTER 6: SYNTHESIS AND DISCUSSION

AKNOWLEDGEMENTS

I dedicate my dissertation to my mother, Pamela, for her steadfast strength and love and unimaginable commitment to me and to every dream I ever dreamt. Also, to my father,

Leonard, for instilling in me, by example, courage, a sense of capability and unabashed adventure. To my earth mother, Elizabeth, for her sensibility and loving perspective, to my many siblings and few friends for their unending support, constant encouragement and the occasional stiff drink. To my advisor, Dr. Ferenc de Szalay, who taught me that I was better than a B. He taught me to love the pursuit of knowledge and gently led me to the gift of teaching. He also taught me that, at times, a killer fashion sense and emphatic jazz hands do matter. Finally, I dedicate my years of doctoral research to my three beautiful and brilliant children, Evyn

Elizabeth, Maxwell Vincent and Isadora Bijoux. In addition to being my very best field companions and late night desk mates, they have been an endless fountain of joy and a source of strength to go on when the pages of this manuscript seemed too far in the distance. They allowed me the ability to trudge forward through swamps when I just couldn’t take one more soil sample or pull one more sled. In them, I found the energy to stay awake for just “one more hour” to finish identifying yet another unknown maggot or tiny unintelligible brown beetle. And now, because of Evy, Max and my little Bijoux, I mustered the ability to put words on these pages even on days when there were no words left.

In the process of getting my doctorate, I acquired much more than an expert level of knowledge about headwater wetland ecology. I leave my program more worldly, more

x empathetic, and tough in ways I didn’t imagine were possible. I am infinitely grateful that I never knew what I was getting myself into for if I had known I most likely would have chosen something less challenging, but no doubt, less fulfilling.

CHAPTER 1

INTRODUCTION

Dissertation Goals

My dissertation examines how flood pulsing in headwater riparian wetlands impacts invertebrate and plant communities. For this, I have conducted field and laboratory projects using descriptive and manipulative experimental designs. In my first study, I used wetland mesocosms at the on-campus Herrick Aquatic Ecology Research Facility

(HAERF) to compare how flood pulsing and static water levels affect diversity, abundance, biomass, and life history traits of wetland and upland plants and aquatic invertebrates. My second study examined how Flood pulsing determines plant zonation using mesocosms at HAERF. In my third study, I examined how diversity, abundance,

and life history traits of emerging adult invertebrates were affected by flood pulsing hydrology in both natural flood-plain wetlands at Mud Brook Preserve (MBP) and mesocosms at the HAERF. Finally, my fourth study tested how coarse woody debris in riparian floodplains affects the physical environment and concomitant invertebrate communities.

Definition of Flood Pulse Concept

The Flood Pulse Concept (FPC) describes the ecological impacts of the hydrological linkages between river and floodplain habitats (Junk et al. 1989). Flood pulsing occurs when heavy rains or snowmelts cause rivers to overflow their banks and flood adjacent riparian habitats. As a result, water levels in riparian wetlands are high during flooding events but gradually decrease due to drainage and evapotranspiration. This variable hydrology creates a mosaic of habitat types within the floodplain which include: (a) permanently flooded lotic habitats in the river channel (b) permanent wetlands in depressions with impermeable soils (c) intermittently flooded wetlands at higher elevations and (d) terrestrial habitats in higher elevations that receive only short-term and sporadic inundation during extreme floods (Naiman and Decamps 1997, Lamberti et al. 2010).

Moreover, the FPC emphasizes that rivers, floodplain wetlands and uplands are integrated components of a single system. The holistic approach of the FPC is useful 3

when studying the ecology of riparian wetlands unlike isolated wetlands (e.g. bogs, vernal pools), riparian wetlands are strongly impacted by the adjacent lotic and terrestrial habitats (Lamberti et al. 2010, Malard et al. 2006). For example, organic

(organisms, detritus) and inorganic matter (water, nutrients, sediments) are exchanged between the river channel habitat and permanent pools and the intermittently flooded floodplain and terrestrial habitats during flood pulse events (Bengino and Sommer 2008,

Polis et al. 1997, Steiger et al. 2005, Thoms 2003). When population levels in one habitat type are largely supported by energy inputs from a second habitat type, the external input is defined as "resource subsidy.” Resource subsidies can alter community interactions (e.g., interspecies competition) and trophic structures (Polis et al. 1997,

Rundio and Lindley 2012). Flood pulsing facilitates reciprocal resource subsidies between riparian wetlands and riverine habitats, and therefore may determine productivity of key species in these interconnected habitats.

Ecological and Economic Importance

Across the globe, the amount and integrity of riparian systems have been reduced by anthropogenic factors such as altered flow regimes, diking, and pollution from urban or agriculture land use. For example, flow regimes in 77% of the large river systems in the United States and Canada have been altered by diking and damming (Dynesius and

Nilsson 1994). Impacts to the hydrologic regime have caused extensive changes to 4

habitat conditions (Mensing et al. 1998, Quinn et al. 2000, Robertson et al. 2001,

Sheldon et al. 2002). Wetlands provide many economically valuable ecosystem services including improved water quality by sequestering chemical inputs. Alteration of hydrologic regimes can adversely alter chemical and nutrient cycles in riparian wetlands and other important ecosystem functions (Costanza et al. 1998, Pinay et al. 2002,

Thoms 2003).

Unimpounded riparian systems have higher productivity than isolated systems

(Bunn and Arthington 2002, Ward and Stanford 1995, Ward 1999). This may be due to changes in flood timing and duration, or fluvial processes such as erosion or deposition

(Adis and Junk et al. 2002). For example, river fish in intact riparian systems feed on the abundant plant and animal food resources in the floodplain during key periods of their lifecycles (e.g. reproduction, growth of juveniles). Junk and Welcomme (1985) found that Amazon River fish grew mostly during the rainy season when they moved into the floodplain (i.e., these river fish relied on resource subsidies from floodplain wetlands). Also, alternating wet and dry periods in intermittently flooded wetlands increases nutrient release from decomposing detritus, which results in high primary and secondary productivity (Molles et al. 1998). For example, 67-95% of aquatic invertebrate production in a riparian system in the southeastern United States occurred in floodplain wetlands, and only ~1% was from the river channel (Smock 1994). This indicates that terrestrial animals (e.g. birds) that feed on invertebrates in riparian zones would also be impacted by factors that alter floodplain productivity. 5

Biodiversity is also higher in riparian wetlands than other types of wetlands

(Bornette et al. 1998, Harper et al. 1997, Tockner and Ward 2000). High biodiversity is in part due to the overlap of riverine, wetland and terrestrial species (Junk 2006, Merritt and Cummins 2008). Furthermore, fluvial processes such as erosion and deposition create a mosaic landscape associated with high biodiversity (Batzer et al. 1999, Huryn and Gibbs 1996, Tockner et al. 2000). A loss of connectivity can lead to decreases in biodiversity and increases of exotic species in riparian wetlands (Bunn and Arthington

2002, Kozlowski 2002, Pringle 2001, Tockner et al. 1999b, Ward et al. 1999).

Impact on Environmental Characteristics in Riparian Wetlands

Dissolved Oxygen and Desiccation

Riparian wetlands cycle between wet and dry periods, and therefore, oxygen levels are highly variable in the water column and the wetland soils. Dissolved oxygen levels are generally low in flooded wetlands because O2 diffusion is ~10,000 times slower through the water column than the atmosphere, and available O2 is quickly used up by microbial decomposition of the abundant organic matter (Mitsch and Gosselink

2007). However, wetland dissolved oxygen levels may temporarily increase when river flooding flushes oxygenated water into the floodplain. Intermittent desiccation in floodplains during draw downs is another stress for aquatic species. Therefore, the 6

biota in these habitats must be mobile or able to tolerate the stresses of a habitat that alternates wet and dry periods (Batzer et al. 1999, Benke 2001, Blom and Voesenek

1996).

Chemical Conditions

Flooding will also input dissolved and particulate matter that alters soil and water chemistry. Flooding may bring in chemicals and nutrients, which become concentrated when water levels drop. As dissolved oxygen becomes depleted, some of these chemicals become toxic as they are reduced (e.g. ammonia, hydrogen sulfide, ferrous iron). Many wetlands act as sinks for sediments and nutrients (e.g. carbon, nitrogen).

This can be negatively affected by a loss of natural hydroregime (Thoms 2003, Tockner et al. 1999a).

Sedimentation, Erosion and Deposition

Strong floods also erode wetland soils or deposit organic debris (leaves, coarse woody debris) and sediment (sand and gravel) from the main channel. Erosion creates scour holes and meanders in the floodplain. Sediment deposition can range from millimeters to several meters annually and is dependent upon local watershed land use and climate (Naiman and Decamps 1997). While these factors usually add stressors to 7

the biota, they also increase overall habitat heterogeneity (Ward and Stanford 1995,

Steiger et al. 2005, Alsfeld et al. 2009), and they can provide nutrient resources for wetland biota (Junk et al. 1989, Ward et al. 2002).

Effects of Flood Pulsing on Biota

The abiotic factors (mentioned above) will have a strong impact on plant and invertebrate communities in floodplain wetlands. While flood pulsing can be beneficial to the biota (e.g. increased nutrient availability, moisture), it also creates conditions of anoxia, turbidity, conductivity, harmful chemical compounds, erosion and deposition that stress organisms (Hillman and Quinn 2002). Floodplain organisms survive these conditions with morphological, behavioral and life history adaptations (Junk et al.

1989). Below I outline many of the important adaptations of plants and invertebrates to the stresses of flood pulsing.

Plants Adaptations and Responses to Abiotic Stresses

Flooding, Dissolved Oxygen and Desiccation

Soil anoxia is an adverse stressor for wetland plants because it prevents cell metabolism in the roots, thereby blocking moisture and nutrient uptake. Some species 8

such as Spartina alternifolia (Mitsch and Gosselink 2007) use anaerobic metabolic pathways (Kennedy et al. 1992), but this is ~19x less efficient than aerobic respiration.

Therefore, most plants have other adaptations to survive in anoxic soils. The vine,

Milkens scandens, increases oxygen diffusion through its stem to its roots by increasing aerenchyma tissue and stomata (Moon et al. 1993). Other plants such as sedges

(Cyperacea), Phragmites spp. and willows, Salix spp., use adventitious roots on their stem to facilitate oxygen diffusion to roots (Moog and Janiesch 1990). Other plants including Typha spp. and Spartina spp. transport oxygen to their roots that diffuses into the soil to create an oxidized rhizosphere (Mitch and Gosselink 2002).

Seedling survival and subsequent community structure is strongly determined by soil submergence (Gurnell et al. 2007). When flood regimes are predictable, some annual plants avoid stressful periods by surviving the flooded season as dormant seeds

(Mitsch and Gosselink 2007). In these habitats, flood events are usually most common in winter to early spring when most perennial plants are dormant (Mitsch and Gosselink

2007). A study by Gurnell et al. (2007), showed that drained soil samples had higher seed germination, lower seedling mortality and greater plant diversity than waterlogged and submerged samples. Most dormant wetland seeds can survive desiccation (~92% of 8,000 species) (Tweedle et al. 2003), but seedlings are usually intolerant of dry soils.

Therefore, many plants time their life cycle so that the seeds germinate and grow quickly when conditions are most favorable (Johansson and Nilsson 2002, Kozlowski

2002). In habitats with unpredictable flood regimes, plants respond to occasional 9

flooding by delaying flowering or seed production until floods recede (Van der Sman et al. 1993).

Erosion and Deposition

Riparian wetland plants are greatly affected by hydrogeomorphic processes, including erosion and deposition (Naiman and Decamps 1997, Tabacchi et al. 1995).

Sediments from the river channel settle out with decreasing flow velocity and can affect the seed, seedling and adult stages (Gleason et al. 2003, Goodson et al. 2001,

Kozlowski 2002). Sediment deposition alters light penetration, oxygen distribution and soil properties including bulk density, moisture, salinity, and nitrogen (Byrd and Kelley

2006, Mahaney et al. 2004). Riparian vegetation with an extensive root system can affect sediment retention by decreasing flow velocity during flood events, and can reduce erosion (Mitsch and Gosselink 2007).

The impacts of sedimentation and erosion will alter wetland community structure. Most studies have shown that sedimentation will decrease seedling numbers

(Dittman and Neely 1999), bury seeds and reduce germination (Jurik et al. 1994). Seed burial depth of only 0.5 cm can reduce germination by 91.7% (Gleason et al. 2003).

Seeds that are resistant to sedimentation are usually large (Dittmar and Neeley 1999,

Gleason et al. 2003 Jurik et al. 1994). For example, anthropogenic sedimentation in a salt marsh caused an increase in opportunistic cattails (Byrd and Kelly 2006). Scouring 10

will also remove and kill seedlings, but the open bare soils it produces provide areas for seed establishment (McBride and Strahan 1984).

Sedimentation of mature plants usually causes a loss of growth due to reduction of oxygen to roots. Furthermore, substantial reductions in decomposition occur when detritus is buried, which can reduce nutrient availability for growing plants (Vargo et al.

1998). However, responses to sedimentation are species dependant and compounded by other variables including nutrient limitations (Mahaney et al. 2004). Riparian trees along ephemeral streams showed significant losses in root biomass and root nutrient content with as little as 0.3 cm of deposited sediments (Cavalcanti and Lockaby 2005).

In contrast, negligible losses and minor increases in plant biomass were detected with

~2cm of deposited sediments (Koning 2004).

Invertebrate Responses to Abiotic Stresses

Flooding, Dissolved Oxygen and Desiccation

Aquatic insects have different types of respiratory systems that are beneficial in both river and wetland habitats in floodplains (Merritt and Cummins 2008, Tronstad et al. 2005a). Many beetles and hemipterans use air stores or a plastron to carry air bubbles below the surface, and mosquitoes and other fly larvae, such as rat-tailed maggot, Eristalis tenax, use siphons to breathe atmospheric air at the water's surface. 11

Mansonia mosquito larvae have spine-like abdominal siphons that pierce the aerenchyma of cattail to exploit interstitial oxygen. Other aquatic taxa have cutaneous respiration to absorb dissolved oxygen across their cuticle or external gills

(Ephemeroptera, damselflies). Other species are specialized for the anoxic stresses in wetlands such as chironomid midge larvae that use hemoglobin in their blood to store oxygen. Behavioral adaptations to low dissolved oxygen includes increasing water flow, and therefore contact with oxygen, with body undulations (Odonata: Damselflies).

Floodplain wetland invertebrates are often generalists that tolerate broad ranges of dissolved oxygen (Batzer and Wissinger 1996), but there are usually community shifts in the most hypoxic wetland environments (Spieles and Mitsch 2003).

In floodplains, aquatic wetlands invertebrates must tolerate intermittent desiccation after the flood waters recede (Williams 1997). Studies have shown that key characteristics of floodplain hydrology are timing, duration, and predictability of flooding events. For example, riparian wetlands with predictable flood periods are inhabited by invertebrate taxa with morphological, physiological, and behavioral adaptations to avoid or survive flood stresses (Batzer and Wissinger 1996). Whereas, invertebrate populations in habitats with unpredictable flood periods often are killed off during flooding, but they can quickly rebuild population numbers with high fecundity or rapid recolonization (Adis and Junk 2002, Naiman and Decamps 1997).

Mobile invertebrates (e.g. winged insects) can escape to permanent habitats during draw-downs. Invertebrates that do not migrate must survive desiccation with 12

various morphological, behavioral and life history adaptations (Batzer and Wissinger

1996, Bunn and Arthington 2002). Much like plant seed banks, invertebrate "egg banks" comprised of dormant or in resistant life stages are sources for recolonization after habitats are reflooded (Brock et al. 2003, Dietz-Brantley et al. 2002, Tronstad et al.

2005b). These egg banks can remain dormant through numerous flooding and drying cycles for years (Brock et al. 2003, Trondstad et al. 2005b). In egg bank studies, frequency and duration of flooding was positively correlated with the number and diversity of invertebrates emerging after reflooding, but semi-aquatic and terrestrial species had more variable responses to rehydration than aquatic species (Dietz-Brantley et al. 2002, Tronstad et al. 2005b).

Changes in flood regime have caused invertebrate communities to respond unexpectedly to flood events. For example, in several reaches of a river system in

Australia, unimpacted reaches showed community level changes in composition in response to flood events whereas communities in impacted reaches showed little or no response to flood events (Quinn et al. 2000).

Flood permanence in a floodplain varies with geomorphic characteristics, and this affects invertebrate community structure. Deeper areas are flooded the longest and invertebrates in these microhabitats tend to be strongly affected by infrequent draw-downs (Jeffries 1994). Droughts in deep wetlands cause high reductions in species diversity and abundance, but over time population dynamics become strongly dependent upon interspecies interactions such as competition, predation, reproduction 13

(Jeffries 1994). For example, there is a general pattern of invertebrate and vertebrate predators recolonizing more slowly than their prey when wetlands are reflooded.

Higher than expected invertebrate numbers in newly flooded wetlands have been attributed to low numbers of predators (Dietz-Brantley et al. 2002), which leads to increased intraspecific and interspecific competition for resources (Van de Meutter et al. 2005). As a result, predation may have greater effects on invertebrate community structure than flood pulsing in permanently flooded riparian wetlands (Batzer and

Wissinger 1996, Corti et al. 1997).

In habitats that receive a shallow flood, taxa can survive frequent draw-downs with resting stages; however prolonged periods of high water can eliminate some taxa

(Batzer and Wissinger 1996). Taxa that inhabit intermittently flooded wetlands are rapid and cyclic colonizers (Batzer and Wissinger 1996). Flight polymorphisms are common in some taxa (Hemipterans and Beetles) that over-winter in permanent habitats, and fly to temporary habitats during flood phases to breed (Batzer and

Wissinger 1996). After aerial colonizers arrive at a wetland they use environmental variables as colonization cues. Furthermore, the incidence of adult egg laying can be affected by habitat characteristics such as depth and plant cover (Stagliano et al.. 1998).

For example, invertebrate communities that colonized areas of high plant cover were more diverse but had lower biomass than those that colonized areas of open water (de

Szalay and Resh 2000). 14

Larval development rates and adult insect emergence will vary temporally and spatially within a wetland (Stagliano et al. 1998, Whiles and Goldowitz 2001, Whiles and

Goldowitz 2005). While aquatic chironomid larvae can following receding water levels and emerge later as adults from deeper areas, semi-aquatic and terrestrial chironomid species will emerge soon after flooding from inundated soils (Tronstad et al. 2005a).

The highest biomass of emerging insects was associated with the longest flooding periods (i.e. 296-365 days per year), but highest diversity was correlated to intermediate flood periods (Whiles and Goldowitz 2001, Whiles and Goldowitz 2005).

Emerging insect biomass is also negatively correlated to distance from flood source, and therefore, flood frequency and duration (Mackenzie and Kaster 2004). As temporary wetlands begin to dry, some taxa increase developmental rates by selecting more nutrient rich food resources. For example, caddisflies (Trichoptera: Asynarchus) switched to cannibalism resulting in faster developmental rates (Wissinger et al. 2004).

Moreover, changes in abiotic cues (turbidity, temperature, salinity, conductivity) that are associated with draw-downs can increase the amount of emergence (Batzer and

Wissinger 1996). Insects that emerge from wetlands can be specifically associated with particular hydrology regimes and invertebrate community tropic guild structure is strongly linked to flood frequency (Whiles and Goldowitz 2005).

Coarse woody debris (CWD) from fallen trees is an important structural component of riparian floodplains. CWD provides complex habitat structure, reduces 15

erosive forces of floods and provides physical and microclimatic refugia for organisms

(Dechene and Buddle 2010, Naiman and Decamps 1997).

In large river systems, CWD is an important invertebrate habitat because it enhances food resources for detritivores, and provides a refuge from predators and desiccation stresses during draw downs (Harper et al. 1997, Tockner et al.. 2002).

However, the ways in which biotic community structure interacts with CWD in intermittently flooded headwater wetlands has been less studied. Additionally, human impaction of river systems can cause a reduction in connectivity, which can result in a concomitant loss in CWD input and movement across the floodplain (Gurnell et al.

2002).

Physical attributes of frequent or high magnitude flood events (erosion, sedimentation, rapid changes in water chemistry) can reduce the abundance and diversity of invertebrates (Uetz et al. 1979). CWD is a natural occurring mediator of these effects. Under intermediate flood disturbances small scale spatial refuges (like

CWD) can help to mediate some of the stressors caused by Flood pulsing (Harper et al.

1997, Pollock et al. 1998). CWD can act as a refuge to both semi-aquatic and aquatic invertebrates that become stranded on drying flood plains following a flood event. Soil conditions under CWD can be cooler, have higher organic content and can retain more moisture than the soils in the immediate area that are not associated with CWD (Gurnell et al. 2002). These conditions are conducive to aquatic and semi-aquatic invertebrates that need to aestivate or desiccate until the flood plain becomes re-flooded. Some 16

mosquitoes, (e.g. Aedes), are dependent on flooding and drying cycles to stimulate life cycle changes and thrive under these conditions. However, many inhabitants of flood plains can only tolerate the harsh condition by seeking refugia, aestivating and desiccating. In many cases, CWD provides the necessary refugia for these organisms

(Isopoda, Cladocera, etc.) in large order systems (Harper et al. 1997).

Furthermore, Pollock et al. (1998), showed that plant species richness and productivity was highest in areas that had intermediate flood frequency and CWD.

Increased abundance and diversity in plants and invertebrates are attributed to increased habitat heterogeneity caused by CWD (Vivian-Smith 1997). In addition, the physical structure of CWD can act to slow flows across flood plains, which is also thought to reduce the stress of flood events (Gurnell et al. 2002). It is in this capacity that CWD is thought to partially mediate some negative effects of flood pulsing.

Although most wetland invertebrates have extensive physiological, behavioral and morphological adaptations to overcome the stresses of flooding, species assemblages of semi-aquatic and terrestrial floodplain soil invertebrates can be affected by flood events (Plum 2005). Following flood events the soil fauna shows a rapid community shift to dominance by well-adapted and opportunistic species (Plum 2005).

While flood events can cause a decline in their populations, recolonization is rapid from diapausing individuals. For example, terrestrial earthworms (Lumbricidae) and potworms (Enchytraidae) greatly contribute to soil processes are common in floodplain soils and their populations are strongly linked to flood regime. Survival of worms has 17

also been related to soil type with higher survival in mineral soils than peat soils (Plum and Filser 2005).

Erosion and Deposition

Like wetland plants, deposition of sediments and organic debris can also bury sessile benthic invertebrates (Gleason et al.. 2003). Invertebrate egg banks can be decreased by sedimentation as low as 0.5 cm (Euliss and Mushet 1999, Gleason et al.

2003). However, the juveniles and adult stages are often mobile and less affected by erosion or sedimentation. Therefore, many taxa have life cycles timing which enable them to avoid stressful periods in which their eggs would be exposed to erosion or deposition (Batzer and Wissinger 1996). Increased turbidity can also affect invertebrate communities by decreasing hunting efficiency of fish or invertebrate predators (e.g. Odonata: Zygoptera) (Timms and Boulton 2001, Trebitz et al. 2007, Van de Meutter et al. 2005).

It is important to point out that most of the studies mentioned above were conducted in large-order river systems (e.g., The Amazon, Mississippi, and Rhine Rivers), and much less is known about the effects of Flood pulsing in headwater ecosystems (Freeman et al. 2007). Large rivers tend to remain within their deeply incised channels but flooding events during the rainy season or after spring thaw can last up to several months

(Tockner et al. 2000). As a result, plants and invertebrates often have predictable life- 18

cycles with periods of reproduction and emergence timed to avoid the most stressful periods (Adis and Junk 2002, Johansson and Nilsson 2002). In contrast, flooding in headwater streams is shorter and less predictable due to smaller catchment size, lower bank height, and a narrower stream channel (Bayley 1995, Tockner et al. 2000). This limits the time that aquatic invertebrates can access resources in the floodplain and makes desiccation stress more unpredictable. Achieving a better understanding of the ecology of headwater stream systems would be invaluable because headwaters comprise approximately 2/3 of the total length of river systems (Dynesius and Nilsson

1994, Freeman 2007). Therefore, the ecology of headwaters creeks will strongly impact the ecology occurring downstream in the large rivers (e.g. by providing organic matter, see Vannote et al. 1980).

CHAPTER 2

USING MESOCOSM EXPERIMENTS TO TEST RESPONSES OF PLANTS AND

INVERTEBRATES TO FLOOD PULSING IN HEADWATER WETLANDS

1Abstract

Flood pulsing is well-known to be a key environmental factor that structures biotic communities in large-order river systems, but this study focuses on the effects of flood pulsing in headwater systems. I used 10 mesocosm wetlands (10 m x 20 m) to test two treatments: a Flood pulse regime in which natural flood events caused water levels to fluctuate and a Static regime in which water levels remained stable. Abiotic characteristics, plants and aquatic invertebrate communities were monitored from 2002 through 2005. The Flood pulse treatment had minimal impacts on environmental

1 Drinkard, M.K., M.W. Kershner, A. Romito, J. Neiset, and F.A. de Szalay. 2011. Responses of plant and invertebrate assemblages to water-level fluctuation in headwater wetlands. Journal of North American Benthological Society 30: 981-996.

conditions of permanent pools, and submersed plant and aquatic invertebrate communities were similar in each treatment. This suggested that there were minimal resource subsidies from the floodplain into the pools. However, flood pulsing caused pronounced changes to plant communities in the intermittently-flooded zone (IFZ) above the stable water line. Overall plant diversity was higher in Static wetlands, and % bare ground was higher in Flood Pulse wetlands suggesting that the short, stochastic floods were a strong environmental stressor. In Flood Pulse wetlands, plants had to survive frequent submergence, which also reduced introduced, weedy, and upland plant taxa. The NMDS ordination found distinct plant communities in each treatment indicating that the abiotic stresses caused pronounced changes in the floodplain community. The NMDS identified many obligate wetland plants as indicator taxa in

Flood Pulse wetlands (e.g. Juncus canadensis, Ludwigia palustris, Dulicheum arundinaceum, Eleocharis obtusa, Carex crinita, C. lupulina, C. vulpinoidea) but many

Facultative or Upland species were indicators in Static wetlands (Cirsium arvense,

Eupatoriadelphus maculatus, Plantago lanceolata, Bidens frondosus, Melilotus officinalis, and Mentha arvensis, Daucus carota, Poa palustris). Many of the traits that characterized plant species in the Flood Pulse wetlands (e.g. Obligate wetland taxa, perennial, native and non-weedy species) are considered beneficial from a management perspective.

Introduction 21

The Flood Pulse Concept (FPC) was developed to model the ecological linkages when rains or snowmelt cause rivers to overflow their banks into riparian floodplains

(Junk et al.1989, Odum et al. 1995). The floodplain contains an Intermittently Flooded

Zone (IFZ) with temporarily flooded wetlands and uplands and also permanently flooded pools. Studies on the FPC have emphasized that these habitats are integrated components of a single ecosystem because flood pulsing connects the river channel with the floodplain (Junk et al.1989, Bayley 1995, Odum et al. 1995, Naiman and Decamps

1997, Tockner et al. 2000, Bunn and Arthington 2002).

The physical action of flood pulsing strongly impacts abiotic conditions (e.g., soil oxygen and moisture, nutrient availability, and chemical transformations) (Junk et al.

1989). River overflow can increase sediment deposition, which can kill seedlings or benthic invertebrates (Gleason et al. 2003). As water levels recede, dissolved chemicals

(e.g. nutrients, salinity) become concentrated in floodplain pools (Mitsch and Gosselink

2007), and floodplain biota must be adapted to tolerate this variation between the wet and dry periods (Batzer and Wissinger 1996, Robinson et al. 2002). Therefore, flood pulsing can be considered a powerful environmental sieve that determines which species establish and persist in riparian wetlands (Van der Valk 1981, Chase 2007).

However, inputs of nutrients also increase resources for some wetland biota (Junk et al.

1989, Polis et al. 1997, Tockner et al. 1999, Baxter et al. 2005). Furthermore, resource subsidies from the river to the wetland can modify biotic interactions such as 22

competition (Polis et al. 1997, Baxter et al. 2005). As a result, connectivity between the river and the floodplain also controls community metrics such as biodiversity, productivity, and ecological interactions, and riparian habitats have unique abiotic characteristics and biotic communities (Smock 1994, Polis et al. 1997, Tockner et al.

2000, Ward et al. 2002).

The FPC suggests that plants in flood pulsing wetlands access increased nutrients but also experience more abiotic stresses (desiccation, anoxia, sedimentation and erosion) than in isolated wetlands (Junk et al. 1989). For example, sediment deposition during floods will alter light penetration and soil properties (Mahaney et al. 2004, Byrd and Kelley 2006), which can decrease seed germination and seedling survival (Jurik et al

1994, Dittman and Neely 1999, Gleason et al. 2003). Adaptations that are useful in riparian wetlands include desiccation tolerance, rapid growth or seed germination after floodwaters recede (Blom and Voesenek 1996, Vretere et al. 2001). Some plants use anaerobic metabolic pathways (Kennedy et al. 1992), oxidized rhizospheres with aerenchyma tissue (Moon et al. 1993, Vretare et al. 2001), and adventitious roots

(Moog and Janiesch 1990) to survive anoxic soils during flooded periods.

Invertebrates are also important components of riparian ecosystems, and they respond quickly to changes in abiotic conditions and macrophyte communities (Batzer and Wissinger 1996, Naiman and Decamps 1997, Smock 1999, Benke 2001, Malmqvist

2002). The FPC suggests that aquatic invertebrate abundance will be highest when they can exploit detrital resources (e.g., terrestrial plant litter) in the floodplain during flood 23

events (Polis et al. 1997, Huryn and Gibbs 1996, Baxter et al. 2005). Past studies in large-order riparian systems found higher invertebrate abundance and diversity in Flood

Pulse wetlands than other wetland types (Naiman and Decamps 1997, Smock 1999).

Most studies on the FPC have been conducted in large order rivers (Tockner et al. 2000), which typically experience seasonally predictable flooding that may last up to several months (Tockner et al. 2000). Less is know about how flood pulsing controls biotic community structure in headwater systems. Floods in headwater streams are less predictable and shorter in duration than in large order rivers due to differences in catchment size, bank height, and stream morphology (Tockner et al. 2000). These differences will affect ecosystem attributes (e.g. nutrient availability, leaf litter decomposition, soil and water anoxia, and erosion) that are controlled by the extent and timing of inundation (Bayley 1995, Naiman and Decamps 1997).

In this study, I used replicated mesocosm wetlands along a headwater creek to test the impacts of flood pulsing on abiotic conditions and the wetland biota. I expected that distinct abiotic conditions in Flood pulsing Wetlands and in Static

Wetlands without flood pulsing would lead to different plant and invertebrate communities. All mesocosms had a permanently flooded pool, and the Flood Pulse wetlands had a Intermittently Flooded Zone above the permanent pool. I also expected that the aquatic organisms in permanently flooded wetland pools would respond differently than communities in the Intermittently Flooded Zone (IFZ). The hypotheses were: 24

H1: Resource subsidies during floods will enhance IFZ habitat quality for wetland plants, and this will create unique emergent plant communities and lead to higher plant diversity and biomass in Flood Pulse wetlands than Static Wetlands.

H2: Exchange between the IFZ and the permanently flooded pools during floods will increase nutrient levels in the pools and allow invertebrates to access food resources in the IFZ. This will lead to higher diversity and abundance of submersed plants, algae, and aquatic invertebrates in the Flood Pulse wetlands.

Methods

Study Site

The studies were conducted at the Art and Margaret Herrick Aquatic Ecology

Research Facility (HAERF) at Kent State University (Portage Co., Ohio). The facility included ten independently flooded wetland mesocosms (total area: 2000 m2). The

HAERF was built at the edge of a ~1 ha wooded area with hardwood trees (e.g., Populus deltoides, Quercus palustris, Acer rubrum, Prunus serotina), shrubs and herbaceous plants (e.g., Viburnum recognitum, Cornus amomum, Solidago spp., Andropogon scoparius, Agrostis alba, Poa spp.). Allerton Creek flowed through the site; this 2nd- order perennial creek was ~10 cm deep during base flow, 2-4 m wide, with a ~1 km2 25

watershed. There are small (~0.1–0.3 hectare) natural floodplain wetlands imediately upstream and downstream from the facility.

Ten earthen 10 m x 20 m wetland mesocosms were excavated around a 100 m X

10 m dammed stream pool that was the water source for the mesocosms. The mesocosm sides were sloped at 0.3-m linear fall to 1-m vertical run, and maximum water depth was 1.7 m (total volume = ~2.5 X 105 l). Each mesocosm had separate inlet and outlet water control structures that used stacked PVC boards to adjust the inflow and outflow of water (Figure 1).

All mesocosms were seeded with a mix of 20 wetland plant species (Onoclea sensibilis, Carex crinita, C. vulpinoidea, C. lupulina, Eleocharis palustris, Scirpus cyperinus, S. acutus, Cyperus esculentes, Leersia oryzoides, Poa palustris, Elymus riparius, Pontederia cordata, Juncus effusus, Sagittaria latifolia, Sparganium americanum, Solidago patula, Polygonum lapathifolium, P. pennsylvanicum, Bidens cernua, Eupatorium perfoliatum; 5% by weight for each species) in June 2002 (4.5 kg seeds/basin) and March 2003 (2.25 kg seeds/ mesocosms). All species are native to

Ohio and they have a range of flooding tolerances from moist-soil to permanently flooded species.

In all mesocosms, I adjusted the height of the boards all inlet structures to be 5 cm above the height of the stream pool, which prevented water from entering during base stream flow. During moderate rain events, runoff caused the stream pool to back up

>10 cm, and each mesocosms received approximately the same amount of water 26

Fig.1. Cross-section along the longitudinal axis of a Static Wetland and a Flood pulse

Wetland at HAERF. A. is the outlet water control structure. Note that in the Static

Wetland, the stop logs are set to keep water levels ~80 cm. In the Flood pulse Wetland, additional stop logs allow water levels to rise to 140 cm, but water drains slowly back down to 80 cm through a 1.3 cm hole in one stop log. B. is the inlet water control structure, which has stop logs set 5 cm above the water level in the stream. Arrows show direction of flow during a flood events as water spills in the inlet structure and discharges through the outlet structure. See text for explanation of non-flooded zone,

Intermittently Flooded Zone and permanently flooded pool. Dimensions are not drawn to scale

until water levels receded. I randomly designated each mesocosm as either a Flood pulse Wetland (N=5) or as a Static Wetland (N=5). In Static Wetlands, I added boards in the outlet structures to maintain a depth of 80 cm in the permanent pool. Any water entering during storms spilled over these boards, and these mesocosms maintained relatively stable water levels. In Flood Pulse wetlands, I added boards in the outlet control structure to fill the wetlands to 140 cm, but I included a board with a 1.3 cm hole that slowly released water down to the 80 cm water level. Therefore, water spilling into the Flood Pulse wetlands was retained up to 140 cm in the IFZ, but it was slowly released until baseline water levels in the permanent pool were the same as in the Static Wetlands. As a result, the Static Wetlands consisted of two habitats: a permanent pool and non-flooded banks. The Flood Pulse wetlands consisted of three habitats: permanent pool, IFZ and non-flooded upper banks.

Physicochemical Measurements

Water levels in each wetland were measured about three times each week with staff gauges from May 2003 to October 2006. Basin morphology was measured to determine the amount of surface area in the floodplain and the pool of each mesocosm, and the ratio was approximately 1:1.3. Basin morphology and the water level data were also used to describe the hydrology in each treatment. I estimated the percent of time water levels were below the stable level or flooded in 3 elevations in the IFZ. The 29

IFZ was divided into three sub-zones by measuring elevations in linear distance along the bank up from the stable water line: 1) Low (from the stable water line to 0.67 m), 2)

Mid (from 0.67 m to 1.33 m), and 3) High (from 1.37 m to 2.0 m).

Physical parameters (turbidity, pH, dissolved oxygen, conductivity, and temperature) were measured with field meters (Hach Turbidimeter, Model No. 52600-

00; Oakton pHtestr; YSI Inc., Model 85) weekly from May to October in 2004, every other week in 2005, and every third week in 2006. Water samples to measure available nutrients (SRP, Nitrate, and Ammonia) were collected in June, July, August and

September of 2004 and processed on an autoanalyzer (Lachat Instruments, Inc., FIA

8000 series Milwaukee, WI.).

Plant and Algae Communities

Chlorophyll a levels of periphyton were measured in June, August and October

2004. On each sample period, I attached 8 microscope slides on a wooden stake at 60 cm from the bottom of the wetland. Slides were collected after two weeks and wrapped in foil and kept frozen in the laboratory. I scraped the slides with a razor blade into a vial with methanol, and measured chlorophyll a levels with a flourometer (Turner

Designs TD-700) following standard procedures (EPA Method 445.0, Arar and Collins

1997). 30

Emergent plants in the Intermittently Flooded Zone were sampled from 2002–

2005. In September 2002, all plant species were catalogued in each wetland. All species were identified using published information (Boutin and Keddy 1993, Crow and

Hellquist 2000, Voss 2001, Chadde 2002), and voucher specimens were stored in the

Kent State University Herbarium. In August and October 2003, August and October

2004, and June and September 2005, I measured percent cover of emergent plants in each wetland. On each date, I placed a 1-m2 quadrat (0.5 m x 2 m; W x L) at three randomly chosen locations on the north bank of each wetland and visually quantified percent cover. The quadrat extended from the permanent pool to the upper edge of the Intermittently Flooded Zone. I also measured above-ground biomass in October

2003 and 2004 and September 2005 by clipping all plants at ground-level in one randomly chosen quadrat in each wetland and separated them by species. Biomass samples were dried (72 h at 60°C) and weighed to determine dry weight.

Emergent plants growing in the permanent pool were measured in September of

2004 and 2005. I placed the 1-m2 quadrat from the stable edge of the permanent pool towards the midline of the permanent pool at six random locations. Stem counts of all emergent plants were counted at each location. Submersed macrophytes in the permanent pool were measured in June, July and August of 2005. On each date, I sampled three random locations per wetland with a 25-cm diameter pipe embedded into the sediments. All submersed plants were removed from the pipe, rinsed of all debris and organisms, and refrigerated until they were processed in the laboratory. 31

All plants were separated to species and dried to determine biomass (72 h at

60°C). Plant percent cover data were also pooled into ecological categories by wetland indicator status (Obligate, Facultative Wetland, Facultative, Facultative Upland, and

Upland), nativity (Native species or Introduced species), weediness (Weedy and Non- weedy), anaerobic tolerance (High, Medium, Low, None), and classification (dicot or monocot) (Andreas et al. 2004, USDA website: http://plants.usda.gov)).

Aquatic Invertebrates

Aquatic invertebrates in the permanent pool were sampled in June and October

2003, and April, July, and October 2004. I used a D-frame sweep (500 µm mesh) to take

1-m sweeps at four random locations along the North edge of each mesocosm. Sweep nets were drawn parallel along the shoreline in ~30 cm water depth. Samples were collected when the permanent pools were near their maximum water levels to prevent dilution during flooding events. Sweep samples were preserved with 100% ethanol, washed in 500 µm mesh sieve, and sorted under a microscope. Large samples were sub-sampled by randomly choosing 1/4 of the material in the sieve. If the first sub- sample contained less than 200 invertebrates, additional 1/4 portions were sorted until at least 200 invertebrates were collected or the entire sample was processed.

Invertebrates were identified to the lowest possible taxon using taxonomic keys (Thorp and Covich 1991, Merritt and Cummins 2008) and then counted. Invertebrate species 32

were also pooled into ecological categories by functional feeding groups (shredders, scrapers, collectors, predators, parasites, piercers) (Thorp and Covich 2001, Triplehorn and Johnson 2004, Merritt and Cummins 2008), and the percent of total invertebrate numbers of each ecological category was calculated.

Statistical Analyses

I defined the dominant plant and invertebrate taxa as those found in at least 4 of the 10 mesocosms. Plant and invertebrate diversity were estimated by calculating

Shannon Diversity Indices (H’), Evenness (J’) and species richness (Zar 1999). Sørensen's similarity coefficients (QS) were calculated to compare the similarity between plant communities.

All data were checked were tested for normality using the Kolmogorov-Smirnov

Test. Non-normal data were transformed; count date were log (x+1) transformed and percent data were arcsine transformed (Zar 1999). Transformed data were retested for normality, and I used the data type that best approximated normality. Levels of

Chlorophyll a, physico-chemical parameters (nutrients, turbidity, pH, dissolved oxygen, conductivity, and temperature), plant data (ecological categories, % bare ground, total biomass, diversity), and invertebrate data (ecological categories, total numbers, diversity) were analyzed with repeated-measures ANOVAs for each year. If there was a

Date X Treatment interaction, I used Student’s T-tests to compare the treatments on 33

each sampling date. In each year, I also used Student’s T-Tests to compare the mean percent of time each treatment was flooded in each elevational zone (below the permanent pool water level, and in the Low, Mid and High sub-zones of the IFZ). I used a Bonferroni correction (i.e., p<0.0125 was considered significant) because I ran a separate T-tests for each of the four water depth zones. I ran the statistics using SPSS software (Edition 15.0, 2007).

I compared plant communities between treatments with ordination using non- metric multidimensional scaling (NMDS) using Sørensen's distances (PC-ORD, Version 5,

McCune and Mefford 2006). Ordinations for each sampling date were performed with percent cover data of the dominant taxa including the treatment factor as a covariate. I then identified community groups by performing multi-response permutation procedure (MRPP) (PC-ORD, Version 5, McCune and Mefford 2006) with Euclidean distance.

Results

Abiotic conditions

The Flood pulse and Static Wetlands had very different hydrologies, and patterns were similar among the three years of the experiment (Figure 2). After storms, Flood

Pulse wetlands usually were 20-50 cm above the permanent pool, and it took 4-10 days 34

Fig.2. Hydrograph showing mean water levels in all mesocosms at HAERF from May

2003 – Oct 2005. Data show changes above or below the permanent pool water line in each treatment. 35

Flood Pulse 80

Static

70 60 50 40 30 20

Change from Stable (cm) Stable from Change 10 0 -10

-20

05 03 03 03 03 04 04 04 04 04 05 05

03 03 04 04 03 03 04 04 04 04 04 05 05 05 05 05 05 05

------

J J J J J J J J

F F

A A S A S A A S

N D N D

O O O

M M M M M 36

for water levels to draw down. Static Wetlands remained near the permanent pool levels except when intense storms (<5 cm/hr) raised water levels for a few hours. Water levels were below the permanent pool for equal periods in Static Wetlands and in Flood

Pulse wetlands (T-tests; all N.S.). The length of time that water levels were within the

Low sub-zone of the IFZ were similar in Flood Pulse wetlands and in Static Wetlands (T- tests; all N.S.). However, Flood Pulse wetlands had water levels in the Mid and High sub-zones of the IFZ much more frequently than Static Wetlands (Table 1).

Although hydrology was different between treatments, abiotic conditions in the permanent pools were similar on most dates. Dissolved nutrients (nitrate, soluble reactive phosphorus), temperature, dissolved oxygen, and pH were never different between treatments (all p > 0.05). However, mean (+/- 1 SE) turbidity was slightly higher in Flood Pulse wetlands (12.9 +/- 2.3 NTUs) than Static Wetlands (8.5 +/- 1.1

NTUs) in 2005 (RM-ANOVA: F1,8 = 7.590, p < 0.05). There was a significant Date X

Treatment interaction for Conductivity in 2004 (RM-ANOVA: F19,152 = 1.991, p < 0.01).

Mean (+/- 1 SE) conductivity was moderately higher in Flood Pulse wetlands (495 +/-

11.7 µS/cm) than Static Wetlands (354.8 +/- 17.7 µS/cm) on 20 August (T-test: t1,8 =

5.389, p < 0.001), and slightly higher in Flood Pulse wetlands (448 +/- 13.3 µS/cm) than in Static Wetlands (408.8 +/- 7.1 µS/cm) on 27 August (T-test: t1,8 = 2.935, p < 0.05) in

2004.

Table 1. Percent of time mesocosms were at different water levels in each treatment.

Water depth are below the permanent pool baseline water level, and three sub-zones in the IFZ: Low (Stable water line to 0.67 m up the bank), Mid (0.67 m to 1.33 m up the bank), and High (1.33 m to 2.0 m up the bank). T-tests were used with a Bonferroni correction to give a significance value of p<0.0125.

Year Depth FP Mean [SE] ST Mean [SE] t 1,8 p

2003 Below Stable 15.64 [4.1] 21.45 [4.3] -0.9783 0.357

Low 58.06 [4.6] 76 [4.3] -2.8536 0.021

Mid 20.12 [1.5] 1.09 [0.4] 11.9711 0.000

High 4.48 [1.6] 0.24 [0.1] 2.5873 0.032

2004 Below Stable 20.32 [6.8] 23.33 [6.4] -0.3238 0.754

Low 61.90 [7.2] 75.71 [6] -1.4761 0.178

Mid 14.44 [2.4] 0.95 [0.4] 5.4925 0.001

High 3.02 [0.9] 0 [0] 3.2827 0.011

2005 Below Stable 21.88 [4.9] 30.94 [8] -0.9655 0.363

Low 59.06 [2.4] 67.65 [7.5] -1.0908 0.307

Mid 15.88 [3.7] 0.94 [0.6] 4.0333 0.004

High 2.82 [1.1] 0.35 [0.2] 2.1268 0.066

Aquatic Invertebrates in the Permanent Pool

Eighty-nine aquatic invertebrate taxa were collected in permanent pools during this study. Chironomidae were the most common taxa in both treatments, and other common taxa included Oligochaeta, Amphipoda (Taltridae, Hyalella), Ceratopogonidae,

Bivalvia (Sphaeriidae), Planorbidae (Gyraulus), Physidae (Physella), Ceonagrionidae

(Enallagma), Caenidae (Caenis), and Baetidae (Baetis). Aquatic invertebrate communities were similar in Flood pulse and Static Wetlands. The NMDS ordinations did not show any differences in invertebrate community types between flood pulsing and static wetlands. The MRPP results were all N.S.

There were a few minor differences in invertebrate community structure between treatments. In 2003, piercing insects were more abundant in Static Wetlands and shredders were higher in Flood Pulse wetlands, but they were not different in 2004

(Table 2). Scrapers were more abundant in Flood Pulse wetlands in 2004 (Table 2).

Total invertebrate abundance was not different between treatments in 2003 or 2004.

Evenness (J’) and Shannon Index (H’) were higher in Static Wetlands in 2003, but there were no significant differences in 2004. There was a treatment X date interaction for

Richness in 2004 with the number of species being higher in Static Wetlands (38.60 +/-

3.22) than in Flood Pulse wetlands (21.80 +/- 1.17) in July (T-test; t1,4= -3.505, p < 0.01).

Table 2. Comparisons of aquatic invertebrate communities in Flood pulse (FP) and Static

(ST) wetlands using Repeated Measures ANOVA. Abundance data are mean number of individuals/sweep sample. Trophic guild data are shown as percent of total.

Interaction RMA

FP Mean [SE] ST Mean [SE] F1,8 p F1,8 p

Abundance 2003 1624.30 [265.2] 1134.20 [443.2] 0.3105 0.593 0.9435 0.360

2004 3305.33 [988.2] 2497.20 [510.9] 0.1486 0.863 0.5298 0.487

Evenness J' 2003 0.54 [0.0] 0.68 [0.0] 3.0818 0.117 5.6761 0.044

2004 0.65 [0.0] 0.70 [0.0] 0.3433 0.715 0.6435 0.446

Shannon Index

H' 2003 1.69 [0.1] 1.97 [0.1] 1.8682 0.209 7.2455 0.027

2004 2.09 [0.1] 2.22 [0.1] 2.2759 0.135 1.1927 0.307

Species Richness 2003 19.40 [0.9] 17.70 [1.1] 1.8466 0.211 1.2989 0.287

2004 23.73 [1.0] 28.13 [2.5] 7.2025 0.006 4.1784 0.075

Collector 2003 13.65 [4.3] 12.19 [3.7] 0.2280 0.646 0.1480 0.711

2004 22.34 [3.2] 16.28 [2.2] 0.5010 0.615 2.8080 0.132 42

Predator 2003 15.17 [2.0] 19.95 [3.3] 2.8460 0.130 1.0910 0.327

2004 20.18 [2.7] 25.23 [3.2] 0.4690 0.513 1.1200 0.321

Parasite 2003 0.00 [0.0] 0.10 [0.1] 1.0000 0.347 1.0000 0.347

2004 0.24 [0.2] 0.07 [0.0] 0.8070 0.464 0.8070 0.395

Piercer 2003 2.02 [0.4] 7.91 [1.7] 0.0070 0.935 12.7520 0.007

2004 1.86 [0.5] 4.61 [1.2] 0.5000 0.615 3.8010 0.087

Shredder 2003 0.10 [0.0] 0.00 [0.0] 0.2770 0.613 8.3050 0.020

2004 0.16 [0.1] 0.35 [0.1] 1.4600 0.262 2.3170 0.167

Scraper 2003 32.53 [7.7] 24.46 [4.5] 0.1880 0.676 0.5730 0.471

2004 16.93 [4.1] 9.15 [2.5] 0.3160 0.589 8.2390 0.021

Plant Communities in the Permanent Pool

Plant communities in the permanent pools were similar between treatments.

There were no differences in mean (+/- 1 SE) algae chlorophyll levels (Flood pulsing wetlands: 279 µg/L (+/- 58.28); Static wetlands (318.89 µg/L (+/- 71.12); Repeated

Measures ANOVA F1,8=0.47, p=0.52; Interaction F1,8=1.13, p=0.35 ). The most common submersed plant species was Elodea canadensis in all wetlands. There was no difference in submersed macrophyte biomass in core samples (Flood pulsing wetlands:

78.73 g/m2 (+/- 43.97); Static wetlands (54.21 g/m2 (+/- 10.39); Repeated Measures

ANOVA F1,8=0.07, p=0.80; Interaction F1,8=1.92, p=0.18 ). The dominant emergent plants in the permanent pools included Sparganium americanum, and Sagitaria latifolia.

Stem counts of these species ranged from (0-416; 0-189 plants/m2, respectively). There were no significant differences of stem counts between treatments for any emergent plant species (all RM ANOVAs were N.S.).

Plant Communities in the Intermittently Flooded Zone

Emergent plant communities in the Intermittently Flooded Zone were diverse with 156 taxa collected (Appendix A). Common species included many obligate wetland species (e.g., Bidens cernua (OBL), and facultative wetland species (e.g., Juncus tenuis (FAC), Phalaris arundinacea (FACW), Scirpus cyperinus (FACW)). 44

Multivariate ordinations found clear differences between treatments. NMDS explained from 72.0% to 85.6% of the variation of plant communities (Figure 3). MRPP identified distinct plant community groups in each treatment on all dates except

October 2004. There were different indicator taxa in different years (Figure 3). In

2003, Flood Pulse wetlands were identified by presence of Juncus canadensis, Ludwigia palustris and Polygonum punctatum. Static wetlands were identified by Cirsium arvense, Eupatoriadelphus maculatum, Juncus effusus, J. tenuis, Plantago lanceolata,

Poa palustris, and Symphyotrichum lanceolatum. In 2004, indicator taxa were

Dulicheum arundinaceum, Eleocharis obtusa, Juncus tenuis, Potentilla candensis, and

Scirpus atrovirens in Flood Pulse wetlands, and Bidens frondosus, Eupatoriadelphus maculatum, Galium aparine, Melilotus officinalis, and Mentha arvensis in Static wetlands. In 2005, indicator taxa were Bidens cernua, Carex crinita, C. lupulina, C. vulpinoidea, Dianthus armeria, Polygonum hydropiperoides, Scirpus cyperinus, and

Sisyrinchium angustifolium in Flood Pulse wetlands. Indicator taxa in Static wetlands were Daucus carota, Melilotus officinalis, Panicum acuminatum, Poa palustris,

Ranaculus repens, and Taraxacum officinale. In all years, bare ground was associated with Flood Pulse wetlands (Figure 3). Sørensen's similarity coefficient between Flood

Pulse wetlands and Static wetlands was high (QS= 0.89) in 2002 before the flood pulsing treatment was initiated. However, similarity decreased (QS= 0.72, 0.75, and 0.78 in

2003, 2004, and 2005, respectively) after the treatment was initiated. 45

Fig.3. NMDS ordinations of plant communities in Flood Pulse (black squares) and Static

(white squares) wetlands and MRPP statistical results. Dominant plant indicator species associated with Flood pulse or Static Wetlands are labeled with four letter codes that are listed in Appendix A. The code BARE signifies bare ground. 46

2003 August 2003 October

PLLA MRPP MRPP SPAR EUMA

p<0.005 OXST p<0.05 CIAR JUEF ACRU POPA TRPR POPA 80 JUEF 80 JETE ACRU PAAC EUGR OXST SPAR PAAC SYLA SYLA POPE POPS EUGR TRPR CALR BARE EUMA JETE LUPA

POPU CIAR PLLA JUCA PLMA 40 RUCR POPU POPE 40 PHAR PLMA BARE CALR TASP

Axis 2 (54.5%) 2 Axis SALA POPS Axis 2 (36.1%) 2 Axis LUPA

0 0

0 40 80 0 40 80 Axis 1 (45.5%) Axis 1 (25.3%)

2004 August 2004 October

MRPP MEAR MRPP p<0.05 p>0.05

80 POPA 80 POPA MEAR ACRU BIFR POPE GAAP MEOF POPR GAAP PLMA ULAM PLMA POSA CATR POPE BIFR SIAN PAAC SAEX EUMA PHAR SOCA HE15 MEOF LYVI SYLA AMAR TAOF ELOB BARE 40 40 DUAR SAEX TRPR CAVA COAM BYSP PRVU POCA ELOB

Axis 2 (10.3%) 2 Axis POCA Axis 2 (43.1%) 2 Axis BARE DUAR CALR AMAR PHAR

0 0

0 40 80 0 40 80 Axis 1 (32.5%) Axis 1 (61.9%)

2005 June 2005 September

MRPP MRPP RARE DACA p<0.05 MEOF p<0.05 CATR ELRE HE28 TAOF 80 TRPR RARE 80 ULAM EUMA PLLA CALR PLMA PAAC MEOF POPA EUMA ACRU POPA SCVA CYST

BYSP DIAR SOCA PODE OXST CACR CAVA BICE HE21 BICE JETE 40 CAVA 40 SALA CACR BARE JETE CALU POCA POHY CYST PAAC

PHAR BARE Axis 2 (34.2%) 2 Axis PRVU (40.0%) 2 Axis SCCY

SIAN PRVU

LYSA 0 0

0 40 80 0 40 80 Axis 1 (37.6%) Axis 1 (45.6%)

Key community attributes in the Intermittently Flooded Zones also differed between treatments. There was much more bare ground in Flood Pulse wetlands in all years (Figure 4). From 10-35% of the soil surface was bare in Flood Pulse wetlands, but plants covered most of the ground in Static Wetlands (<9% bare ground) (Figure 4).

Despite the pronounced differences in plant cover, dry weight biomass was not significantly different between treatments (Figure 4). Mean [SE] plant species richness varied between treatments and years: 2003 Flood Pulse 17.7 [0.4], Static 25.0 [0.1] (F 1,8

= 10.368, p = 0.012); 2004 Flood Pulse 27.9 [0.8], Static 32.3 [1.2] (F 1,8 = 4.054, p =

0.079) ; 2005 Flood Pulse 32.1 [0.4], Static 35.0 [1.9] (F 1,8 = 1.268, p = 0.293). There was an Date X Treatment interaction in 2004: August Flood Pulse 18.4 [1.8], Static 25.2

[4.2], F 1,8 = 0.054, p = 0.606), (F 1,8 = 0.054, p = 0.606); October Flood Pulse 17.0 [1.6],

Static 24.8 [1.7], F1,8 = 0.054, p = 0.606), (F 1,8 = 2.010, p = 0.018). Shannon’s Diversity

(H’) was higher in Static Wetlands in 2005 (Figure 4). Furthermore, mean [1 SE]

Evenness (J’) was also higher in Static wetlands; 2003 Flood Pulse 0.748 [0.33], Static

0.664 [0.23] (F 1,8 = 0.20 , p = 0.666); 2004 Flood Pulse 0.682 [0.44], Static 0.801 [0.13]

(F 1,8 = 0.345, p = 0.576); 2005 Flood Pulse 0.615 [0.06], Static 0.882 [0.07] (F 1,8 =

6.8394, p = 0.035).

There were clear impacts on the life history traits of the communities in each treatment (Table 3). By 2005, Obligate Wetland plants were much more abundant in

Flood Pulse wetlands than Static Wetlands, and Facultative plants occurred in equal 48

Fig.4. Plant community in Flood pulse and Static Wetlands in 2003-2005. A. Mean (+/-

1 SE) grams of total dry weight biomass. All t-tests were not significant (p >0.05). B.

Mean (+/-1 SE) percentage of bare ground. All t-tests were significant (p <0.05). C.

Mean (+/-1 SE) Shannon’s Diversity. Differences were not detected in 2003 or 2004 but

2005 data were significantly different (p = 0.01)

800 Flood Pulse All N.S. 700 A. Static

600

500 400 300

Biomass (g) Biomass 200 100 0 2003 2004 2005

50 B. All p< 0.05

10 Bare Ground (%) Ground Bare 0 2003 2004 2005

2.50 N.S. p = 0.01 C. N.S. 2.00

1.50

1.00

0.50 Shannon's Diversity (H') Diversity Shannon's 0.00 2003 2004 2005 50

Table 3 Life history traits for the emergent plant community in the IFZ of Flood pulse

(FP) and Static (ST) wetlands using Repeated Measures ANOVA. Data are mean percent of total plant cover.

Interaction RMA

FP ST

Year Mean [SE] Mean [SE] F1,8 p F1,8 p

Wetland Designation

Obligate 2003 54.31 [4.1] 35.91 [3.4] 0.0063 0.939 13.5468 0.006

2004 29.87 [3.8] 19.72 [2.3] 0.0485 0.831 2.6532 0.142

2005 24.52 [3.0] 12.56 [1.9] 0.5369 0.485 11.9855 0.009

Facultative Wetland 2003 25.43 [4.5] 32.26 [3.1] 0.0705 0.797 0.9286 0.363

2004 29.22 [2.8] 26.10 [3.2] 1.6772 0.231 0.3290 0.582

2005 25.49 [4.4] 31.27 [2.3] 0.5372 0.485 0.8258 0.390

Facultative 2003 11.08 [2.9] 18.32 [3.1] 0.1591 0.700 1.5848 0.244

2004 10.29 [2.3] 5.94 [0.8] 4.4324 0.068 2.0679 0.188 52

2005 12.51 [2.4] 23.23 [1.9] 1.6957 0.229 10.5128 0.012

Facultative Upland 2003 5.44 [2.7] 7.05 [1.8] 4.4149 0.069 0.1212 0.737

2004 18.20 [2.2] 33.19 [2.9] 0.6896 0.430 11.8148 0.009

2005 20.86 [4.3] 20.52 [1.6] 3.7657 0.088 0.0112 0.918

Upland 2003 0.97 [0.3] 3.41 [0.7] 0.1353 0.723 6.7204 0.032

2004 4.86 [1.1] 5.21 [1.9] 0.3303 0.581 0.0121 0.915

2005 1.40 [0.4] 2.40 [0.7] 0.1903 0.674 2.2978 0.168

Weedy 2003 40.43 [3.3] 45.08 [3.5] 0.0203 0.890 0.6400 0.447

2004 27.05 [2.8] 42.71 [2.4] 0.1311 0.727 15.7347 0.004

2005 38.68 [5.1] 39.10 [3.6] 0.2141 0.656 0.0072 0.934

Introduced 2003 3.18 [1.1] 5.59 [0.9] 7.5358 0.025 1.5029 0.255

2004 12.08 [1.2] 19.25 [2.5] 1.1022 0.324 5.9748 0.040

2005 10.38 [2.2] 13.38 [2.6] 0.0180 0.897 1.0042 0.346

Annual 2003 18.71 [2.5] 11.99 [1.1] 0.5481 0.480 3.2505 0.109 53

2004 12.19 [1.5] 30.08 [2.9] 6.2889 0.036 34.7077 0.000

2005 6.81 [1.5] 8.30 [3.0] 0.1274 0.730 0.3095 0.593

Perennial 2003 77.58 [2.5] 84.93 [1.5] 0.5007 0.499 4.0447 0.079

2004 80.69 [2.0] 60.83 [3.1] 1.9130 0.204 21.5183 0.002

2005 77.20 [8.7] 81.12 [3.5] 0.4601 0.517 0.1611 0.699

Dicot 2003 59.42 [3.2] 56.57 [3.2] 0.4567 0.518 0.0902 0.772

2004 44.88 [2.6] 64.13 [2.6] 0.3579 0.566 22.9415 0.001

2005 36.48 [6.2] 49.82 [3.7] 2.2099 0.181 5.0983 0.059

Monocot 2003 34.01 [2.6] 32.84 [1.9] 0.2508 0.630 0.2430 0.635

2004 47.14 [2.0] 32.44 [2.7] 0.0615 0.810 14.9036 0.005

2005 53.47 [7.6] 49.86 [3.8] 0.0566 0.819 0.0625 0.810

proportions in both treatments. In contrast, Facultative Upland and Upland plants were a greater proportion of the community in Static Wetlands. Weedy and species were more abundant in Static Wetlands than in Flood Pulse wetlands. Annuals were more abundant in Static Wetlands, and the proportion of Perennial plants was higher in Flood

Pulse wetlands. Dicots were more abundant in Static Wetlands and Monocots were more abundant in Flood Pulse wetlands.

Discussion

In previous studies, flood pulsing strongly impacted the biota in large order riparian systems (Benke 2001, Bunn and Arthington 2002, Malmqvist 2002, Ward et al.

2002). In this study, I detected that flooding also causes pronounced differences in emergent plant communities in the Intermittently Flooded Zone (IFZ) along a headwater creek, as predicted in the first hypothesis. However, the second hypothesis was not supported because plant and invertebrate communities and environmental conditions were similar in permanently flooded wetlands with or without flood pulsing. This suggests that ecological impacts of the short, unpredictable flood events in headwater riparian wetlands may be less important in stable, permanent pools than above the water line in the IFZ.

Impacts of Flooding in the IFZ

Multivariate ordination of the emergent plant communities detected clear distinctions between the Flood pulse and Static wetlands in all years. This illustrates that the treatment created a strong selective pressure on the regional species pool within the first year of the study. The data suggests that tolerance to submergence and soil anoxia are important adaptive traits in headwater wetlands. For example, I found a higher proportion of flood-intolerant taxa (Facultative Upland and Upland species) in

Static wetlands but flood-tolerant taxa (Obligate wetland) in Flood Pulse wetlands.

Therefore, it is clear that effects of frequent short-term flooding in headwater systems can be sufficient to suppress intolerant plant species.

The changes I found in community structure would probably impact key ecosystem functions in riparian habitats. The proportion of the community dominated by perennial and monocot species was often higher in Flood Pulse wetlands. Perennials generally have more below-ground biomass that persists through the winter (Roumet et al. 2006). This may help stabilize soils during heavy spring floods. Monocots can also reduce erosion because grasses have deep fibrous root systems. Moreover, potential wetland water budgets may have been affected because Obligate plants are characteristically more efficient at water uptake and storage (Tiner 2006).

Furthermore, annual plants were more abundant in Static wetlands and these typically have high seed production that provides food for granivorous mammals and waterfowl

(Silvertown 1981).

A flood pulsing hydrology in large order river systems has been associated with higher plant growth (Ward and Stanford 1995), but I did not detect any differences in total biomass between Flood pulse and Static wetlands. This was surprising because I measured 20-45% more bare ground in the Flood Pulse wetlands. The data suggests that individual plants in Flood Pulse wetlands may have been more robust, and thereby the total amount of biomass was similar in both treatments.

The data indicate that impacts of water levels fluctuation in headwater systems can be stressful to plant communities because there was more bare ground and lower emergent plant species diversity in Flood Pulse wetlands. This stressful hydrology may have increased seedling mortality and/or reduced in plant growth rates (Van Der Sman et al. 1993, Casanova and Brock 2000, Robertson et al. 2001, Johansson and Nilsson

2002), which has been documented in natural floodplain ecosystems (Blom 1999).

Potentially important stressors include intermittent submergence and sedimentation.

The Flood Pulse wetlands water levels were frequently up to 75 cm above the permanent pool, and series of storms caused extended floods for periods of 2-8 days.

As a result, the percent of time wetlands were flooded deeper than 20 cm was <2% in

Static wetlands but 18-25% in Flood Pulse wetlands. Anoxia in flooded soils will impair root cell metabolism and build up toxins (e.g., ammonia, hydrogen sulfide) (Justin and

Armstrong 1987, Vartapetian and Jackson 1997). Light attenuation during submergence can also increase mortality by reducing photosynthesis (Lenssen et al.

1998). Furthermore, I observed heavy silt deposits on germinating seedlings after spring

floods subsided, and silt deposition reduces seedling survival and decreases plant growth rates (Jurik et al. 1994, Ewing 1996, Gleason et al. 2003). Although erosion can be another important stressor in many riparian systems (Ward et al. 2002, Steiger et al.

2003), it was not important in the experiments because flow velocities were never high enough to cause erosion (M. Drinkard, pers. observ.).

Impacts of Flooding in the Permanent Pool

The results show that although flooding in headwater wetlands can be stressful to plants in the IFZ, it is less important in permanent pools in floodplains. Although aquatic invertebrate diversity was lower in permanent pools of Flood Pulse wetlands in

2003, invertebrate diversities were the same in 2004 suggesting that any stressful impacts were transient. Furthermore, submersed plant communities and algal biomass were never different. Although turbidity, pH and dissolved oxygen were different at times between treatments, turbidity was always low, pH was circumneutral, and dissolved oxygen levels were high in both treatments. Therefore, treatments did not have major impacts on key habitat conditions.

It was not very surprising that I did not detect substantial changes in submersed plants or algae because the pools were never drawn down, but I expected that there would be a beneficial impact on aquatic invertebrates via access to the floodplains.

Although aquatic invertebrates were slightly more diverse in Static wetlands and there

were some differences in trophic groups in 2003, total numbers were never different.

Furthermore, NMDS ordinations did not detect any significant differences in invertebrate assemblages in any year. This was unexpected because others have found that differences in the frequency of flooding will create unique invertebrate communities (Reese and Batzer 2007).

There are several potential reasons that there were few differences in aquatic invertebrates. First, there are probably fewer stressful impacts of fluctuating water levels in permanent pools than in intermittently inundated habitat. Furthermore, biotic factors (e.g. fish predation) can be very important in stable habitats (Chase 2007), and these may have been similar in all of the permanently flooded pools. The design of the mesocosms may have also impacted the results. For example, overland flow can import high amounts of materials into floodplain pools (Jenkins and Boulton 2003), but the mesocosms received their inputs through the water control structure. Moreover, the amount and quality of resources subsidies between habitats influences the impact of flood pulsing (Junk et al. 1989). While the permanent pools were comparable in size to those occurring in natural floodplains (M. Drinkard, pers. observ.), the IFZ in the mesocosms were only about 2 m wide. The pool to floodplain ratio in the mesocosms was only 1:1.3, and therefore resource subsidies into the permanent pool may have been minimal. Therefore, I acknowledge that flood pulsing may be more important in natural floodplain pools than I found in this study. Further research would be useful to

test the ecological significance of resources subsidies under different pool to floodplain ratios or hydroperiods in headwater systems.

Management Implications

It has become widely understood that the duration and frequency of flooding is an important variable in large order riparian systems (Ewing 1996, Casanova and Brock

2000). Flooding regimes are often manipulated to enhance valuable functions such nutrient retention or wildlife habitat (de Szalay et al. 2003), and a better understanding of flood pulsing is needed to guide management efforts (Barry et al. 2004, Freeman et al. 2007, Lake et al. 2007). Low-order streams have unique habitat conditions, which will alter the biotic responses of flood pulsing. Therefore, developing comprehensive watershed management strategies will require testing flooding in headwater systems.

However, a lack of control over flooding and a lack of replicates can make it challenging to test the influence of hydrology in natural systems.

The frequent, stochastic, short-term flooding I created in the mesocosms closely simulated the hydrology expected in temperate low-order streams (Middleton 2002).

Therefore, the results show that native and non-weedy taxa can become more abundant in Flood pulsing wetlands. This may be because native species are better adapted to natural flood regimes (Tabacchi 1995). Colonization by exotics can impair wetland functions (Tabacchi et al. 2000), and several of the exotics abundant in Static

wetlands (e.g., Cirsium arvense, Melilotus officinalis, Daucus carota) are considered noxious pests. In contrast, the indicator species in Flood Pulse wetlands included many that are important wildlife food plants (e.g., Bidens cernua, Carex lupulina, C. vulpinoidea, Polygonum hydropiperoides, P. punctatum, Scirpus cyperinus). Therefore, the results suggest that creating a natural flood regime in managed or restored wetlands can be an important tool to promote desired plant communities in low-order systems.

Acknowledgements

I thank Brendan Morgan, Jennifer Clark, Constance Hausman, Justin

Montemarano, Douglas Kapusinski, Leonard N. Drinkard, III, Lauren Drinkard, Karen

Montgomery, Rachel Johnson, Megan Meuti, Mike Woods, Natasha Wingerter, and

Emily Faulkner for field assistance. This research was supported by the Art and

Margaret Herrick Endowment for the Aquatic Ecology Research Facility.

Appendix A. Mean (1 SE) percent cover of dominant plant taxa in Flood pulse and Static wetlands in 2003-2005. Codes are the four letter code used in the NMDS ordination graph. Taxa that were not indicator species in the NMDS analysis are not shown on the

NMDS plots.

Specimen Code FP Mean [SE] ST Mean [SE]

Acer rubrum ACRU 2003 0.10 [0.1] 0.23 [0.1]

2004 0.10 [0.1] 0.23 [0.1]

2005 0.10 [0.1] 0.37 [0.1]

Ambrosia artemisiifolia AMAR 2004 0.40 [0.2] 0.10 [0.1]

Bidens cernua BICE 2003 7.73 [1.3] 11.87 [1.3]

2005 1.23 [0.3] 0.33 [0.1]

Bidens frondosus BIFR 2004 0.23 [0.2] 3.43 [1.2]

Bryophyta spp. BYSP 2005 2.03 [0.6] 1.40 [0.4]

Carex crinita CACR 2005 4.47 [2.1] 1.10 [0.4]

Carex lupulina CALU 2003 15.33 [4.2] 5.47 [1.2]

2004 1.57 [0.4] 0.70 [0.2]

2005 8.07 [2.7] 1.67 [0.8]

Carex lurida CALR 2005 2.23 [1.2] 3.93 [1.8]

Carex tribuloides CATR 2004 0.87 [0.3] 1.07 [0.3]

2005 0.03 [0.0] 0.08 [0.0]

Carex vulpinoidea CAVA 2004 6.87 [1.5] 4.33 [0.9]

2005 8.30 [1.7] 2.40 [0.6]

Cirsium arvense CIAR 2003 0.93 [0.5] 2.80 [0.3]

2005 1.97 [0.7] 1.80 [0.4]

Cladophora sp. CLSP 2004 2.40 [0.9] 2.10 [0.5]

Cornus amomum COAM 2004 1.37 [0.6] 1.33 [0.2]

Cyperus strigosus CYST 2003 7.83 [1.7] 12.33 [2.1]

2005 0.70 [0.3] 0.90 [0.4]

Daucus carota DACA 2004 0.60 [0.4] 0.53 [0.3]

2005 1.20 [0.6] 5.97 [2.8]

Dianthus armeria DIAR 2005 1.77 [1.4] 0.73 [0.4]

Dicots unknown DIUN 2005 1.90 [0.7] 2.77 [1.0]

Dulichium arundinaceum DUAR 2004 0.60 [0.2] 0.27 [0.2]

Echinochloa crus-galli ECCR 2004 1.700 [1.1] 1.03 [0.2]

Eleocharis obtusa ELOB 2004 1.50 [0.2] 0.97 [0.3]

Elymus repens ELRE 2005 0.27 [0.2] 0.13 [0.1]

Elymus riparius ELRI 2005 0.10 [0.1] 0.17 [0.1]

Eupatoriadelphus maculatus EUMA 2003 1.20 [0.8] 5.57 [1.5]

2004 0.03 [0.0] 0.43 [0.2]

2005 0.70 [0.2] 8.57 [1.2]

Euthamia graminifolia EUGR 2003 3.90 [1.7] 5.70 [0.9]

2005 0.93 [0.4] 1.17 [0.6]

Galium aparine GAAP 2004 0.03 [0.0] 5.67 [1.0]

Herbaceous 15 HE15 2004 1.17 [0.2] 1.03 [0.2]

Herbaceous 21 HE21 2005 0.57 [0.4] 0.17 [0.1]

Herbaceous 27 HE27 2005 0.57 [0.3] 0.27 [0.1]

Herbaceous 28 HE28 2005 0.03 [0.0] 0.27 [0.2]

Herbaceous 5 HE05 2003 0.17 [0.1] 0.63 [0.2]

Juncus candensis JUCA 2003 6.20 [2.2] 2.57 [1.8]

Juncus effusus JUEF 2003 2.37 [1.0] 11.60 [2.5]

2005 6.60 [2.2] 8.43 [0.9]

Juncus tenuis JETE 2003 3.93 [1.7] 13.53 [4.0]

2005 7.27 [1.3] 3.40 [1.2]

Ludwigia palustris LUPA 2003 2.63 [0.9] 0.00 [0.0]

Lycopus virginicus LYVI 2004 5.07 [1.5] 2.33 [0.6]

Lythrum salicaria LYSA 2005 0.37 [0.2] 0.53 [0.5]

Melilotus officinalis MEOF 2004 2.40 [0.8] 4.17 [0.6]

2005 0.20 [0.1] 1.67 [0.8]

Mentha arvensis MEAR 2004 0.30 [0.2] 2.57 [1.3]

Monocot 4 MO04 2005 0.87 [0.5] 1.10 [0.8]

Monocot 7 MO07 2005 0.30 [0.1] 1.20 [0.8]

Monocots Unknown MOUN 2005 1.97 [0.7] 2.83 [1.0]

Oxalis stricta OXST 2003 0.10 [0.1] 0.50 [0.2]

2005 1.30 [0.4] 1.73 [0.7]

Panicum acuminatum PAAC 2003 2.87 [1.1] 4.57 [1.3]

2004 0.87 [0.3] 1.60 [0.6]

2005 0.07 [0.0] 3.93 [1.5]

Phalaris arundinacea PHAR 2003 8.57 [3.6] 9.97 [4.8]

2004 0.50 [0.4] 0.17 [0.1]

2005 5.47 [2.3] 4.53 [1.8]

Plantago lanceolata PLLA 2003 0.23 [0.1] 1.33 [0.6]

2005 0.73 [0.3] 1.37 [0.7]

Plantago major PLMA 2003 1.30 [1.0] 1.03 [0.4]

2004 0.93 [0.6] 0.93 [0.9]

2005 0.17 [0.1] 0.33 [0.1]

Poa palustris POPA 2003 0.00 [0.0] 2.63 [0.6]

2004 0.30 [0.2] 0.53 [0.2]

2005 4.43 [1.9] 10.63 [3.5]

Poa pratensis POPR 2004 0.03 [0.0] 0.47 [0.1]

2005 0.43 [0.3] 0.20 [0.1]

Polygonum hydropiperoides POHY 2005 1.37 [0.6] 0.00 [0.0]

Polygonum pensylvanicum POPE 2003 0.30 [0.2] 0.37 [0.2]

2004 0.43 [0.2] 1.40 [0.3]

Polygonum punctatum POPU 2003 7.60 [1.3] 3.33 [1.4]

2004 0.50 [0.2] 1.00 [0.6]

Polygonum sagittatum POSA 2004 0.90 [0.3] 2.30 [1.1]

Populus deltoides PODE 2005 0.90 [0.4] 0.40 [0.1]

Potamogeton pusillus POPS 2003 6.33 [2.3] 13.93 [4.2]

Potentilla canadensis POCA 2004 0.90 [0.3] 0.00 [0.0]

2005 1.53 [0.7] 1.90 [0.7]

Prunella vulgaris PRVU 2004 0.50 [0.2] 0.70 [0.4]

2005 2.03 [0.9] 0.53 [0.3]

Ranunculus repens RARE 2004 0.40 [0.2] 1.03 [0.2]

2005 0.07 [0.1] 0.33 [0.1]

Rumex crispus RUCR 2003 0.97 [0.3] 0.53 [0.2]

2005 0.43 [0.3] 0.47 [0.2]

Sagittaria latifolia SALA 2003 2.93 [1.2] 5.07 [2.5]

2005 2.43 [0.8] 1.57 [0.6]

Salix exigua SAEX 2004 0.37 [0.2] 0.17 [0.1]

Scirpus atrovirens 2004 1.63 [0.7] 1.27 [0.5]

Scirpus cyperinus SCCY 2003 1.83 [1.3] 1.07 [0.6]

2005 11.07 [2.9] 8.83 [3.0]

Scirpus validus SCVA 2005 0.10 [0.1] 0.47 [0.3]

Sisyrinchium angustifolia SIAN 2004 12.07 [3.6] 6.07 [2.6]

2005 1.70 [0.6] 0.57 [0.2]

Solidago canadensis SOCA 2004 4.27 [1.5] 4.00 [1.3]

2005 14.17 [3.5] 15.13 [2.0]

Solidago rugosa SORU 2004 0.13 [0.1] 0.30 [0.2]

Sparganium americanum SPAR 2003 0.20 [0.1] 4.30 [1.8]

Symphyotrichum lanceolatum SYLA 2003 0.50 [0.3] 2.80 [0.9]

2004 3.10 [0.9] 2.17 [0.6]

Symphyotrichum pilosum SYPI 2005 3.40 [1.1] 5.43 [2.7]

Taraxacum officinale TAOF 2004 0.37 [0.1] 0.53 [0.1]

2005 0.23 [0.1] 0.57 [0.3]

Taraxacum sp. TASP 2003 0.87 [0.4] 0.87 [0.3]

Trifolium pratense TRPR 2003 0.27 [0.2] 2.07 [0.6]

2004 2.20 [0.5] 2.93 [1.1]

2005 1.97 [0.7] 3.07 [1.4]

Trifolium repens TRPR 2005 2.37 [1.1] 1.47 [0.7]

Ulmus americana ULAM 2004 0.47 [0.2] 0.73 [0.1]

2005 0.80 [0.3] 0.97 [0.2]

CHAPTER 3

ZONATION OF PLANT COMMUNITIES CAUSED BY HYDROLOGICAL STRESSES IN

HEADWATER RIPARIAN WETLANDS

Abstract

In most riparian wetlands, plant community structure is controlled by abiotic stressors influenced by the timing of flooding events. Important stressors, including desiccation, soil anoxia, submergence, erosion, and sedimentation, act as an environmental sieve on the potential pool of colonists producing a realized community that is highly resistant to stresses. The intensity of these deterministic abiotic

conditions should vary along an elevation gradient due to changes in the length and frequency of inundation. This study examined the effects of a flood pulsing hydrology on plant communities in large-scale (10 m x 20 m) wetlands. I sampled plant communities for two years in Flood Pulse wetlands that fluctuating after storm events and Static wetlands that had stable water levels. I also compared plant community structure (percent cover of dominant species, percent cover of taxa grouped by their wetland indicator status, percent cover of bare ground, total biomass, species richness) in Low, Mid and High elevation zones (from 0, 66, and 132 cm above the baseline water levels, respectively) in each treatment. Hydrological stresses caused by flood pulsing strongly affected plant community structure. Percent cover of bare ground was higher and richness was lower in Flood Pulse wetlands than Static wetlands, which suggested that plants that could not tolerate stochastic flooding were eliminated. However, plant biomass was highest in Low elevation zone, which was frequently inundated and received the most nutrient inputs. The variable hydrology in Flood Pulse wetlands selected for Obligate wetland plants but Static wetlands were dominated by Facultative

Wetland, Facultative and Upland plants. Furthermore, each elevation zone had distinct species assemblages in Flood Pulse wetlands but not in Static wetlands. The most flood-tolerant species were consistently found in the Flood Pulse / Low and Flood Pulse

/ Mid elevation zones, and the least flood-tolerant species occurred in the Static / High elevation zones. Different species assemblages were found in High elevation zones in

Flood Pulse wetlands that were only flooded 3% of the time, and Static wetlands that

were flooded <1% of the time. This illustrates that even subtle differences in flooding frequency on this headwater creek strongly affected plant community structure. The higher abundance of some non-native wetland plant species in Static wetlands demonstrates the importance of maintaining natural flooding regimes in headwater ecosystems.

Introduction

It has been well established that biotic factors such as interspecific competition and herbivory affect plant community distribution in many terrestrial ecosystems

(Connell 1983; Sih et al. 1985). Furthermore, some aquatic habitats with predictable flood regimes or stable water levels are also strongly structured by biotic factors

(Lenssen et al. 1999; Barry et al. 2008; Sanderson et al. 2008). However, abiotic stresses are often the primary factors structuring plant communities in ecosystems with variable water regimes like floodplain wetlands (Weiher et al. 1998). In these systems, alternating periods of flooding and drawdown act as an environmental sieve that eliminates species that lack adaptations to stochastic changes in environmental conditions. Stresses during flooding include anoxic soils, accumulation of toxic reduced chemicals (e.g. ammonia, hydrogen sulfide), and low light levels below the water. In contrast, stresses during drawdowns are desiccation and erosion of sediments

(Middleton 1999; Casanova & Brock 2000; Hefting et al. 2004). Reduced plant diversity in these systems can further reduce the importance of inter-specific competition (Fraser

& Karnezis 2005; Lenssen & de Kroon 2005; Fraser & Miletti 2007).

Plants in intermittently flooded zones (IFZ) of riparian wetlands need adaptations for stresses caused by flooding and draw downs (Weiher & Keddy 1995; Bunn &

Arthington 2002). Common traits in wetland plants are shallow root systems, aerenchyma tissue, ability to use anaerobic and aerobic cell metabolic pathways, and life cycle synchronicity (e.g., seed emergence) stimulated by drawdown (Blom &

Voesenek 1996). Physiological and morphological responses can be induced by changing water levels, and some species respond to differences in water level as little as

1 cm (Fraser & Miletti 2007).

The unique environmental conditions in riparian floodplains strongly affect key biotic properties. The Flood pulse Concept (FPC) posits that floodplain community structure, biodiversity and productivity are largely determined by the flood regime (Junk et al. 1989; Poff et al. 1997; Tockner et al. 2000). For example, flood-pulsing can promote high biodiversity and productivity because it allows nutrient subsidies to enter from the river channel (Tockner et al. 1999; Junk & Wantzen 2004). However, frequency and duration of inundation decrease at higher elevations in wetlands, and gradients of stresses and benefits can form zones with different biotic communities.

Plant zonation along elevation gradients has been described many wetland systems including salt marshes (Silverstri et al. 2005), prairie potholes (Galatowitsch &

van der Valk 1995), tropical forested floodplains (Junk et al. 1989), Great Lake coastal wetlands (Barry et al. 2004), and large, temperate river systems (Tockner et al. 1999;

Boudell & Stromberg 2008). Hydrological stresses usually are most important at lower elevations, but other factors such as biotic interactions become important at higher elevation (Lenssen et al. 2004; Fraser & Miletti 2007; Sanderson et al. 2008). For example, flood-intolerant plants in wetlands with fluctuating water levels are uncommon at lower elevations (Cassanova & Brock 2000; Battaglia & Collins 2006), and species richness is highest at higher elevations (Zelnik & Carni 2008). However, less is known about impacts of hydrology in ecosystems with short unpredictable flooding, such as floodplains along headwater creeks.

In this study, I examined how flood regimes affected the structure of riparian plant communities along a headwater creek. I also examined if species tolerance to flooding varied among different plant communities in zones along an elevation gradient.

I established wetlands with a flood-pulsing hydrology (Flood Pulse wetlands) and stable water levels (Static wetlands) and tested the following hypotheses:

H1: Due to the different hydrologies, Flood Pulse wetlands and Static wetlands will have distinct plant communities. Flood Pulse wetlands will have more diverse plant communities with higher biomass than Static wetlands.

H2: Fluctuating water levels will create distinct plant communities along an elevation gradient. Plant communities with lower diversity and higher biomass will

occur at lower elevations that are more frequently inundated by flooding. Plants that tolerate soil flooding will grow at higher elevations in Flood Pulse wetlands than in Static wetlands.

Methods

Study Site

This study was conducted at the Art and Margaret Herrick Aquatic Ecology

Research Facility (HAERF) on the Kent State University campus (Portage Co., Ohio). In

2001, 10 earthen wetland mesocosms (10 m x 20 m) were excavated adjacent to a dammed stream pool (100 m X 10 m), which was the water source for the mesocosms.

The banks of mesocosms were sloped inwards, and the water depth in the center was

1.7 m (total volume = 2.5 X 105 l).

In spring of 2002 and 2003, each mesocosm was seeded with a mix of native wetland plant species. The species ranged from obligate wetland plants that tolerate extended flooding or permanently saturated soils (Bidens cernua, Carex crinita, C. vulpinoidea, C. lupulina, Eleocharis palustris, Juncus effusus, Leersia oryzoides,

Pontederia cordata, Sagittaria latifolia, Schoenoplectus acutus, Sparganium americanum, Solidago patula) to facultative wetland plants that generally are found in intermittently flooded habitats (Cyperus esculentes, Elymus riparius, Eupatorium

perfoliatum, Onoclea sensibilis, Polygonum lapathifolium, P. pennsylvanicum, Scirpus cyperinus, Poa palustris). The plant communities were not manipulated after that date, and additional species colonized from nearby habitats. See Drinkard et al. (In Press) for more details on the study site.

Design of hydrology treatments and plant sampling

In 2002, I randomly assigned mesocosms into two hydrology treatments: Flood

Pulse wetlands (N=5) and Static wetlands (N=5). Each mesocosm had separate inlet and outlet water-control structures that used stacked PVC boards to adjust the inflow and outflow of water. I set PVC boards in the inlet structures at 5 cm above the stream pool so that all mesocosms only received water inputs during storm flooding. In Static wetlands, outlet water control structures were set so that they did not retain floodwaters above a maximum depth of 80 cm. In Flood Pulse wetlands, additional PVC boards were added that retained floodwaters up to 140 cm, but a 1 cm hole was drilled through a board at 80 cm depth. Therefore, water levels in these wetlands could rise sharply after a storm and slowly drop back down to a depth of 80 cm. As a result, the

Static Wetlands had two distinct habitats: a permanent pool and non-flooded banks.

The Flood Pulse wetlands had three habitats: a permanent pool, an intermittently flooded zone (IFZ), and non-flooded banks. These treatments were maintained from

2002 to 2006, which allowed plant communities to become established.

Emergent plant communities above the permanent pool were sampled in

September 2005 and September 2006. On each date, I visually quantified percent cover in a 1-m2 (0.5 m x 2 m; W x L) quadrat at three randomly chosen locations on the north bank of each wetland. The quadrat was laid on the mesocosm bank perpendicular to the 80-cm water line of the permanent pool. The quadrat was dividing into three elevation zones: 1) Low Zone - from the baseline water line to 66 cm; 2) Mid Zone - from

66 to 132 cm; and 3) High Zone- from 132 cm to 200 cm. In Flood Pulse wetlands, the quadrats spanned the IFZ and extended into the non-flooded banks.

All plants were identified using taxonomic keys to species (Crow & Hellquist

2000; Voss 2001; Chadde 2002), and current nomenclature was from PLANTS website of the USDA (USDA, NRCS 2011). Voucher specimens are stored in the Kent State

University Herbarium. I also measured above-ground biomass in September 2005.

Plants in one randomly chosen quadrat in each wetland were clipped at ground-level and separated by elevation zone. Dry weight of each species was measured after drying for 72 h at 60°C.

I used the Wetland Indicator Status (WIS) of each taxon to estimate their tolerance to soil saturation (Reed 1997). The WIS classifies taxa by their occurrence in wetlands; Obligate (OBL) plants are found in wetlands over 99% of the time, Facultative

Wetland (FACW) plants are found in wetlands between 67% and 99% of the time,

Facultative (FAC) plants are found in wetlands between 33% and 67% of the time,

Facultative Upland (FACU) plants are found in wetlands between 1% and 33% of the

time, and Upland (UPL) plants are found in wetlands less than 1% of the time (Reed

1997). Generally, the more common a species is in wetlands (i.e., OBL taxa), the more tolerant it is to stresses in saturated soils (Tiner 1991).

The water depth in each mesocosm was measured 3-5 times per week using a fixed staff gauge. I estimated the slope angle of the bank in each mesocosm using a carpenter’s level and meter stick. This angle was used to determine the water depth that would flood each elevation zone in each mesocosm. I then calculated the percent of time that each elevation zone in each mesocosm was flooded using the water level data and the estimated bank slope angles.

Statistical Analysis

All data were tested for normality with the Kolmogorov-Smirnov Test. Non- normal percent data were arcsine transformed and count data were log (X+1) transformed (Zar 1999). I used the transformed or raw data that more closely approximated a normal distribution.

In each year, I used 1-way ANOVAs to compare the mean percent of time each elevation zone in each treatment was flooded (below the permanent pool water baseline level, and the Low, Mid and High elevation zones). I defined the common plant taxa as those species that were found in at least 3 of 5 wetlands of one treatment or comprised at least 5% of total percent cover in a treatment or elevation zone. In

each year, percent cover, biomass, and species richness were compared between treatments and elevation zones using 2-way ANOVAs. Also, percent cover data of all species were pooled by WIS, and percent cover of each category was compared between treatments and elevation zones with 2-Way ANOVAs. All significant ANOVAs were followed by Tukey’s HSD tests to make pairwise comparisons among means. All univariate statistics were all run on SPSS software (version 15.0, 2007).

I also used multivariate methods to examine if distinct community types were found in each treatment / elevation zone habitat type. Non-metric multidimensional scaling (NMDS) with Sørensen distances were run on PC-ORD software (Version 5). The ordination procedure used a random start and 250 runs. The number of dimensions retained in the final ordination was determined by including all axis that reduced stress

(i.e. increased the model’s goodness of fit) by at least 5 (on a scale of 0 to100) and yielded a significant model (p<0.05). Ordinations were performed on each sampling date with the percent cover data of common taxa. Treatment / elevation zone habitat type was included as a covariate. I tested whether community groups were distinct by performing multi-response permutation procedure (MRPP) with Euclidean distance. A

MRPP was run on each year’s data to compare treatment effects on assemblage composition. I also identified indicator species of each treatment / elevation zone habitat type with an Indicator Species Analysis (Dufrene & Legendre 1997). Indicator values show which taxa are common and abundant in plant community. I ran all multivariate statistics with PC-ORD (Version 5).

Results

Abiotic conditions

In both years, there were distinctly different hydroregimes in the Flood pulse and Static Wetlands (Figure 1). After storms, water levels in Flood Pulse wetlands were typically 20-60 cm above baseline water levels, and it took 4-10 days for levels to decrease to the baseline levels. Static Wetlands remained near baseline water levels except for a few hours after intense storms (<5 cm rainfall / hr). In 2005, levels decreased below baseline water levels in all wetlands due to evapotranspiration during an extended summer dry period (Figure 1). The estimated average bank slope angle of all mesocosms was 26º, which meant a 66 cm-wide elevation zone would be fully submerged when water depths increased by ~29 cm.

As a result of these patterns, the corresponding elevation zones in Flood pulse and Static wetlands had different hydrologies in 2005 and 2006 (Table 1). Static

Wetland water levels remained near the baseline water level or within the Low Zone

95% - 99% of time, and the Mid Zone and High Zones were rarely (<3% of the time) flooded. Flood Pulse wetlands experienced a greater range of flooding, and water levels were at or below the Low Zone for 67% - 81% of the time. The Mid Zone was often flooded (16% - 23% of the time) in Flood Pulse wetlands, and this zone was

Fig.5. Mean relative water levels in Flood Pulse and Static wetlands. Hydrograph shows change from the baseline water line from January 2005 – December 2006. The depth of the baseline water line in each mesocosm is “0” on the Y-axis. Across all mesocosms, the Low Zone extended from the baseline water level to an average depth of 29 cm, the

Mid Zone extended from 29 to 58 cm, and the High Zone extended from 58 cm to 87 cm above the baseline water line.

70 Flood Pulse

60 Static

10 Relative Water Level (cm) Level Water Relative

-10

-20

05 06

06 05 05 05 05 05 05 05 06 06 06 06

05 05 05 06 06 06

- -

------

Jul Jul

Apr Oct Apr Oct

Jan Jun Jan Jun

Feb Feb

Mar Mar

Nov

Aug Sep Aug Sep

Dec

May May

Table 4. Student’s t-test results showing mean [SE] percent of time the elevation zones in each treatment were flooded. FP = Flood Pulse wetlands, ST= Static wetlands.

Statistic results are comparing values in each elevation zone across treatments in each year.

Zone FP ST t 1,8 P

Below Baseline 21.88 [4.9] 30.94 [8.0] -0.966 0.363

Low Zone 59.06 [2.4] 67.65 [7.5] -1.091 0.307

2005 Mid Zone 15.88 [3.7] 0.94 [0.6] 4.033 0.004

High Zone 2.82 [1.1] 0.35 [0.2] 2.127 0.066

Below Baseline 25.85 [5.3] 26.18 [10.3] -0.028 0.978

Low Zone 41.46 [10.7] 68.78 [11.5] -1.741 0.12

2006 Mid Zone 23.25 [5.2] 1.95 [0.8] 4.027 0.004

High Zone 9.27 [5.3] 2.93 [1.6] 1.145 0.285

significantly more flooded than in Static Wetlands in both years. The High Zone was occasionally flooded (3-9%) of the time, but this was not different than in the Static

Wetlands.

Plant community responses

The plant community included 98 taxa, of which 42 were considered common

(Table 2). Thirty five common taxa had different abundances among elevation zones and/or treatments. Bidens cernua, Carex lurida, C. vulpinoidea, Cyperus strigosus,

Populus deltoides, Polygonum hydropiperoides, and unknown dicot seedlings had higher numbers in Flood Pulse wetlands, and Acer rubrum, Agropyron repens, Daucus carota,

Panicum acuminatum, Poa palustris, Symphyotrichum lanceolatum and S. pilosum, and unknown monocots seedlings were higher in Static wetlands. Plants that were more abundant in the Low zone were Agrostis alba, Carex lurida, Euthamia graminifolia

Juncus effusus, Ludwigia palustris, Sagittaria latifolia, and Scirpus cyperinus. Plants that were more abundant in the Mid zone were Acer rubrum, Solidago canadensis, Populus deltoides and unknown dicot seedlings. Plants that were more abundant in the High zone were Andropogon gerardii, Eupatoriadelphus maculatum, Panicum acuminatum,

Poa palustris, Polygonum pensylvanicum, Potentilla canadensis, Plantago lanceolata,

Solidago canadensis, and Symphyotrichum pilosum.

Table 5. Mean [SE] percent cover of the common taxa in 2005 and 2006 in each treatment / elevation zone habitat type. Statistics are results of 2-way ANOVAs for treatment (Flood pulse vs. Static), elevation zone (Low vs. Mid vs. High Zone), and treatment x elevation zone interaction effects. Elevation zones were compared with

Tukey’s HSD tests when there was a significant zone effect.

Low Zone Mid Zone High Zone Treatment Zone Interaction Tukeys HSD

Low Low Mid

F1,24 p F2,24 p F2,24 p vs vs vs

Common Taxa FP ST FP ST FP ST Mid High High

Acer rubrum 2005 0 [0] 0 [0] 0 [0] 0.07 0 [0] 0 [0] 1 0.327 1 0.383 1 0.383

[0.1]

2006 0 [0] 0 [0] 0 [0] 0.47 0 [0] 0 [0] 7.54 0.011 7.54 0.003 7.54 0.003 0.007 1.000 0.007

[0.2]

Agrostis alba 2005 0 [0] 0.08 0 [0] 0 [0] 0 [0] 0 [0] 6 0.022 6 0.008 6 0.001 0.016 0.016 1.000

[0.1]

2006 0 [0] 0 [0] 0 [0] 0 [0] 0 [0] 0 [0] 0.53 0.473 4.79 0.018 0.53 0.594

Agropyron 2005 0 [0] 0.2 0 [0] 0 [0] 0 [0] 0.13 6 0.022 2.67 0.09 2.67 0.116 repens [0.1] [0.2]

2006 0 [0] 0 [0] 0 [0] 0 [0] 0 [0] 0 [0]

Andropogon 2005 0 [0] 0 [0] 0 [0] 0 [0] 0 [0] 0.13 2.67 0.116 2.67 0.09 2.67 0.09 gerardii [0.1]

2006 0 [0] 0 [0] 0 [0] 0 [0] 0.47 1.67 [1] 1.1 0.304 3.48 0.047 1.1 0.348 1.000 0.077 0.077

[0.5]

Bidens cernua 2005 0.33 0.27 0.93 0 [0] 0.33 0 [0] 8.99 0.006 1.37 0.273 2.99 0.069

[0.2] [0.2] [0.3] [0.1]

2006 0.2 0 [0] 0.33 0 [0] 0.07 0 [0] 5.4 0.029 0.8 0.461 0.8 0.461

[0.1] [0.2] [0.1]

Bryophyta spp. 2005 0.8 0.4 1.4 [0.5] 0.47 0.47 0.33 3.98 0.057 1.61 0.22 0.92 0.411

[0.4] [0.3] [0.2] [0.2] [0.2]

2006 3.33 4.73 5.6 [2.6] 4.6 [2] 1.73 3 [0.5] 0.16 0.695 1.29 0.293 0.31 0.737

[1.9] [1.6] [0.6]

Carex crinita 2005 0 [0] 0.67 0 [0] 0 [0] 0 [0] 0 [0] 3.33 0.08 3.33 0.053 3.33 0.053

[0.4]

2006 0.13 1.93 0 [0] 0 [0] 0 [0] 0 [0] 2.18 0.153 2.87 0.076 2.18 0.135

[0.1] [1.2]

Carex lurida 2005 9.73 2.33 4.53 0.53 1.4 [0.9] 0 [0] 21.8 <0.00 10.7 <0.00 0.81 0.457 0.171 0.171 1.000

[2.8] [1.4] [1.6] [0.2] 3 1 1

2006 13 [2.9] 7.93 5.93 9.13 0.67 0 [0] 0.15 0.702 7.65 0.003 1.2 0.318 0.523 0.002 0.032

[1.5] [2.7] [4.9] [0.5]

Carex lupulina 2005 0 [0] 4.07 0 [0] 0 [0] 0 [0] 0 [0] 2.32 0.141 2.32 0.12 2.32 0.12

[2.7]

2006 0 [0] 0 [0] 0 [0] 0 [0] 0 [0] 0 [0]

Carex 2005 2.07 1.8 3.27 [1] 1.27 2.33 0.13 4.59 0.043 0.77 0.475 0.78 0.47 vulpinoidea [1.5] [0.5] [0.5] [0.8] [0.1]

2006 2.73 2.8 5.93 5.47 3.6 [1.3] 3.6 [1.8] 0.01 0.933 1.23 0.311 0.01 0.989

[2.5] [0.8] [2.7] [1.9]

Cirsium 2005 0 [0] 0.47 0.8 [0.4] 1 [0.4] 1.53 1 [0.1] 0.02 0.895 3.3 0.054 0.81 0.458 arvense [0.1] [0.8]

2006 0.53 0.13 1.73 0.13 0.73 1.33 0.63 0.436 0.55 0.583 1.17 0.328

[0.5] [0.1] [1.6] [0.1] [0.4] [0.4]

Cyperus 2005 0.47 0.27 0.8 [0.2] 0.07 0 [0] 0 [0] 6.03 0.022 4.52 0.022 2.98 0.07 0.904 0.066 0.026 strigosus [0.2] [0.2] [0.1]

2006 0 [0] 0.47 0.27 0.13 0 [0] 0 [0] 0.54 0.471 0.92 0.41 1.44 0.256

[0.4] [0.2] [0.1]

Daucus carota 2005 0.07 1.2 0.4 [0.3] 3.93 1.53 4.93 9.65 0.005 3.27 0.055 0.18 0.836

[0.1] [0.6] [2.3] [0.8] [2.1]

2006 0 [0] 2.2 0.53 6.27 [3] 2 [0.7] 5.67 17.6 <0.00 2.92 0.073 1.45 0.254

[0.9] [0.5] [3.8] 9 1

Dicots 2005 0.73 1.2 1.47 1.93 1.6 [0.4] 2.07 2.6 0.12 3.47 0.047 0 1 0.118 0.056 0.925 unknown [0.2] [0.4] [0.2] [0.4] [0.4]

2006 2.2 2.13 9 [1.7] 6.6 7.4 [1.5] 3.87 4.89 0.037 13.2 <0.00 1.27 0.298 <0.00 0.012 0.145

[0.6] [0.4] [0.8] [1.1] 1 1

Elymus 2005 0 [0] 0 [0] 0 [0] 0.13 0 [0] 0 [0] 2.67 0.116 2.67 0.09 2.67 0.09 riparius [0.1]

2006 0.07 0.6 0 [0] 0 [0] 0 [0] 0 [0] 3.76 0.064 5.88 0.008 3.76 0.038 0.018 0.018 1.000

[0.1] [0.3]

Euthamia 2005 0 [0] 3.53 0.2 [0.1] 3.8 0.27 0.33 37.5 <0.00 5.95 0.008 8.92 0.001 0.207 <0.00 0.023 graminifolia [0.9] [1.4] [0.2] [0.3] 5 1 1

2006 0 [0] 11.4 0 [0] 5.73 0.2 [0.2] 1.13 174. <0.00 40.9 <0.00 44.1 <0.00 0.055 0.039 0.986

[0.9] [0.8] [0.5] 86 1 1 2 1

Eupatoriadelp 2005 0 [0] 0 [0] 0.67 0.6 1.2 [0.3] 2.13 0.96 0.342 5.44 0.033 1.27 0.276 0.947 0.025 0.012 hus maculatus [0.4] [0.4] [0.6]

2006 0 [0] 0 [0] 0.87 1.2 1.53 0.67 0.27 0.61 4.29 0.026 1.08 0.355 <0.00 <0.00 0.002

[0.3] [0.5] [0.8] [0.3] 1 1

Juncus effusus 2005 2.6 5.87 1.93 1.87 0.87 0.4 [0.2] 1.22 0.28 6.55 0.005 2.06 0.149 0.073 0.004 0.434

[0.8] [1.9] [1.1] [0.7] [0.5]

2006 10.4 17.4 11.53 3.67 12.8 0.33 0.67 0.42 4.71 0.019 1.81 0.186 0.229 0.014 0.37

[4.3] [5.2] [6.4] [1.4] [7.8] [0.3]

Juncus tenuis 2005 0.8 0.87 3.33 1.6 [1] 2.07 0.93 2.09 0.162 2.15 0.138 0.67 0.521

[0.2] [0.6] [1.4] [0.6] [0.5]

2006 0.33 2.53 4.13 3.47 4.13 2.13 0.03 0.871 2.21 0.131 1.71 0.202

[0.3] [1.5] [0.7] [0.6] [2.1] [0.8]

Ludwigia 2005 0 [0] 0 [0] 0 [0] 0 [0] 0 [0] 0 [0] ...... palustris 2006 0.53 0 [0] 0 [0] 0 [0] 0 [0] 0 [0] 3.88 0.061 3.88 0.035 3.88 0.035 0.06 0.06 1.000

[0.3]

Lythrum 2005 0.13 0 [0] 0 [0] 0 [0] 0 [0] 0 [0] 1 0.327 1 0.383 1 0.383 salicaria [0.1]

2006 2 [1.1] 0 [0] 0 [0] 0 [0] 0 [0] 0 [0] 3.21 0.086 3.21 0.058 3.21 0.058

Mentha 2005 0 [0] 0 [0] 0.07 1.53 0.93 5.53 1.99 0.171 1.84 0.18 0.9 0.421 spicata [0.1] [1.2] [0.6] [4.1]

2006 0 [0] 0 [0] 0 [0] 7.6 0.07 13.4 [9] 3.9 0.06 1.22 0.314 1.19 0.32

[5.6] [0.1]

Monocots 2005 0.4 0.47 1.4 [0.3] 2.47 2.13 2.93 5.9 0.023 22.2 <0.00 1.27 0.299 <0.00 <0.00 0.176

Unknown [0.2] [0.1] [0.4] [0.2] [0.5] 1 1 1

2006 0.93 1.93 2.73 7 [0.4] 4.73 [1] 5.47 7.46 0.012 10.5 0.001 2.41 0.111 0.002 0.001 0.963

[0.3] [0.4] [0.8] [1.6]

Oxalis stricta 2005 0 [0] 0 [0] 0.33 0.27 0.47 0.2 [0.1] 1.04 0.318 3.79 0.037 0.54 0.589 0.083 0.05 0.966

[0.2] [0.2] [0.2]

2006 0 [0] 0 [0] 0.33 1.2 0.6 [0.3] 1.13 1.93 0.178 2.65 0.091 0.56 0.577

[0.1] [0.8] [0.5]

Panicum 2005 0 [0] 0 [0] 0 [0] 0 [0] 0 [0] 0 [0] ...... acuminatum 2006 0 [0] 0 [0] 0 [0] 1.33 0.6 [0.5] 11.87 4.47 0.045 3.96 0.033 3.2 0.059 0.96 0.044 0.077

[0.7] [5.9]

Phalaris 2005 1.2 2.07 4.07 1.47 1.6 [1.4] 4.33 0.07 0.787 0.46 0.634 1.64 0.214 arundinacea [0.6] [0.9] [2.4] [1.1] [1.8]

2006 2.47 5.73 9.53 6.53 7.87 3.2 [1.7] 0.5 0.484 0.24 0.792 0.05 0.949

[1.2] [2.9] [7.1] [3.4] [7.8]

Plantago 2005 0 [0] 0 [0] 0.2 [0.1] 0.87 0.4 [0.2] 1.27 2.82 0.106 2.57 0.098 0.74 0.487 lanceolata [0.6] [0.6]

2006 0 [0] 0 [0] 0.07 1.47 0.6 [0.3] 2.2 [1.4] 3.37 0.079 4.45 0.023 0.84 0.443 0.351 0.017 0.278

[0.1] [1.2]

Plantago 2005 0 [0] 0 [0] 0.07 0 [0] 0 [0] 0 [0] 1 0.327 1 0.383 1 0.383 major [0.1]

2006 0 [0] 0 [0] 0 [0] 0.33 0 [0] 0 [0] 1 0.327 1 0.383 1 0.383

[0.3]

Poa palustris 2005 0.2 1.93 1.13 2.53 6.6 [2.7] 11.8 5.87 0.023 20.5 <0.00 1.12 0.342 0.451 1.000 1.000

[0.2] [0.8] [0.5] [0.9] [1.6] 1

2006 0 [0] 6.33 3.27 8.6 22.6 18.47 6.19 0.02 8.45 0.002 3.5 0.046 0.023 0.023 1.000

[1.9] [1.5] [3.7] [9.8] [7.2]

Poa pratensis 2005 0 [0] 0 [0] 0.33 0.47 0 [0] 0 [0] 0.15 0.701 5.43 0.011 0.15 0.861 1.000 0.057 0.057

[0.3] [0.2]

2006 0 [0] 0 [0] 0.27 1.6 0.32 0.02 [0] 1.8 0.192 4.96 0.016 3.79 0.037 1.000 0.208 0.208

[0.2] [0.7] [0.2]

Polygonum 2005 1.8 0 [0] 0.87 0 [0] 0.67 0 [0] 10.2 0.004 1.02 0.377 1.02 0.377

93 hydropiperoid [0.9] [0.4] [0.3] 9 es 2006 0.17 0.01 [0] 1.93 0 [0] 0 [0] 0 [0] 3.02 0.095 2.35 0.117 2.37 0.115

[0.1] [1.2]

Polygonum 2005 0 [0] 0 [0] 0.33 0.4 1.87 [1] 0.73 0.98 0.332 4.65 0.02 1.18 0.326 0.85 <0.00 <0.00 pensylvanicum [0.3] [0.2] [0.3] 1 1

2006 0 [0] 0 [0] 0.8 [0.6] 1.67 1.67 1.4 [0.6] 0.17 0.684 3.73 0.039 0.5 0.615 0.242 0.001 0.058

[0.9] [0.8]

Populus 2005 0 [0] 0 [0] 0.13 0 [0] 0 [0] 0 [0] 1 0.327 1 0.383 1 0.383 deltoides [0.1]

2006 0 [0] 0 [0] 0.8 [0.3] 0 [0] 0 [0] 0 [0] 5.43 0.028 5.43 0.011 5.43 0.011 0.018 0.848 0.060

Potentilla 2005 0 [0] 0 [0] 0 [0] 0 [0] 0.73 0.27 0.86 0.363 3.95 0.033 0.86 0.436 0.686 0.018 0.106 canadensis [0.5] [0.2]

2006 0 [0] 0 [0] 0 [0] 0 [0] 1.87 0 [0] 2.04 0.166 2.04 0.152 2.04 0.152

[1.3]

Prunella 2005 0 [0] 0 [0] 1 [0.4] 0.4 2 [1] 0.4 [0.3] 3.23 0.085 2.91 0.074 1.31 0.289 vulgaris [0.4]

2006 0 [0] 0 [0] 0.8 [0.8] 1.6 1.6 [1.6] 0.27 0.06 0.804 1.06 0.363 0.77 0.473

[1.1] [0.3]

Sagittaria 2005 1.4 0.4 0 [0] 0 [0] 0 [0] 0 [0] 2.26 0.146 7.33 0.003 2.26 0.126 0.008 0.008 1.000 latifolia [0.6] [0.3]

2006 0.06 [0] 0.02 [0] 0 [0] 0 [0] 0 [0] 0 [0] 0.9 0.353 3.57 0.044 0.9 0.421 0.073 0.073 1.000

Scirpus 2005 10.6 7.13 1 [0.8] 0.53 0 [0] 0 [0] 0.37 0.548 6.96 0.004 0.26 0.777 0.014 0.007 0.954 cyperinus [4.8] [4.2] [0.5]

2006 21.2 [9] 37 1.2 [0.8] 0.47 0 [0] 0 [0] 0.1 0.752 22 <0.00 0.41 0.666 <0.00 <0.00 0.577

[10.8] [0.2] 1 1 1

Scirpus validus 2005 0.13 0.73 0 [0] 0 [0] 0 [0] 0 [0] 1.05 0.317 2.18 0.135 1.05 0.367

[0.1] [0.6]

2006 0.04 [0] 0.07 0 [0] 0 [0] 0 [0] 0 [0] 0.07 0.792 1.97 0.161 0.07 0.931

[0.1]

Sisyrinchium 2005 0 [0] 0 [0] 0.27 0.13 0.53 0 [0] 3.17 0.087 1.65 0.213 1.65 0.213 angustifolia [0.2] [0.1] [0.3]

2006 0 [0] 0 [0] 2.33 0.6 0.73 0.07 2.15 0.155 4.69 0.019 0.57 0.575 0.016 0.549 0.144

[1.7] [0.3] [0.6] [0.1]

Solidago 2005 0.6 2.2 5.53 8.07 11.2 7.87 0.02 0.89 6.26 0.007 0.91 0.417 0.074 0.005 0.483

95 canadensis [0.5] [1.4] [1.8] [2.1] [4.6] [1.5]

2006 0 [0] 3.07 3.87 12.4 11.93 18 [3.5] 2.69 0.114 4.67 0.019 0.19 0.825 0.308 0.015 0.285

[1.5] [3.5] [3.8] [8.6]

Symphyotrichu 2005 0 [0] 0 [0] 0 [0] 0 [0] 0 [0] 0 [0] ...... m lanceolatum 2006 0 [0] 0.53 0.53 3.6 0.13 1.67 5.28 0.031 1.07 0.358 0.17 0.847

[0.3] [0.5] [2.8] [0.1] [1.1]

Symphyotrichu 2005 0.4 0.87 1.67 4.07 3.8 [1.4] 4.47 0.79 0.382 5.49 0.011 0.47 0.628 0.056 0.011 0.758 m pilosum [0.3] [0.6] [0.7] [1.7] [2.4]

2006 0.27 0.6 2.13 5.53 4.07 [1] 9.13 6.51 0.017 14.8 <0.00 0.67 0.519 0.021 <0.00 0.046

[0.3] [0.3] [1.6] [2.4] [1.6] 1 1

Taraxacum 2005 0 [0] 0 [0] 0.07 0.33 0.4 [0.2] 0.53 1.07 0.31 4.42 0.023 0.36 0.703 0.425 0.018 0.228 officinale [0.1] [0.2] [0.3]

2006 0 [0] 0 [0] 0.13 0.73 0.33 0.47 2.75 0.11 3.57 0.044 1.52 0.238 0.061 0.089 0.981

[0.1] [0.3] [0.2] [0.2]

Trifolium 2005 0.33 1.67 1.13 1.07 1.73 0.67 0 0.993 0.46 0.639 1.53 0.237 pratense [0.3] [1.2] [0.7] [0.5] [0.5] [0.3]

2006 0 [0] 0.67 0 [0] 0.15 0.2 [0.2] 0.6 [0.4] 2.29 0.143 0.54 0.592 0.3 0.742

[0.7] [0.1]

Trifolium 2005 0 [0] 0 [0] 0.07 0.53 0.13 0.27 1.41 0.247 1.1 0.35 0.68 0.517 repens [0.1] [0.5] [0.1] [0.2]

2006 0 [0] 0 [0] 0.53 3 [1.9] 0.53 2.07 1.93 0.178 2.85 0.077 0.48 0.623

[0.3] [0.4] [1.1]

Ulmus 2005 0 [0] 0 [0] 0.2 [0.2] 0.2 0.73 0.6 [0.2] 0.06 0.806 4.86 0.017 0.06 0.94 0.638 0.015 0.106 americana [0.2] [0.4]

2006 0 [0] 0 [0] 0.27 1.27 0.8 [0.5] 1.67 3.06 0.093 4.09 0.030 0.78 0.471 0.204 0.024 0.54

[0.3] [0.5] [0.7]

The 2-way ANOVA did not detect a difference of total plant biomass between treatments in 2005 (Figure 2). However, there was a difference among elevation zones.

The Low zone had the highest biomass, and the Mid and High zones had similar biomass

(Table 3). The amount of bare ground was higher in Flood Pulse wetlands than in Static

Wetlands in both years (Figure 3). There was no difference of bare ground among elevation zones in either year (Table 3). Species richness was similar in all treatments throughout the study, but it was different among zones in 2005. None of the pair wise comparisons of richness between zones were significantly different, although the comparison between Low and Mid Zones was almost significant: p <0.051 (Figure 4,

Table 3). No significant interactions were detected for any comparison (Table 3).

The proportions of plants grouped by WIS were very different between elevation zones and treatments (Figure 5). Obligate Wetland (OBL) plants were a greater portion of the total community in Flood Pulse wetlands (Table 3). Facultative Wetland (FACW),

Facultative (FAC) plants and Upland (UPL) plants had greater proportions in Static wetlands. The OBL taxa were more abundant in the Low Zone, and Facultative Upland

(FACU), FAC, and UPL plants were more abundant in the Mid and High zones (Table 3).

No significant interactions were detected for any comparison (Table 3).

The NMDS ordination showed clear separation of the different treatments and elevation zones and explained a high amount of the observed variance among plant communities in 2005 and 2006 (Figure 6). In 2005, the best NMDS model was a 2- dimensional ordination that explained 84.3% of the observed variation in plant

Fig.6. Mean (+/-1 SE) total biomass per sample (0.33 m2 quadrant) in the treatment / elevation zones in 2005. Flood pulse and Static wetlands were not different. Elevation zones with different letters are significantly different.

300 2005 Flood Pulse a Static

250

200 b 150 a,b

Biomass (g) Biomass 100

0 Low Mid High

100

Fig.7. Mean (+/-1 SE) Percent bare ground in the treatment / elevation zones in 2005 and 2006. Flood Pulse wetlands had more bare ground than Static wetlands in both years. Elevation zones were not different in 2005 or 2006.

101

16 2005 Flood Pulse

14 Static

4 Percent Bare Ground Bare Percent 2

0 Low Zone Mid Zone High Zone

30 2006

10 Percent Bare Ground Bare Percent 5

0 Low Zone Mid Zone High Zone

102

Fig.8. Mean (+/-1 SE) species richness per sample in the treatment / elevation zone habitat types in 2005 and 2006. Richness was different in 2005, but pairwise Tukey’s tests were not significant.

103

18 Flood Pulse 2005 16 Static

14 12 10 8 6

Species RIchness Species 4 2 0 Low Zone Mid Zone High Zone

20 2006 18

14 12 10 8

6 Species Richness Species 4 2 0 Low Zone Mid Zone High Zone

104

Fig.9. Plant community in each treatment / elevation zone habitat type in 2005 and

2006. Values are the mean percent (+/-1 SE) of total plant cover that was made up by each Wetland Indicator Status (WIS) category. FP = Flood Pulse wetlands, ST= Static wetlands. Percentages do not add up to 100% because the amount of bare ground is not included.

105

106

Fig.10. NMDS ordinations of plant communities in the treatment / elevation zone habitat types. Symbols are: Flood pulse/Low; Static/Low; Flood pulse/Mid;

Static/Mid; Flood pulse/High; Static/High.

107

108

Table 6. 2-way ANOVAs comparing communities characteristics in treatment / elevation zone habitat types. WIS are plant cover grouped by their Wetland Indicator Status.

Statistics are results of 2-way ANOVAs for treatment (Flood pulse vs. Static), zone (Low vs. Mid vs. High Zone), and treatment x zone interaction effects. Tukey’s HSD tests were run when there was a significant zone effect.

109

Treatment Zone Interaction Tukey’s test

Mid Low vs. Low vs. vs. Mid High F1,24 p F2,24 p F2,24 p High

Bare Ground 2005 20.27 <0.001 2.42 0.110 0.36 0.701

2006 28.78 <0.001 2.27 0.125 2.35 0.117

Richness 2005 0.48 0.497 3.54 0.045 0.07 0.929 0.051 0.12 0.902

2006 0.78 0.385 2.72 0.086 0.13 0.882

Biomass 2005 1.39 0.250 4.26 0.026 0.67 0.522 0.038 0.059 0.978

WIS

Obligate 2005 6.46 0.018 11.88 <0.001 0.27 0.763 0.003 <0.001 0.155

110

2006 3.00 0.096 6.75 0.005 1.44 0.256 0.13 0.001 0.047

Facultative Wetland 2005 8.64 0.007 1.08 0.356 0.15 0.864

2006 1.82 0.190 3.03 0.067 1.98 0.161

Facultative 2005 2.19 0.152 8.15 0.002 0.03 0.967 0.003 0.001 0.648

2006 7.77 0.010 9.79 0.001 0.38 0.689 0.016 <0.001 0.083

Facultative Upland 2005 0.04 0.836 11.60 <0.001 3.39 0.051 0.001 <0.001 0.345

2006 1.77 0.196 5.52 0.011 0.17 0.844 0.029 0.004 0.376

Upland 2005 0.95 0.338 4.11 0.029 0.63 0.540 0.063 0.097 0.825

2006 5.32 0.030 2.28 0.124 1.00 0.383

111 communities. Axes 1 and 2 explained 67.1% and 17.2%, respectively of the observed variation (Figure 6). NMDS stress was 13.3, which indicates a moderately high goodness of fit in this model. In 2006, the best NMDS model was a 2-dimensional ordination that explained 77.6% of the observed variation in plant communities. Axes 1 and 2 explained

70.7% and 6.90%, respectively of the observed variation (Figure 6). NMDS stress was

13.1, which indicates a moderately high goodness of fit in this model.

The elevation zones and treatments were delineated along the two ordination axes. The plant communities in the Flood pulse / Low and Static / Low habitat types were the most widely separated groups in both years. In 2005 and 2006, MRPP indicated that all elevation zones in Flood Pulse wetlands were different from each other (Table 4). In contrast, the three elevation zones in Static wetlands were not different from each other in both years. Furthermore, Low and High zones in Flood

Pulse wetlands were distinctly different from all elevation zones in Static wetlands.

Indicator Species Analysis (Table 5) showed which species were associated with each treatment / elevation zone habitat type. A maximum indicator value of 100% occurs when a species is abundant in all replicates of a treatment / elevation zone habitat type. In this experiment, indicator values ranged from 23% to 75%. Indicator species in the Flood Pulse / Low zone were mostly OBL plants (e.g. Carex lupulina,

Polygonum hydropiperoides, Sagittaria latifolia) but OBL (C. crinita, C. lurida) and FACW plants (Eupatoriadelphus maculatum, Juncus effusus, Scirpus cyperinus) were indicator species in the Static / Low zone. In the Flood Pulse / Mid zone, indicator species were

112

Table 7. Results of pairwise comparisons of plant communities with MRPP tests in 2005 and 2006. Treatments were Flood Pulsing (FP) and Static (ST); Zones were Low, Mid,

High elevation zones. The T statistic describes the separation between the groups.

Large negative T values indicate the groups are well separated. The agreement (A) statistic describes within-group similarity: A = 1 if all samples within a treatment group have the same species composition; A = 0 if the samples have the same amount of heterogeneity as expected by chance; and A < 0 if heterogeneity is greater than expected by chance. The p value is the probability that the T statistic was this value by chance alone.

113

2005 2006

T A p T A p

FP Low vs. ST Low -3.121 0.102 0.009 -4.740 0.158 0.002

FP Low vs. FP Mid -3.381 0.100 0.008 -3.054 0.076 0.008

FP Low vs. ST Mid -5.179 0.229 0.002 -5.450 0.219 0.002

FP Low vs. FP High -4.986 0.195 0.002 -4.299 0.158 0.003

FP Low vs. ST High -5.519 0.300 0.002 -5.565 0.277 0.002

ST Low vs. FP Mid -3.676 0.093 0.002 -4.058 0.139 0.005

ST Low vs. ST Mid -3.296 0.089 0.007 -3.868 0.150 0.007

ST Low vs. FP High -4.397 0.125 0.001 -3.824 0.144 0.006

ST Low vs. ST High -4.936 0.189 0.002 -4.862 0.233 0.003

FP Mid vs. ST Mid -2.920 0.074 0.013 -2.382 0.057 0.020

FP Mid vs. FP High -0.876 0.020 0.184 0.201 -0.007 0.542

114

FP Mid vs. ST High -4.633 0.157 0.002 -4.025 0.124 0.003

ST Mid vs. FP High -0.636 0.015 0.229 -0.859 0.018 0.189

ST Mid vs. ST High -1.380 0.041 0.094 -0.792 0.024 0.195

FP High vs. ST High -0.920 0.026 0.167 -1.345 0.033 0.099

115

Table 8. Indicator species in each treatment / elevation zone habitat type. Treatments were Flood Pulsing (FP) and Static (ST); elevation zones were Low, Mid, High elevation zones. WIS is the Wetland Indicator Value for each species. Indicator Value is a measure of relative frequency and abundance within a habitat type. A maximum value of 100% would occur if a species if found at high numbers in all replicates of a habitat type.

116

Treatment/ zone Indicator Species WIS Indicator Value P value

2005

FP/Low Bareground - 27.4 0.016

Carex lupulina OBL 38.7 0.004

Polygonum hydropiperoides OBL 45.1 0.015

Sagittaria latifolia OBL 74.6 <0.001

ST/Low Agrostis alba FACW 60.0 0.015

Carex crinita OBL 60.0 0.015

Carex lurida OBL 60.0 0.014

Eupatoriadelphus maculatus FACW 39.9 0.014

Juncus effusus FACW 32.2 0.007

117

Scirpus cyperinus FACW 38.6 0.032

Scirpus validus OBL 46.8 0.044

FP/Mid Bidens cernua OBL 33.0 0.045

Cyperus strigosus FACW 47.6 0.008

ST/High Daucus carota FAC 33.7 0.055

Euthamia graminifolia FAC 38.0 0.014

Unidentified Monocot spp. - 24.7 0.024

Poa palustris FACW 36.5 0.002

Ulmus americana FACW 40.2 0.016

FP/Low Bareground - 35.6 0.048

Carex lupulina OBL 31.8 0.045

Ludwigia palustris OBL 60.0 0.017

Lythrum salicaria FACW 60.0 0.013

118

Polygonum hydropiperoides OBL 68.8 0.001

Sagittaria latifolia OBL 80.0 0.001

ST/Low Elymus riparius FACW 50.2 0.014

Eupatoriadelphus maculatus FACW 49.5 <0.001

Scirpus cyperinus FACW 41.8 0.028

FP/Mid Unidentified Dicot spp. - 23.2 0.028

Polygonum pensylvanicum FACW 60.0 0.015

ST/Mid Acer rubrum FAC 80.0 0.001

Unidentified Monocot spp. - 24.6 0.014

ST/High Andropogon gerardii FAC 45.3 0.045

Panicum acuminatum FAC 54.5 0.005

Plantago lanceolata UPL 45.7 0.008

Solidago canadensis FACU 34.1 0.043

Symphyotrichum pilosum FACW 34.3 0.006

119

OBL (Bidens cernua) or FACW (P. pensylanicum, Cyperus strigosus). There were no indicators in the Static-Mid zone or in the Flood pulse-High zone. In the Static / High zone, indicator species ranged widely in their tolerance of flooding. For example, indicator species were FACW (Poa palustris, Ulmus americana, Symphylotrichum pilosum), FACU (Solidago canadensis), FAC (Daucus carota, Euthamia graminifolia), and

UPL (Plantago lanceolata).

Discussion

The Flood pulse and Static treatments created distinctly different environmental conditions for the riparian plant community. Flooding from the headwater creek frequently inundated all but the highest elevation zones in the Flood Pulse wetlands, and flood events in these mesocosms lasted longer than in Static wetlands. The Static wetlands were flooded into the Mid zone less than 2% of the time in both years, whereas the Mid zone in Flood Pulse wetlands were flooded 16-23% of the time.

Therefore, the stresses and benefits of flooding varied in intensity along the elevation gradient and between treatments.

Although I used mesocosms, I feel I tested of the impacts of hydrological stresses in headwater wetlands under realistic conditions. First, plant communities had several years to become established on the banks of the mesocosm wetlands. For example, I

120 only seeded 20 species in 2002-2003, but 98 taxa were present by 2006. Second, the timing of flood events in the mesocosms was caused by natural fluctuations of the headwater stream. Third, water depth changes after flooding were up to 135 cm, which is within the range reported in natural floodplains (Shipley et al. 1991; Johansson

& Nilsson 2002; Bayley & Guimond 2009). Therefore, the results of this study probably closely approximate the complex dynamics of biotic responses to water fluctuations indicative of natural headwater systems in the temperate climate of the northeastern

United States.

When I compared species assemblages, I detected pronounced changes at the population and community level among elevation zones and treatments. This supported both of the hypotheses and indicated that the flood-pulsing hydrology acted as an environmental filter to select for specific plant traits. The first hypothesis predicted a divergence in plant communities in Flood pulse and Static mesocosms. I found that 17 common taxa had different abundances between treatments.

Furthermore, the differences in species occurrence could be related to their life history traits. Indicator species in the Static treatment ranged from Obligate wetland plants

(OBL) to Upland plants (UPL), where Flood Pulse wetlands were dominated by OBL plants. These traits are consistent with the expectation that soils in the Flood Pulse wetlands would be exposed to prolonged and repeated flooding, which would exclude species that do not tolerate soil anoxia.

121

I expected that taxa richness would be higher in the Flood Pulse wetlands because water level fluctuation usually increases habitat heterogeneity and biodiversity

(Junk et al. 1989; Tockner 2000; Brose 2001; Malard et al. 2006). I found evidence that flood pulsing increased habitat heterogeneity because there were distinct communities in each elevation zone (see below) in Flood Pulse wetlands suggesting that microhabitat differences were starting to develop. However, richness was similar in the Flood Pulse wetlands and Static wetlands. Furthermore, in a related study (Drinkard et al. In press),

I found that Shannon’s Diversity was higher in Static wetlands than Flood Pulse wetlands, which was the opposite of my expectations. There are several potential explanations for this pattern. First, I conducted the study in newly constructed mesocosms, whereas other studies sampled well-established plant communities.

Diversity can fluctuate widely in the first few years of community establishment.

Therefore, I cannot predict the long-term changes in response to the treatment.

Second, and more importantly, the intermittently flooded zone in the mesocosms was only ~2 m wide. Most natural riverine systems have wider floodplains that can have many different microhabitats (Shipley et al. 1991; Bayley & Guimond 2009). Thus, natural floodplains are large enough to support more types of plant communities

(Schneider 2001; Chase & Leibold 2002). It is also possible that flood-pulsing reduces biodiversity at small-scales (within patches), but the habitat heterogeneity caused by flood-pulsing increases biodiversity at large scales (between patches). For example,

Desilets & Houle (2005) found that flood pulsing increased diversity across the entire

122 floodplain but reduced diversity near the river edge where stresses were most intense.

I acknowledge, therefore, that the mesocosms tested within-patch dynamics of plant communities, but large-scale field trials are still necessary to understand regional impacts of flood-pulsing.

In my second hypotheses, I predicted that distinct plant zonation would develop along the elevational gradient. As expected, over half (29 taxa) of the common species in the Flood Pulse wetlands had different abundances among elevation zones. For example, many sedges (Carex lurida, Cyperus strigosus, Scirpus cyperinus) and rushes

(Juncus effusus) were abundant in the Low zone and rare or non-existent in the High zone. These taxa were all OBL or FACW, indicating that they tolerated soil anoxia. In contrast, many flood intolerant forbs (Potentilla canadensis, Plantago lanceolata,

Solidago canadensis, Symphyotrichum pilosum, Taraxacum officinale) and grasses

(Andropogon gerardii, Panicum acuminatum) became increasingly prevalent in the High zones. Abrupt changes in life history traits along elevation gradients have been found in salt marshes where soil salinity changes due to evapotranspiration (Pennings &

Callaway 1992; Silvestri et al. 2005) and freshwater systems where soil anoxia is a common abiotic factor (Galatowitsch & van der Valk 1996; Vivian-Smith 1997; Seabloom

& van der Valk 2003).

The indicator species analysis showed that flood-tolerant OBL species were only found in Low zone in Static wetlands, but in the Low and Mid zones in Flood Pulse wetlands. Thus, flood-pulsing promoted wetland plants occurrence at higher elevations

123 than the static treatment. Furthermore, NMDS found species assemblages in the Low,

Mid and High elevation zones were distinct in Flood-pulsing wetlands but not in Static wetlands. Therefore, the impacts of elevation interacted with the treatment effect.

Others have found plant zonation was restricted to areas with frequent flooding

(Seabloom et al. 1998; Dwire et al. 2004; Bowles et al. 2005; Zelnik & Carni 2008), but it is important to note that I detected differences across relatively minor differences in hydrology. For example, plant communities in High elevation ones were different between Flood pulse and Static wetlands even though flooding times were similar (e.g.,

3% in Flood pulse / High elevation zone vs. <1% in Static / High elevation zone).

Although impacts of flood pulsing are widely known to occur in large order systems

(e.g., Junk et al. 1989), the study shows the importance of short stochastic flooding in headwater systems on plant community assembly.

I also predicted that more frequent and longer periods of inundation at lower elevations would affect plant biomass and species richness. As predicted I did find higher biomass in the Low zone. Many plants common in the Low zone such as Scirpus cyperinus and Juncus effusus formed dense robust clumps (M. Drinkard, pers. observ.).

This was probably because species that tolerate saturated soils can benefit from the nutrient inputs in this zone (Jackson & Colmer 2005). I also expected that plant diversity would be lower in Low zone. In both years, richness was slightly lower in the

Low zones (10-11 species/sample) than the Mid and High zones (12-15 species/sample).

There was a significant difference among zones in 2005, but none of the pair wise

124 comparisons with Tukey’s tests were significant. However, the comparison of Low zone and Mid Zone biomass was marginally non-significant (Tukey’s test, P<0.051).

Therefore, the general pattern supported the hypothesis that flood pulsing will reduce diversity through stressful conditions caused by repeated submergence during floods.

Flood-pulsing can have both beneficial (elevated soil moisture, nutrient subsidies, reduced competition) and stressful (soil anoxia, sedimentation, erosion) impacts in riparian wetlands (Tockner et al. 1999; Benke et al. 2000). In this study, I did not directly test the underlying causes that created the plant zonation, but I have strong indirect evidence suggesting that flood-pulsing in this headwater system acted as a stressor. There was more bare ground and lower plant richness in the Flood Pulse wetlands, and these attributes are correlated with higher habitat stress in riparian systems (Toogood & Joyce 2009). The stresses of flooding probably increased seedling mortality in the wetlands, because mature plants usually survive short flood events

(Blom & Voesnek 1996; Blom 1999). Therefore, some of the impacts I observed may be different in wetlands dominated by mature perennial plants (e.g. forested floodplains).

Management Implications

Riparian wetlands are vitally important for many species of fish, waterfowl, and invertebrates, but less is known about headwater streams than high-order river systems. The results can help guide the development and application of management

125 strategies for headwater ecosystems. I found that the flood-pulsing hydroregime created by the second order creek favored plant species with desirable traits. For example, many species that were more abundant in Flood Pulse wetlands (Bidens cernua, Carex lurida, C. vulpinoidea, Cyperus strigosus, Polygonum hydropiperoides,

Sagittaria latifolia) are important food plants for waterfowl (Fredrickson & Taylor 1982;

Delnicki & Reinecki 1986; Sheeley & Smith 1989; Strader & Stinson 2005). Taxa that are wetland indicators (i.e. OBL, FACW) were found at higher elevations in Flood Pulse wetlands than Static wetlands, which shows that flood-pulsing should create more wetland habitat. Therefore, floodplains connected with the river channel may provide better habitat for wetland-dependent species than floodplains that are diked to stabilize water levels.

Furthermore, most taxa in Flood Pulse wetlands were native plants. In contrast, several common taxa in Static treatments were exotic weeds: Agropyron repens (quack grass), Plantago lanceolata (narrowleaf plantain), Daucus carota (Queen Anne’s Lace)

(USDA, NRCS 2011). Others (e.g., Kercher et al. 2007) found that a natural flood- pulsing hydrology in riparian wetlands promoted native wetland species and reduced noxious pests. However, I also found that exotic purple-loosestrife, Lythrum salicaria, was an indicator species in the Low elevation zone in Flood Pulse wetlands. This species is an aggressive invasive in wetlands throughout the U.S. (Blossey et al. 2001). It can reduce wetland diversity by establishing monotypic stands, especially in disturbed habitats (Blossey et al. 2001). Possibly, the higher amounts of bare ground in the Flood

126

Pulse wetlands allowed its seedlings to become established. Therefore, it would be useful to further test if the amount of bare ground caused by flooding regimes in riparian wetlands affects the distribution of Lythrum and other noxious species.

127

Acknowledgements

I thank Brendan Morgan, Karen Montgomery, Constance Hausman, Jennifer

Clark, Douglas Kapusinski, Justin Montemarano, Leonard N. Drinkard II, and Leonard N.

Drinkard III for field assistance. This research was supported by funds from the Art and

Margaret Herrick Aquatic Ecology Research Facility grant program.

CHAPTER 4

USING MESOCOSMS TO TEST IF FLOOD-PULSING AFFECTS EMERGENCE OF AQUATIC

INSECT IN HEADWATER WETLANDS

Abstract

Hydroregime is an important factor in insect emergence in riparian wetlands. In this study, I examined the impact of hydrology on invertebrate emergence in four habitat types along headwater creeks. I used floating traps to sample invertebrates that emerged from: 1) a permanent pool in a natural riparian wetland, 2) an intermittently flooded zone in the same natural riparian wetland, 3) mesocosms with a

Flood-pulsing hydrology, and 4) mesocosms with a static hydrology. Chironomid midges were the most common taxa collected at all habitats; several dipterans and a

128

129

few other invertebrate taxa correlated with each habitat types. Multivariate analyses found that invertebrate communities were different between the permanent pool and intermittently flooded zone in the natural wetland. This shows that flood-pulsing hydrology of headwater creeks created unique macroinvertebrate assemblages in different floodplain microhabitats. Communities were also different between the habitats at the natural wetland versus the mesocosm wetlands, but not between the

Flood pulseand Static mesocosms. Lack of a difference between mesocosm habitats suggests that the flood-pulsing hydrology had minimal impact on habitat conditions in the permanently flooded pools. The pronounced difference in invertebrate communities in the natural and mesocosm wetlands is more difficult to explain. It may be due to artificial habitat conditions found in the man-made mesocosms. Alternately, it might show an effect of ecological factors such as habitat scale or watershed characteristics, which could not be controlled for in the experimental design.

Introduction

The Flood Pulse Concept (Junk et al. 1989) was developed to test the vital linkages when rivers spill over their bank into the adjacent floodplain. Extensive floods can occur in large order river systems, which create a moving littoral zone that inputs organic matter, invertebrate, and fish into floodplain pools (Junk et al. 1989, Sparks

1995). This increases primary and secondary productivity in flood-pulsing systems.

130

Fluctuating water levels also increase habitat heterogeneity, which leads to higher biodiversity in floodplain ecosystems (Harper et al. 1997, Robinson et al. 2002, Ward et al. 1999).

Headwaters systems play a key role in enhancing downstream water quality and are important to local species diversity and flood control (Tronstad et al. 2005b,

Gladden and Smock 1990). Although many studies have examined flood-pulsing effects in large order rivers, the timing and duration of flooding in headwaters is very different

(Junk 1989, Benke 2001). Riparian floodplains in headwater systems are smaller and have less predictable and shorter flood regimes than in large rivers. Therefore, headwater floodplain communities may respond differently to flood-pulsing than in large rivers, and therefore studies on impacts of flooding on these systems would be valuable.

Invertebrates are an integral component of floodplain biotic communities. They are important components in wildlife food webs, and they affect detrital breakdown rates (Batzer and Wissinger 1996, Malmqvist 2002). As floodwaters recede, aquatic insects must survive concentrations of harmful chemicals (e.g. ammonia, hydrogen sulfide, ferrous iron) and low dissolved oxygen levels in the wetland pools (Mitsch and

Gosselink 2002). During drawdowns, aquatic invertebrates are also exposed to stresses caused by desiccation during stranding. Furthermore, erosion and sedimentation caused by rapid flood surges is problematic for sessile invertebrates (Brock et al. 2003,

Dietz-Brantley et al. 2002, Tronstad et al. 2005b). For example, eggs in the soil are

131

highly susceptible to sedimentation, which occurs in flashy, unpredictably flooded system (Dietz-Brantley et al. 2002, Tronstad et al. 2005b). Thus, aquatic invertebrate abundance is usually higher in wetlands with predictable flood patterns and long-term flooding (Batzer and Wissinger 1996).

Aquatic invertebrates in river floodplains have morphological, physiological, and behavioral adaptations to survive fluctuating habitat conditions (Batzer and Wissinger

1996). Some species survive draw-downs by migrating to deep floodplain pools or to nearby aquatic habitats (e.g., Chironomidae, Hemiptera, Coleoptera).

Microcrustaceans, Aedes (Culicidae), and annelid worms enter into a desiccation resistant dormant stage (Stanczak and Keiper 2004). Many taxa (e.g. Culicidae,

Chironomidae) use high reproductive rates to reestablish populations after they are reflooded (Adis and Junk 2002, Batzer and Wissinger 1996, Naiman and Decamps 1997).

Some taxa undergo a facultative early emergence to escape when conditions become stressful (Merritt and Cummins 1996, Tronstad et al. 2005a). For example, odonate nymphs respond to increased temperature, conductivity, and lower dissolved oxygen during drawdowns by emerging earlier as adults (Tolonen et al. 2003, Trondstad et al.

2005a, Boulton and Lloyd 1992). Odonates also select oviposition sites with stable water levels over those with fluctuating water levels (Bazzanti et al. 2003).

This project compared aquatic insect communities in different floodplain microhabitats along headwater streams. I sampled adult emerging insects in two microhabitats in the floodplain of a small stream: permanent pools and intermittently

132

flooded areas. I also manipulated the hydrology in mesocosms to create mesocosms with flood-pulsing hydrology and others with static water levels. I sampled emerging insects in these mesocosms to examine the impact of the flood-pulsing treatment and to compare the insect communities with the natural floodplain microhabitats. My hypotheses were:

H1: Different communities of emerging insects will be found in different microhabitats

in floodplains. Because hydrology is an important environmental factor, species

assemblages in permanently flooded habitats with fluctuating water levels will

differ greatly from those in intermittently flooded areas or in wetlands with

static water levels.

H2: Intermittent dewatering will act as a stressor to many aquatic species, and thus

communities in intermittently flooded areas will have lower abundance and

diversity than in permanently flooded pools.

Methods

Study Site

The wetland mesocosms used in this experiment were located at Art and

Margaret Herrick Aquatic Ecology Research Facility (HAERF) on the Kent State University

133

campus. The HAERF was constructed in 2002, and it includes ten large earthen wetland mesocosms (10 M X 20 M). Each mesocosm could be independently flooded or drained with inlet and outlet water control structures from Allerton Creek, a second order headwater stream. Water control structures were a stop log design that used stacked

PVC boards to maintain water inputs or outflow at different depths. More information about HAERF is provided in Drinkard et al. (2011).

Inlet water-control structures in all mesocosms were set that water could enter from Allerton Creek during floods caused by storms. Mesocosms were randomly assigned into two habitat types (n=5 each habitat type). In Static wetlands, outlet water control structures were set that the floodwater inputs quickly flowed out and the wetland pool depth remained at a baseline level of ~80cm. Therefore, the water levels rarely inundated the mesocosm’s earthen banks, but the central pool was permanently flooded. In Flood Pulse wetlands, outlet control structures were set to allow the mesocosms to flood to a depth of 140 cm. There was a hole drilled in one of the PVC boards to allow water to slowly flow out until the depths decreased to a baseline level of 80 cm. Water levels in Flood Pulse wetlands usually remain elevated for 4-10 days after a moderate to large storm event.

In 2006, I sampled intermittently flooded wetlands and permanently flooded pools along a headwater creek at Mud Brook Preserve (MBP). The MBP is a 30 hectare preserve in Summit Co., Ohio, which is located 16 km from HAERF. The natural levees along the creek are low, and the floodplain became flooded each time water levels in

134

the creek increased at least 20 cm. At higher elevations, the 500 m wide floodplain was inundated up to 30 cm after large rainstorms, but it was only flooded for 1-3 days after most storm events. In the floodplain, there are small deep pools that remain flooded when the creek water level drop back within its banks. I sampled a 60-cm deep pool that remained flooded throughout the study and the adjoining floodplain. The soils in the intermittently flooded area were flooded frequently and remained saturated throughout the study.

Experimental Design

I compared aquatic invertebrates in four different habitat types at MBP and

HAERF. The hydrology of the Flood Pulse wetlands at HAERF (henceforth HAERF-FP) was the most similar to the permanent pool at MBP (henceforth (MBP-PFP) because both had permanent flooding and contact with the surrounding floodplain during flood events. The two other habitat types had different hydrologies. The Static wetlands at

HAERF (henceforth HAERF-ST) rarely had connection with their floodplain (i.e. the earthen banks of each mesocosm), and the intermittently-flooded zone at MPB

(henceforth MBP-IFZ) was frequently drawn down. Therefore, I expected these two habitats to be different from the first two habitat types and each other.

Hydrology

135

At HAERF, I measured water levels with staff gauges placed in each mesocosm.

Levels were measured 3 times each week from January 2005 to October 2006. At MBP,

I measured water levels with a meter stick when the site was visited during sampling throughout 2006.

Adult insect emergence

I sampled aquatic invertebrates to measured temporal and spatial differences in abundance. I collected invertebrates that emerged from the water’s surface using a floating emergence traps that had a similar design to those by Walton et al. (1999).

The wood frame of the trap was a four-sided pyramid on a square wooden frame (0.5 m x 0.5 m base). The sides were fiberglass insect netting (1 mm mesh size) stapled to the wooden frame. In the mesocosms at HAERF, one emergence trap was set in each

HAERF-FP wetland (n=5) and each HAERF-ST wetland (n=5) at three dates during the growing season in 2005 (5 May, 24 June, 21 July), and 2006 (25 May, 27 July, and 14

September). In MBP, I placed emergence traps at six random locations in MBP-PFP and also MBP-IFZ on three dates in 2006 (26 May, 21 July, 13 September).

All emergence traps remained in the wetland for one week. All invertebrates that were found inside the trap were removed with an aspirator and preserved in 70% ethanol. In the laboratory, these were counted and identified to family in 2005 and to family or genus in 2006 using taxonomic keys (Merritt and Cummins 2008, Thorp and

136

Covich 2001, Johnson and Triplehorn 2004). I defined the dominant taxa as those that comprise at least 2% of all invertebrates collected in a habitat type.

Statistical Analyses

All data were tested for normality using Kolmogorov-Smirnov tests. Non-normal data were log (x+1) transformed. Transformed data were retested for normality, and I used the data type that best approximated normality. Abundances of dominant taxa, total numbers and taxa richness were compared between habitats with Repeated

Measures ANOVA. If treatment effects were significant, I ran Tukey’s HSD tests to make pairwise comparisons of means. All univariate statistics were run on SPSS software

(SPSS version 15.0, 2007).

I also compared the overall community structure invertebrate with multivariate statistics. I used non-metric multidimensional scaling (NMDS) using Sørensen's distances to visually examine for differences in community structure among habitats.

In 2005, I compared the two habitat types at HAERF, and in 2006, I compared the four habitat types at HAERF and MBP. Ordinations were run with the data pooled across all sampling dates in a year, and I included habitat type as a covariate. Next, Multi-

Response Permutation Procedure (MRPP) was used to test if invertebrate communities differed between habitat types in 2005 or 2006. I also identified indicator taxa that correlated with the habitat types using an Indicator Species Analysis (Dufrene and

137

Legendre 1997). The Indicator Species Analysis technique works by determining which taxa are both abundant and frequently occur in a specific habitat type. Multivariate statistics were run with PC-ORD software, version 5 (McCune and Mefford 2006).

Results

Hydrology

In 2005, the flooding regimes in HAERF-FP and HAERF-ST wetlands were clearly different (Figure 1). After storms, HAERF-FP wetlands usually were 20-50 cm above the permanent pool, and it took 4-10 days for water levels to draw down. HAERF-ST wetlands remained near the permanent pool levels except when intense storms (<5 cm/hr) raised water levels for a few hours. In both habitat types, water levels fell slightly below the baseline level of the permanent pool during the dry summer months, but neither habitat was ever completely dewatered. MBP water levels were measured in May, June and July 2006. On these dates, water levels at MBP and HAERF followed a similar pattern of high and low water levels. For example, water levels at all sites would rise after a large rainstorm, and all sites returned to pre-flood levels within 1 to 4 days.

As expected, water levels in the IFZ at MBP would decrease to below the soil surface within several after flood events, but the PFP was never dewatered. There were also

138

Fig.11. Water levels in 2005 and 2006 at Mud Brook Preserve and at the Herrick Aquatic

Ecology Research Facility. Water levels at HAERF are recorded as change from baseline water level in the pools at HEARF. Water levels at MBP were water depths taken at random locations in each habitat type. Habitat types are Flood pulse (HAERF-FP) and

Static wetlands (HAERF-ST) at Herrick Aquatic Ecology Research Facility and

Permanently Flooded Pool (MBP-PFP) and Intermittently Flooded Zone (MBP-IFZ) sites at Mud Brook.

139

140

extended periods of flooding at HEARF and MBP during a period when there were repeated rain storms.

Invertebrate communities

I collected 4,947 invertebrates in 21 families at HAERF in 2005. In 2006, I collected 3,353 invertebrates in 46 families at HAERF and 2,051 invertebrates in 38 families at MBP. The most abundant taxa in all habitat types were chironomid midges

(Diptera: Chironomidae). In 2005, they comprised 94% of all invertebrates in HAERF-FP wetlands and 84% in HAERF-ST wetlands. In 2006, they were 80% of the total collected in HAERF-FP, 79% in HAERF-ST, 82% in MBP-PFP, but only 31% of all invertebrates in

MBP-IFZ. In 2005, the only other dominant taxon was a crane fly (Tipuloidae). In

2006, dominant taxa included several chironomid midges, biting midges

(Ceratopogonidae), dolichipodid flies (Dolichopodidae), collembolans (Entomobryidae), aquatic mites (Acariformes Hydracarina), and other Diptera.

In 2005, total numbers were not different between the two habitat types (Figure

2, Table 1). The mean total number of invertebrates across the 2005 sampling dates were ~120-200 invertebrates per trap. In 2006, total numbers were somewhat lower, and mean numbers per trap were ~ 25-130 invertebrates at HAERF and MBP (Figure 2).

In 2006, total numbers were not different among the habitat types (Figure 2, Table 1).

141

Fig. 12. Invertebrate community numbers are HAERF and MBP in 2005 and 2006. A.

Mean (SE) of total numbers of invertebrates per emergence trap. B. Mean (SE) of the number of species per emergence trap. Habitat types are Flood pulse (HAERF-FP) and

Static wetlands (HAERF-ST) at Herrick Aquatic Ecology Research Facility and

Permanently Flooded Pool (MBP-PFP) and Intermittently Flooded Zone (MBP-IFZ) sites at Mud Brook.

142

300 A. p= All n.s. HAERF FP 250 HAERF ST

200 MBP IFZ

150 MBP PFP

100 Abundance

0 2005 2006

14 B. p= All n.s.

4 Species Richness Species 2

0 2005 2006

143

Table 9. Results of RM-ANOVAS comparing total abundance and species richness at

HAERF in 2005 and in HAERF and MBP in 2006. In 2005, Flood pulse and Static wetlands at Herrick Aquatic Ecology Research Facility were compared. In 2006, Flood pulse and

Static wetlands at Herrick Aquatic Ecology Research Facility and Permanently Flooded

Pool and Intermittently Flooded Zone sites at Mud Brook were compared.

144

Habitat Type Interaction

Year F df p F df p

Abundance 2005 0.764 1,8 0.408 1.103 2,16 0.356

2006 2.828 3,16 0.072 1.996 6,32 0.095

Species Richness 2005 1.09 1,8 0.327 2.537 2,16 0.11

2006 1.81 3,16 0.189 1.401 6,32 0.247

145

Invertebrate biodiversity was not different among habitat types in 2005 or 2006 (Figure

2, Table 1). In 2005, I collected 4-5 taxa per trap in both habitats at HAERF (Figure 2).

In 2006, richness was slightly higher, and I collected 8-10 taxa per trap in all habitats

(Figure 2).

A few of the dominant taxa had different population densities among habitat types. At HAERF, only two taxa were dominant in 2005 (Diptera, Chironomidae,

Tipuloidea), and they were present in similar densities in both treatments. In 2006, numbers of most taxa were not different between treatments at HAERF and MBP.

However, aquatic mites (Acariformes Hydracarina) were more abundant in HAERF-FP, and a crane fly (Tipuloidea, F. Limnoniidae, Erioptera) was more abundant in the

HAERF-ST. A chironomid midge (Tanypodinae, Species 2), was less abundant in MBP-

IFZ than MBP-PFP. Another chironomid midge (Species 1), had a treatment effect, and had highest numbers in MBP-PFP, but the Tukey’s Test did not detect significant differences among habitat types.

Multivariate community analyses

In 2005, the 2 dimensional ordination of the NMDS analysis was significant

(NMDS: stress = 0.8, p < 0.01, Variance = 93.6%) (Figure 3). Most samples varied along axis 1, which explained almost all of the variance between samples. However, the ordination did not distinguish between invertebrate communities in the HAERF-FP and

146

Fig. 13. Invertebrate communities at Herrick Aquatic Ecology Research Facility and Mud

Brook Preserve in 2005 and 2006. In 2005, Flood pulse (HAERF-FP) and Static wetlands

(HAERF-ST) at Herrick Aquatic Ecology Research Facility were sampled. In 2006, Flood pulse and Static wetlands at Herrick Aquatic Ecology Research Facility and Permanently

Flooded Pool (MBP-PFP) and Intermittently Flooded Zone (MBP-IFZ) sites at Mud Brook were sampled. Percent of variance explained by the axes of the 2 dimensional NMDS ordinations are shown on the plots.

147

148

HAERF-ST habitat types. The MRPP analysis also did not detect a difference among communities (MRPP: A = 0.017, p = 0.202), and there were no indicator species for either habitat type.

In 2006, the NMDS analysis compared communities at HAERF habitats (FP and

ST) with those at MBP habitats (IFZ and PFP) (Figure 3). The 2-dimensional solution was significant (NMDS: stress = 17.4, p < 0.01, Variance = 81.7%). Samples collected in MBP were distinguishable from those in HAERF along Axis 2, which explained 42% of the variance of the data. The MBP-PFP and MBP-IFZ samples were clearly distinguishable along Axis 1, which accounted for 40% of the variance. MRPP pairwise comparisons found that communities in all habitats were different, except HAERF-FP and HAERF-ST communities were not different (Table 3).

Indicator taxa analysis revealed that several flies (Diptera), a damselfly (Odonata) and aquatic mites (Acariformes, Hydracarina) were indicators of different habitats

(Table 5). Three chironomid midges and aquatic mites were associated with HAERF-FP habitats. Damselflies (Odonata, Coenagrionidae) and cranefly (Diptera, Tipuloidea,

Limoniidae, Erioptera) were indicators in HAERF- ST habitats. A second cranefly

(Tipluloidea, Cylindrotomidae, sp.1) was the only taxa associated with MBP-IFZ. Taxa that associated with MBP-PFP were 3 species of Chironomidae, biting midges (Diptera,

Ceratopogonidae) and a third cranefly (Tipuloidea, Tipulidae, Tipula). Of these, several species were indicator taxa but were not dominant taxa (Odonata, Coenagrionidae;

149

Table 10. Abundance of dominant taxa at Herrick Aquatic Ecology Research Facility and

Mud Brook Preserve in 2005 and 2006. In 2005, Flood pulse (HAERF-FP) and Static wetlands (HAERF-ST) at Herrick Aquatic Ecology Research Facility were compared. In

2006, Flood pulse and Static wetlands at Herrick Aquatic Ecology Research Facility and

Permanently Flooded Pool (MBP-PFP) and Intermittently Flooded Zone (MBP-IFZ) sites at Mud Brook Preserve were compared. Values are the mean [1 standard error] number per trap on each sampling date. Statistics from the Repeated Measures ANOVAs are given. P values are bold when treatments were significant different (p<0.05). Letters

(a, b, c) above means show differences among treatments using Tukey’s tests.

150

Treatment Date X Treatment HAERF -FP HAERF -ST MBP-IFZ MBP-PFP F df p F df p 2005 Diptera: 178.27 108.13 1.01 0.34 1.70 2,1 0.21 Chironomidae spp. [12.94] [11.43] - - 7 1,8 3 9 6 2 1.20 0.30 1.10 2,1 0.32 Tipuliodea 0.07 [0.51] 8.6 [5.43] - - 3 1,8 5 5 6 4

2006 Arachnida 10.67 [4.22] 0.12 2.25 [2.17] 5.21 1,1 0.01 3.45 4,3 0.04 Acariformes Hydracarina a 3.73 [2.54] b [0.58] b b 2 8 6 2 6 8

Collembola 1.94 1.71 1,1 0.20 1.84 4,3 0.18 Entomobrya 7.87 [4.3] 2.73 [1.7] [1.98] 0.25 [0.67] 4 8 8 3 6 3

Diptera: 7.71 0.40 1,1 0.67 4.36 4,3 0.00 Ceratopogonidae 0.6 [1.36] 2.33 [2.5] [2.98] 6.88 [3.21] 3 8 4 9 6 6 Chironomidae 29.93 [6.51] 22.13 [5.52] 5.65 38.88 1,1 0.04 2.07 4,3 0.10 Chironomidae sp. 1 a a [3.11] a [6.74] a 3.72 8 4 5 6 4 0.41 1,1 0.66 0.61 4,3 0.65 Chironomidae sp. 2 9.4 [3.83] 14.13 [4.69] 0 [0] 0 [0] 1 8 9 2 6 7 0.71 11.56 1.07 1,1 0.36 0.70 4,3 0.59 Chironomidae sp. 3 26.6 [7.45] 16.87 [4.33] [1.08] [3.59] 3 8 3 7 6 2 0.12 11.31 1.83 1,1 0.18 0.87 4,3 0.45 Chironomidae sp. 4 26.87 [7.28] 9.73 [3.58] [0.58] [3.57] 1 8 9 1 6 1

151

2.55 1,1 0.10 1.26 4,3 0.30 Chironomidae sp. 5 6.87 [2.78] 6.67 [3.19] 2 [1.55] 8.94 [3.37] 3 8 6 6 6 7 2.81 1,1 0.08 0.62 4,3 0.63 Tanypodinae sp. 1 1.4 [2.27] 1.87 [2.2] 0 [0] 4 [2.79] 4 8 6 9 6 5 3.31 [2.69] 1,1 0.04 3.22 4,3 0.02 Tanypodinae sp. 2 1 [1.33] a, b 0.8 [1.17] a, b 0 [0] a b 3.75 8 4 7 6 3 1.06 0.83 1,1 0.45 1.54 4,3 0.21 Dolichopodidae 0.33 [0.79] 0 [0] [1.12] 1.63 [1.73] 4 8 0 9 6 0 Tipuloidea Limoniidae 0.24 4.31 1,1 0.02 4.30 4,3 0.02 Erioptera 0 [0] a 4.73 [3.33] b [0.66] a 0.06 [0.5] a 5 8 9 5 6 9

152

Table 11. P-values of MRPP pair-wise comparisons between habitats at HAERF and

MBP. In 2005, Flood pulse (HAERF-FP) and Static wetlands (HAERF-ST) at Herrick

Aquatic Ecology Research Facility were compared. In 2006, Flood pulse and Static wetlands at Herrick Aquatic Ecology Research Facility and Permanently Flooded Pool

(MBP-PFP) and Intermittently Flooded Zone (MBP-IFZ) sites at Mud Brook were compared.

153

MBP IFZ MBP PFP HAERF FP HAERF ST

2005

HAERF FP - - - 0.202

2006

MBP IFZ - <0.001 <0.010 <0.001

MBP PFP - - <0.050 <0.050

HAERF FP - - - 0.1877503

154

Table 12. Indicator species of the habitat types at HAERF and MBP in 2006. Indicator values (IV%) are based on relative abundance and relative frequency, and values of

100% are the best indicators for a given habitats.

155

Habitat Indicator taxa IV % p value

HAERF FP Acariformes Hydracarina 32.3 0.041

Diptera Chironomidae Tanytarsini sp. 1 21.4 0.0124

Diptera Chironomidae sp. 2 31.1 0.0126

Diptera Chironomidae sp. 3 33.0 0.0314

HAERF ST Diptera Tipuloidea Limoniidae Erioptera 25.2 0.035

Odonata Coenagrionidae 26.0 0.0484

MBP IFZ Diptera Tipuloidea Cylindrotomidae sp.1 26.7 0.0126

MBP PFP Diptera Ceratopogonidae 30.5 0.0324

Diptera ChironomidaeTanypodinae sp. 1 33.3 0.002

Diptera ChironomidaeTanypodinae sp. 3 40.0 0.0004

Diptera Chironomidae sp. 4 32.7 0.0338

Diptera Tipuloidea Tipulidae Tipula 20.0 0.05

156

Chironomidae, Tanytarsini sp. 1; Tipuloidea Tipulidae, Tipula; Tipluloidea,

Cylindrotomidae, sp. 1).

Discussion

Several studies have found that flood-pulsing hydrology is a key environmental factor that determines wetland community structure (Benke 2001, Malmqvist 2002,

Robinson et al. 2002), but few have tested this in headwater systems. In this study, I tested the impact of flood-pulsing hydrology by comparing invertebrate communities in the intermittently flooded zone and permanently flooded pools of a natural floodplain wetland and in similar habitats in created mesocosms. Both sites experienced anthropogenic alteration. The mesocosms at Herrick Aquatic Ecology Research Facility were created about 5 years before this study. However, they now support a diverse aquatic invertebrate community (Drinkard et al. 2011). The floodplain at MBP was abandoned agricultural land, but it has largely re-naturalized after farming ceased in the

1980’s. Both locations were flooded from adjacent headwater streams after storms, and the timing of flooding at each site was similar. Furthermore, each location flooded about 60 cm above baseline water levels after large storms. Thus, I could use these data to test impacts of flood-pulsing and also provide information on the limitations and advantages of using mesocosms to examine community-level changes.

157

The natural flood-pulsing hydrology caused differences in emergence patterns of invertebrates in the habitat types I sampled at MBP. The 2006 NMDS and MRPP analyses clearly distinguished the MBP-PFP and MBP-IFZ communities. Indicator taxa analysis found that invertebrates associated with MBP-PFP were several chironomids, a crane fly (Tipuloidea) and a ceratopogonid midge. Another crane fly was associated with MBP-IFZ. A dominant midge also had different densities in these habitat types.

This supported the Hypothesis #1 that invertebrate communities would vary with habitat type in floodplain habitats. Others have examined large-order river floodplains and described changes in invertebrate (Batzer and Wissinger 1996), plants (Battaglia and Collins 2006) and general biodiversity of communities (Bunn and Arthington 2002).

Thus, I show that the short, stochastic flooding regime in the headwater creek at Mud

Brook Preserve also affected invertebrate communities by creating different microhabitat conditions within the floodplain.

Chironomid midges were dominant taxa at all the habitat types I sampled.

Chironomid midges are well adapted to many aquatic habitats types and they are common in most wetland (Oliver 1971). Additional common taxa were other true flies

(Diptera). For example, in Mud Brook Preserve, a crane fly (Tipuloidae,

Cylindrotomidae, Sp. 1) was an indicator species in the IFZ, and several midges, a different crane fly, and biting midges (Ceratopogonidae) were indicators of the PFP.

The changes in the taxa occurred because hydrology acted as an environmental filter

158

that restricted the local species pool (Fairchild et al. 2003). Although the presence or absence of drawdowns was probably the most significant factor this would affect several abiotic conditions such as water chemistry, soil moisture, and nutrient cycling.

In this experiment, I did not examine the response of individual species to each abiotic condition, so I cannot determine which filters were the most important.

Although the difference in hydrology between the IFZ and the PFP at Mud Brook

Preserve affected community structure, it did not change overall richness or total abundance. I had hypothesized that these values would be lower in habitats exposed to stresses during intermittent drawdowns. Chase (2007) showed that dry periods in aquatic systems can serve to filter intolerant species and thereby reduce biodiversity for the intermittently flooded microhabitat. However, many wetland invertebrate species are generalists and can survive under variable habitat conditions (Batzer and Wissinger

1996). For example, aquatic microcrustaceans such as copepods and ostracods enter into a facultative desiccation-resistant dormancy during short-term drawdowns

(Trondstad et al. 2005b). Since many of the dominant taxa were in both habitat types, they must be adapted to both permanent flooding and intermittent drawdowns. As a result, total diversity and total numbers were not greatly affected even though community structure was altered as numbers of individual taxa changed. Another potentially important factor is that species distributions may have been occasionally homogenized across the floodplain during flood events. The habitats were only 50 m

159

apart and floodwaters would physically connect the permanent pool and the IFZ.

Transport of invertebrates across floodplains during floods is an important component of riparian systems. For example, chironomids and other invertebrates can be washed into a floodplain from upstream sources (Beingo and Sommer 2008, Lamberti and

Chaloner 2010). This exchange is seasonally variable (Rundio and Lindley 2011) but has a significant long term impact on the mature community of these systems. (Thomaz et al. 2007, Tockner et al. 2010). The redistribution of organisms by flooding decreases as one moves away from the stream channel (Bright et al. 2010), but the habitats were within 70 m of Mud Brook.

Unlike the results at Mud Brook Preserve, there were no major differences in macroinvertebrate communities between the two habitat types I sampled at HAERF mesocosms. The flood pulse treatment was designed to test the impact of exchange between the mesocosm banks and the permanent pools during floods, but the 2005 and

2006 multivariate analyses did not detect differences in invertebrate communities between the Flood Pulse and Static habitats. Nor were there differences in total numbers or species richness detected in either year. There were a few differences among some taxa: water mites and three chironomid midge species were associated with HAERF-FP, and a crane fly and a coenagrionid damselfly species were associated with HAERF-ST. Thus, the community structure was not greatly altered, in spite of an obvious difference in hydrology between the two habitat types.

160

There are several potential reasons for this lack of an impact of the flood-pulsing treatment. In a previous study, I sampled the water column in Flood pulse and Static wetlands at HAERF, and found that benthic invertebrate communities were similar in these habitat types (Drinkard et al. 2011). In that study, I also found that the emergent plant community on the mesocosm banks responded strongly to flood-pulsing

(Drinkard et al. 2011). The lack of a response in the aquatic invertebrate community that study and this one suggests that microhabitat conditions in the permanently flooded pools were not greatly affected by the hydrology treatment. I did not sample water chemistry in 2006, but dissolved oxygen, pH, conductivity, and temperature were similar in both habitat types in 2005 (Drinkard et al. 2011). Another potential reason for the lack of an effect is that any impacts of flood-pulsing on trophic resources or physical structure could have been minor. For example, overland flow can be an important source of nutrient subsidies to floodplain wetlands during river flooding.

Also, inorganic sediments and organic matter (e.g. coarse woody debris) are redistributed during floods, which adds to habitat heterogeneity. It is widely believed that both factors have a strong impact on macroinvertebrate biodiversity in floodplains

(Polis 1997, Huryn and Wallace 2000, Baxter et al. 2005, Alsfeld et al. 2009), which suggests that the flood-pulsing treatment did not greatly alter these factors. Perhaps the lack of a wide floodplain at HAERF limited the impact of flood-pulsing.

Furthermore, Palmer et al. (2010) recently conducted a meta-analysis of stream restoration projects and found that habitat heterogeneity did not, by itself, have a

161

strong impact on macroinvertebrate biodiversity. This suggests that the current paradigm that habitat heterogeneity in floodplains will always influence macroinvertebrate diversity should be re-evaluated in future studies.

I also hypothesized that communities in permanently flooded sites with fluctuating water levels (i.e. PFP-MBP and HAERF-FP) would be the most similar, and they would differ substantially from those with stable water levels (HAERF-ST) or with drawdowns

(MBP-IFZ). Thus, it is surprising that the communities in both habitats at Herrick

Aquatic Ecology Facility were similar and each was different than both habitats at Mud

Brook Preserve. For example, the 2006 NMDS and MRPP analyses showed clear differences between the Flood Pulse wetlandsat HAERF and the PFP at MBP. Thus, my hypothesis that the two permanently flooded habitat types would be similar was not supported by the data.

There are several potential reasons that HAERF-FP and MBP-PFP did not have similar communities. The most obvious is that HAERF habitats were man-made mesocosms, and this affected some important environmental conditions. The use of mesocosms in ecological research has been debated (Carpenter 1996). Although some studies show mesocosm can simulate how algae and invertebrates respond in natural aquatic systems

(Drenner and Mazmuder 1999, Spivak et al 2005), Carpenter (1996) concluded that common design problems of mesocosms reduce their usefulness in ecological research.

162

If the difference between the HAERF-FP and MBP-PFP habitats was caused by a limitation of the mesocosm design, perhaps the newly created mesocosms at HAERF lacked some habitat or biotic characteristic that was fully developed in a natural wetland. For example, the mesocosms were constructed in upland soils, and hydric wetland soils can take decades to become established (Craft et al. 2002). This would lead to different soil and water chemistry in the mesocosms and the natural floodplain wetlands. Invertebrate communities become established over many years in man- made wetlands (Stanczak and Keiper 2004, Batzer et al. 2006, Stewart and Downing

2008), and the results may have been influenced by the age of the mesocosm wetlands.

However, mesocosms are widely used in ecological research, and there is much evidence that they are a useful tool to test impacts on whole systems. Thus invertebrate communities may not have been greatly affected by the mesocosm approach. One important factor that should be accounted for in any ecological study is the effect of habitat scale on community structure (Drenner and Mazumder 1999).

Scale-dependent effects occur among natural habitats and mesocosms (Spivak et al.

2011). For example, small mesocosms may have greater temperature fluctuations than large mesocosms (Ahn and Mitsch 2002). Thus the scale differences between HAERF

(10 m X 20 m basins) and MBP (the PFP was about 40 m X 125 m) could be important than if I sampled man-made or natural habitats. Although I know that this was a

163

potential problem when I designed this experiment, there were no small permanent pools at Mud Brook that could have been used in this study.

Another potential reason why the biotic communities at HAERF and Mud Brook

Preserve were different can be attributed to watershed characteristics. Most of the watershed of Allerton Creek at HAERF is the Kent State University Campus. Thus there many potential sources of urban pollution such as automobile oil from parking lots, herbicides, fertilizers, and sediments. Mud Brook’s watershed includes a mix of forested areas, agricultural land, and subdivisions, which would have a different impact on water quality, and therefore, may support a different aquatic invertebrate community.

Overall Implications

In this experiment, I found that wetland invertebrate communities vary among habitat types within a headwater floodplain. Under the natural hydrology at Mud Brook

Preserve, invertebrate communities showed clear responses to flood permanence.

Further analysis is needed to test which are the key abiotic conditions that drive community changes. In previous studies I found that mesocosms were a useful tool to test the impact of flood-pulsing on emergent macrophytes. However, in this

164

experiment, limitations of the mesocosm experimental design may have reduced some of the potential effects of flood-pulsing. Thus, further studies should examine how flood-pulsing effects are influenced by the redistribution of materials within the floodplain.

165

Acknowledgements

I would like to thank Brendan Morgan, Jennifer Clark, Doug Kapusinski, Justin

Montemarano, Leonard Neil Drinkard 2 & 3, and Natasha Wingerter, Mike Woods, and

Emily Faulkner for their help in the field and lab. This grant was support by the Art and

Margaret Herrick Endowment fund for the Herrick Aquatic Ecology Research Facility.

CHAPTER 5

THE IMPORTANCE OF COARSE WOODY DEBRIS AS INVERTEBRATE MICROHABITAT IN

HEADWATER FLOODPLAINS

Abstract

Coarse woody debris (CWD) from fallen trees is an important physical component of riparian floodplains. Decaying logs create a unique microhabitat moderating the environmental stresses caused by intermittent flooding and dewatering.

During drawdowns, aquatic and semi-aquatic macroinvertebrate use CWD as a refuge from desiccation stress. I examined effects of CWD on soil dwelling macroinvertebrates in the floodplain along a headwater stream, Mud Brook, in Ohio. In Spring 2006, I simulated tree-fall events by placing 100 cm X 35 cm (L X Dia.) logs in the floodplain

166

167

(Added CWD) and comparing them to nearby areas without logs (Added Control). I also sampled natural logs (Natural CWD) and areas without logs (Natural Control) in a different location in the floodplain. In 2006 and 2007, I sampled soil invertebrates with pitfall traps and measured abiotic conditions in all treatments. Abiotic conditions were not different under CWD and their associated Control treatment, but they were different in the area with Natural CWD and the area with Added CWD. Invertebrate species richness was also not different between CWD treatments and the associated

Control areas. However, there were again differences between the areas with Natural

CWD and Added CWD. Invertebrate abundance did not show any consistent effect of the presence of CWD or floodplain location. Multivariate analyses show that there were different invertebrate communities in the four treatments, but the most consistent differences were between the two locations in the floodplain. Therefore, location within the floodplain had a stronger impact than presence of CWD, suggesting that abiotic factors that were not influenced by fallen logs were probably important.

Indicator species analysis found some differences among treatments. Some detritus feeders (e.g. ptilid beetles, collembolan springtails) were associated with Natural CWD or Added CWD treatments on some dates. However, herbivores and predators were also associated with CWD, and therefore there was no clear impact of the CWD on trophic groups. Presence of logs may not have been important because CWD could have been moved during floods or because logs were not have decayed enough to influence soil environmental conditions. Furthermore, the effect of shading by dense

168

vegetation may have masked any impact of CWD. Although I found few local-scale impacts of CWD on invertebrate communities in this floodplain wetland, a useful future study could test habitat-wide impacts of CWD.

Introduction

Forested riparian floodplains providing provide ecologically and economically important services because they store river overflow during floods, provide fish and wildlife habitat, act as sinks and transformers of nutrients, and serve as a source of resource subsidies for adjacent lotic habitats (Junk et al. 1989, Bayley 1995). Habitat attributes such as landscape heterogeneity (Pollock et al. 1998), composition of the forest community (Batzer et al. 2005, Engelhardt and Ritchie 2002) and connectivity to lotic systems (Nilsson and Svedmark 2002) affect the ecological integrity of the floodplains and the services they provide (Bunn and Arthington 2002, Ward and

Stanford 1995, Ward et al. 1999).

Forested riparian zones can be stressful environments for the biota because floods erode soils, remove detritus and uproot plants (Gleason et al. 2003, Kozlowski 2002).

Floods also bring in sediments that bury the seed and invertebrate egg banks, which can be harmed by burial of as little as 0.5 cm of sediments (Gleason et al. 2003).

Submergence and subsequent soil anoxia also inhibit plant growth (Kennedy et al.

169

1992) or cause morphological changes such as increased aerenchyma tissue (Moon et al.

1993). After floodwaters recede, desiccation becomes a stress for aquatic and semi- aquatic animals and plants in the drying floodplain wetlands. Floodplain biota respond by seeking refugia within the floodplain, migrating to other habitats, becoming dormant or timing their lifecycles to avoid unfavorable conditions (Johansson and Nilsson 2002,

Kozlowski 2002). For example, many aquatic invertebrates survive the dry period in resting stages in moist soil (Boulton and Lloyd 1992, Tronstad et al. 2005 a,b) and terrestrial invertebrates raft on floating wood debris during floods (Braccia and Batzer

2001).

Standing snags and fallen logs of dead trees are termed coarse woody debris

(CWD). In floodplains, CWD is an integral structural component that affects flow dynamics, habitat stability, sedimentation patterns, and biotic communities (Gurnell et al., 1995, Naiman and Decamps 1997). The CWD is added to the floodplain by catastrophic events (e.g. fires, logging or floods), herbivory (e.g. beavers or insect infestation) or by natural tree death (Bragg 2000). The longevity of CWD in floodplains varies, but usually it is a stable and lasting component of the habitat (Abbe and

Montgomery 1996). Although some studies found that most CWD was less than 5 years old, others found that logs existed in the floodplain for 100 years or more (Jones and

Daniels 2008, Naiman and Decamps 1997).

The presence of CWD increases floodplain heterogeneity and creates unique microhabitats, which can affect the habitat characteristics. The physical structure of

170

CWD promotes sediment deposition and increases soil stability because it reduces flow velocity (Gippel 1995, Nilsson and Svedmark 2002, Gurnell et al. 2002). Different size classes of CWD provide different ecological functions: larger logs are more likely to create scour pools that are aquatic habitat but smaller logs break down faster and provide several microhabitat for detritivores (Laser et al. 2009)).

The geomorphology of the watershed will affect CWD characteristics and its impact on the floodplain. Size and abundance of CWD typically increases with stream order (Gurnell et al. 2002, McIllroy et al. 2008), but local geomorphology and biological factors (e.g., tree density, elevational gradient, and local climate) alter CWD inputs and log longevity regardless of stream order (Gomi et al. 2002, Piegay 1997). In mid-order streams, CWD alters channel morphology and increases sediment storage in the floodplain because logs deflect river flow (Gurnell et al. 2002). In contrast, in large order rivers, CWD is often relocated during large floods and floating logs have less effect on sediment storage (Nakamura and Swanson 1993). CWD also alters floodplain geomorphology in high gradient headwater streams when logs contact the stream banks, (Gurnell et al. 2002), but its impact in low gradient headwater systems is not well known.

Humans have diked and channelization most of the world’s major river system

(Dynesius and Nilsson 1994), and often they have decreased the amount of CWD in floodplains (Bragg 2000). The amount of CWD produced in a watershed decreases with urbanization, especially when 20% or greater of the watershed is covered by impervious

171

surfaces (Finkenbine et al. 2000). Higher amounts of CWD are carried into floodplains immediately downstream from logged areas (McIllroy et al. 2007, Thompson et al.

2009), but inputs drop off afterwards until the forest has time to regenerate (Webster et al. 2008).

Studies have found that CWD has a strong impact on plant and animal community structure (Nilsson and Svedmark 2002, Pollock et al., 1998). For example, it increases habitat physical complexity, which increases plant and animal biodiversity and abundance (Harper et al., 1997, Pollock et al., 1998, Spieles and Horn 2009, Vivian-Smith

1997). The decaying wood can create a unique microhabitat that mediates some stresses associated with flooding or desiccation (Pollock et al., 1998). Soils under CWD are usually cooler and moister and have higher organic content than exposed soils

(Naiman and Decamps 1997), and this can increase survival of seedlings (Pettit and

Naiman 2006). Additionally, many animals use standing snags and fallen logs as refuges

(Naiman and Decamps 1997). However, biotic responses to CWD vary among taxa. In one study, removing most of the dead trees from a lake littoral zone caused rapid declines in the dominant fish, Perca flavesces, but no response in macroinvertebrate numbers (Helmus and Sass 2008). Another study found that addition of CWD increased

Odonata and Ephemeroptera but did not affect other invertebrate taxa (Alsfeld et al.

2009).

Invertebrate communities in floodplains are key components of riparian food webs and they affect ecosystem processes such as nutrient cycling (Covich et al. 1999).

172

CWD will mediate some of stressors that affect their populations in riparian wetlands

(Harper et al, 1997, Pollock et al, 1998, Tockner et al., 2002). During drawdowns, aquatic and semi-aquatic invertebrates aestivate under CWD until the floodplain becomes re-flooded (Harper et al., 1997). CWD traps organic debris, and therefore increase food resources for detritivores. Additions of CWD have been used to increase invertebrate biodiversity in streams that were channelized for agriculture (Lester et al.

2007). However, the role of CWD in intermittently-flooded riparian wetlands along headwater systems has not, as yet, been well-studied.

Study Objectives and Hypotheses

The study examined effects of CWD on soil dwelling invertebrates in floodplains along a headwater stream. I tested the impact of CWD by placing logs in vegetated areas of floodplains and monitored the environmental conditions and invertebrate communities below the logs. I also compared these factors at the logs with natural logs in a nearby location that had a different plant community. I tested three main hypotheses:

H1: Stressful environmental conditions caused by drawdowns will be reduced in

sheltered microhabitats under CWD. I expect that there will be lower soil

temperatures, higher soil moisture, and higher organic content in soils below

CWD than in adjacent areas.

173

H2: Due to these abiotic differences, I expect that there will be higher abundance and

diversity of soil dwelling invertebrates in areas with CWD than in exposed areas.

Different species assemblages will make up the invertebrate communities found

under CWD and in exposed areas.

H3: The presence of CWD will have a greater effect on invertebrate communities than

location within the floodplain. Therefore, invertebrate communities under

different types of CWD will be more similar to each other than communities in

different locations within the floodplain.

Methods

Study Site

Mud Brook Preserve (MBP) is managed by the Hudson Land Conservancy located in Hudson, Ohio (Portage Co.). Mud Brook is a 2nd order low-gradient stream in the glaciated northeastern portion of Ohio. The stream frequently overruns its bank after high rainfall or snow melt, and it floods a complex of adjacent wetlands ranging from permanently flooded depressions to intermittently flooded areas. The range of hydroregimes provide habitat for a mix of plant communities including emergent marshes, scrub-shrub wetlands, and forested swamps. Because this is a headwater stream, few

174

logs float in from upstream areas. However, many logs are added from natural mortality of trees growing in the floodplain.

This study took place in an intermittently flooded area that was flooded from

Mud Brook in winter and spring. The soils were exposed by midsummer and remained unflooded except during intermittent floods that lasted several days after large summer storms.

Experimental Design

In spring 2006, I used naturally fallen red maple logs (Acer rubrum) from the floodplain as the CWD. The logs I used were all Decay Class 2 indicating that the bark was not attached, but the wood did not flake off when scraped (Pyle and Brown 1998).

I simulated additions of CWD by cutting ten 100 cm X 35 cm log sections (L X dia.) and placing the sections in randomly chosen sites in the MBP floodplain. Each log was tethered to stakes within a 1 m x 1 m area that was termed the Added CWD treatment

(Figure 1). I randomly chose a control area 2 m away that lacked logs, and marked off a

1 m X 1 m area with stakes. This was termed the Added Control Treatment.

In September 2006, I located 10 naturally felled Decay Class 2 A. rubrum logs in a nearby site in the Mud Brook floodplain. These 30-40 cm dia. logs ranged from 3 m to 15 m long. I staked off a 1 m X 1 m portion of each log, which was termed the Natural

CWD treatment. I also randomly chose a control area 1 to 2 m away that lacked logs

175

that was termed the Natural Control sites. Although these areas were near the Added

CWD treatment, there was a potential effect of location on invertebrate community.

Therefore, I also compare invertebrate communities between the two locations.

I sampled plant communities in all sites in September 2006 and May and

September in 2007. On each sampling date I used the boundaries of the 1 m X 1 m area as a quadrat and sampled plant percent cover of all species and the amount of exposed log in the quadrat. I defined the dominant plant taxa as any that comprised 5% or more of the community for that treatment in each year.

Soil abiotic conditions were measured in September 2006, May and September of 2007 in all four treatments. On all sampling dates, I used a thermometer to measure soil temperature at random locations below CWD or in randomly locations within

Control sites. I also collected two soil cores (3 cm x 10 cm; diameter X depth) in each site. Cores were refrigerated in a Ziploc bags until they were processed in the laboratory. I determined soil moisture by weighing the cores before and after drying

(48 hours at 60 oC). In September 2007, I also estimated soil % organic content by loss on ignition using a muffle furnace at 450 oC. Data from both cores at each site were combined and treated as one sample.

Soil invertebrates were sampled with pitfall traps on September 2006 and May and September 2007. In sites with CWD, one trap was placed at a random location adjacent to the log. In Control sites, traps were placed in randomly locations within the staked area. The pitfall traps were a PVC pipe (10 cm x 15 cm; L X Dia.) embedded in

176

the ground flush with the soil surface. I placed a plastic drinking cup with ~60 ml of soapy water into the pipe, and a plastic funnel was placed into cup to collect invertebrates that fell into the opening of the trap. Traps remained in place for 48-72 hours before the contents were sieved through a 250 micron mesh screen, and trap contents were preserved with 70% ethanol. I used taxonomic and ecologic information in Thorp and Covich (1991), Merritt and Cummins (2008) and Johnson and Triplehorn

(2004) to identify invertebrates to the lowest possible taxonomic unit and determine their typical diet items.

Statistical Analyses

All data were tested for normality with Kolmogornov-Smirnov tests (SPSS, ver.

15.0, 2007) and transformed if needed; temperature and invertebrate data were log (X

+1) transformed and percent plant cover, percent soil moisture and percent organic matter data were arcsine transformed. Abiotic conditions, invertebrate abundance and richness were examined using one way ANOVAs in September 2006 and Repeated

Measures ANOVAs in 2007 (SPSS, ver. 15.0, 2007). If there were significant Treatment

X Date interactions, I ran one-way ANOVAs on each date. When ANOVAs detected a significant difference among treatments, I ran Tukey’s HSD tests to make pairwise comparisons among treatments. Temperature data were taken at different times of day in the areas with natural logs (i.e. Natural CWD and Natural Control treatments)

177

and the cut logs (i.e., Added CWD and Added Control treatments). Therefore, I compared Natural CWD vs. Natural Control and Added CWD vs. Added Control separately.

Invertebrate data in September 2006 and May and September 2007 were also analyzed using Non-metric Multidimensional Scaling (NMDS) procedure to look for distinct communities among the treatments (PC-ORD, Version 5, McCune and Mefford

2006). I also used a Multi-Response Permutation Procedure (MRPP) to determine if communities were significantly different among treatments. I used Indicator Species

Analysis on each date to determine if taxa were associated with each treatment (PC-

ORD, Version 5, McCune and Mefford 2006).

Results

Soil environmental conditions

The soil conditions were sometimes different between sites where Added CWD and Natural CWD were located, but they were never different between any CWD area and the corresponding Control areas. Temperatures ranged from 9 oC to 16 oC at all sites, and they were warmest in mid-summer. Temperatures were not different between CWD or Control treatments in either location (2006 Added CWD vs. Added

Control ANOVA: F1,18 = 0.042, p = 0.841; 2007 Added CWD vs. Added Control RM-

178

ANOVA: F1,18 = 0.536, p = 0.474; 2006 Natural CWD vs. Natural Control t- Test: t1,8 =

1.846, p = 0.102; 2007 Natural CWD vs. Natural Control RM-ANOVA: F1,10 = 0.394p =

0.538 )

Percent soil moisture varied by site and by date (Figure 1). In September 2006,

Added CWD sites had higher percent soil moisture than Natural Control sites. The

2007, the Repeated Measures ANOVA had a significant Date X Treatment interaction.

The data was analyzed by date and the results showed that there were no differences in

May 2007. In September 2007, Added CWD sites and Added Control were similar to each other but were higher than Natural CWD sites and Natural Control sites (Figure 1).

In September 2007, percent organic matter data was higher in Natural CWD and Natural

Control sites than in the Added CWD and Added Control sites (Figure 1).

Plant Community Responses

Dominant plants (those greater than 5% of any sample) included Leersia oryzoides, Phalaris arundinacea, Sparganium americanum, Polygonum arifolium and

Sagittaria latifolia. Most plant densities were similar in all sites, but there was less invasive reed canary grass (Phalaris arundinacea) and more native rice cutgrass (Leersia oryziodes) in floodplain near the Natural CWD and Natural Control treatments.

However, plants covered the entire 1 m X 1 m area of all sites including the logs. As a result, total plant percent cover was nearly 100% in all sites.

179

Figure 14. Soil environmental conditions below logs and in control areas at Mud Brook

Preserve in 2006 and 2007. A. Percent Moisture (Mean ± 1 SE). B. Percent Organic

Matter (Mean ± 1 SE). Bars with different letters are significantly different on that date.

180

a a, b NS 70 a, b b a a,b 60 c b,c 50 40 30 20

A. Percent Soil Moisture Soil Percent A. 10 0 Sept 2006 May 2007 Sept 2007

45 40 b b 35 Add CWD a a 30 Add Control 25 Natural CWD 20 15 Natural Control 10

B. Percent Organic Matter Organic Percent B. 5 0 2007

181

Invertebrate Community Responses

Invertebrate community characteristics among treatment were usually different between locations but similar between the CWD treatments and their corresponding

Control areas. Species richness in pitfall traps was usually higher at Natural sites (both

CWD and Control) than either the Added CWD or Added Control treatments (Figure 2).

Richness was never different between the Natural CWD and Natural Control or the

Added CWD and Added Control. Total abundance varied throughout the course of the study, and no clear patterns emerged across dates. Total numbers were higher in the

Added CWD treatment than all other treatments in September 2006. In contrast, total numbers were lowest in Added CWD in May 2007 and not different among treatments in September 2007 (Figure 2).

Multivariate analyses using NMDS and MRPP analysis found that invertebrate communities were different among treatments on all sampling dates. However, patterns were complex and changed among dates (Figure 3; Table 1). In September

2006, the treatments formed four distinct clusters on the NMDS plot. The greatest separation was between the communities in Added CWD and Control CWD treatments versus the Natural CWD and Natural Control treatments. MRPP found all pairwise comparisons were significant except Natural CWD and Natural Control were not different. In May 2007, the patterns on the NMDS plot were less clear, but the Added

CWD and Added Control treatments still clustered separately from Natural CWD and

182

Figure 15. Invertebrate communities collected in pitfall traps at Mud Brook Preserve. A.

Species richness per trap (Mean ± 1 SE). B. Total abundance per trap (Mean ± 1 SE) in

September 2006 and May and September 2007. Bars with different letters are significantly different on that date.

183

25 Added CWD b Added Control

20 a,b Natural CWD a a,b b Natural Control 15 b a a,b a a a,b

10 a A. Species Richness SpeciesA. 5

0 Sept 2006 May 2007 Sept 2007

450 400 b

350

300 a b 250 200 b Abundance b

B. B. 150 100 b b a N.S. 50 0 Sept 2006 May 2007 Sept 2007

184

Figure 16. NMDS analysis of invertebrate communities in September 2006, May 2007, and September 2007. The variance explained by each axis is in parentheses.

185

186

Table 13. Comparisons of invertebrate community at the four habitat types at HAERF and Mud Brook, Ohio. Numbers are p values of MRPP pair-wise comparisons among treatments and the overall MRPP statistic. The numbers in the pair-wise comparisons correspond to treatments: 1= Added CWD; 2 = Added Control; 3 = Natural CWD; 4 =

Natural Control.

187

Date 1 vs. 2 1 vs. 3 1 vs. 4 2 vs. 3 2 vs. 4 3 vs. 4 Overall p

Sept 2006 0.004 0.002 0.001 0.004 0.004 0.357 <0.001

May 2007 0.311 0.042 0.018 0.511 0.239 0.262 0.063

Sept 2007 0.244 0.543 0.044 0.036 0.110 0.022 0.035

188

Natural Control treatments. The MRPP analysis found the Added CWD was different than both Natural CWD and Natural Control treatments. In September 2007, patterns on the NMDs were less clear than on the other dates. MRPP analysis found Added

CWD was different than Natural Control, Added Control was different than Natural

CWD, and Natural CWD was different than Natural Control. Therefore, the only pairwise comparison that was different on all sampling dates was between Added CWD and

Natural Control.

Indicator species found that several invertebrate taxa were associated with different treatments. In September 2006, taxa associated with both CWD treatments were either predators (lycosid spiders, carabid beetles) or detritivores (orobatid mites, ptiliid beetles, collembolans). Added Controls had herbivores (hemipterans, and meloid and chrysomelid beetles) as indicator taxa. However, this pattern changed on later dates.

In May 2007, the indicator taxa in CWD treatments were again detritivores and predators, but these trophic groups were also indicators in Control treatments. In

September 2007, indicator taxa in CWD treatments were herbivores, and indicators in

Control treatments were herbivores and predators. Furthermore, patterns of most taxa were variable across dates (Table 2). For example, lycosid spiders were associated with

Natural CWD in Sept 2006 but with Natural Control in May and September 2007. Also, orbatid mites were associated with Added CWD in September 2006, with Natural CWD in May 2007, and they were not indicator taxa in September 2007.

189

‘

“

Table 14. Indicator Species Analysis for the four habitat types at HAERF and Mud Brook in 2006 and 2007. Indicator Value (IV%) has a maximum value of 100% when all samples of a habitat type contain that species

190

Taxa Trophic group Habitat Type IV% p value

September 2006

Arachnida

Lycosidae Predator Natural CWD 15.3 0.014

Orbatidae Detritivore Added CWD 32.5 0.046

Coleoptera

Ptiliidae Detritivore Added CWD 23.6 0.003

Meloidae Herbivore Added Control 28.7 <0.001

Carabidae Predator Added CWD 21.9 0.006

Chrysomelidae Herbivore Added Control 29.4 <0.001

Collembola

Sminthuridae Detritivore Added CWD 46.9 0.001

Entomobryidae Detritivore Added CWD 34.1 <0.001

May 2007

Oligochaetes Detritivore Added CWD 13.3 0.020

Arachnida

Linyphiidae Predator Natural CWD 13.8 0.008

Lycosidae Predator Natural Control 30.4 0.002

Orbatidae Detritivore Natural Control 33.5 0.025

191

Collembola

Entomobryidae Detritivore Natural Control 32.9 0.049

September 2007

Arachnida

Lycosidae Predator Natural Control 15.9 <0.001

Diptera

Sciomyzidae Predator Natural Control 12.1 0.010

Hymenoptera

Chalcidoidea Herbivore Natural Control 12.3 0.010

Formicidae Predator Natural Control 13.3 <0.001

Orthoptera

Gryllidae Herbivore Natural CWD 21.8 0.001

192

Discussion

The impact of CWD on environmental conditions and biotic communities has been poorly studied in floodplains, although some have found it affects invertebrate communities (Braccia and Batzer 2001, Braccia and Batzer 2008). Because it is an important structural component in other aquatic (Naiman and Decamps1997) and terrestrial habitats (Spies et al. 1988), I expected to find clear impacts on soil environmental conditions and invertebrate communities at Mud Brook. In contrast, I found species assemblages rarely differed in areas with and without CWD, and patterns were not consistent across sampling dates. Furthermore, location within the floodplain had a stronger impact than presence of CWD, and therefore abiotic factors that were not strongly influenced by the presence of fallen logs were probably more important in this floodplain.

Abiotic changes caused by CWD can affect invertebrate community structure

(Dechene and Buddle 2010). I predicted that microhabitats with CWD in the floodplain would have less stressful microhabitat condition (i.e. lower soil temperature, higher soil moisture and higher organic matter) than control sites. Soil temperature affects colonization and site selection of vertebrates (Pittman and Dorcas 2009, Anderssen et al. 2010) and invertebrates (Williams et al. 2007). However, this hypothesis was not supported by the data because I found no differences in soil temperature between sites

193

with and without CWD. Other factors I tested (soil organic matter, and soil moisture) differed between locations but not between CWD and Control sites. This suggests that the presence of CWD did not affect on soil microhabitat conditions. The Added CWD treatment may have had little effect because the logs were present in the floodplain for a short time. However, environmental conditions between Natural CWD and Natural

Control treatments were also not different. Presumably, these logs were in the floodplain for many years, so it was surprising that no differences were detected. Other studies have shown that it can take several years for logs to affect microhabitat abiotic conditions (Hansen et al. 1991). There are several possible reasons for the lack of any strong effect of CWD. First, logs might be frequently moved within the floodplain by floods, and they did not remain in a location long enough to influence soil conditions.

Second, I tested Class II logs, which were still relatively sound. Possibly, logs in later decay stages have a stronger effect because the more decayed wood retains more water and nutrients. Third, I had nearly complete plant cover in all sites. Shading by plants and detritus will moderate soil environmental conditions (Facelli and Pickett

1991). Therefore, the vegetation may have had a stronger effect than CWD on the conditions I monitored. Further studies should examine if CWD has a more pronounced effect in unvegetated areas of the floodplain.

The second hypothesis was that areas with CWD would support different invertebrate communities than areas without CWD. This could have occurred if microhabitat environmental conditions, physical structure, or colonization cues differed

194

between areas with CWD and Control areas. On September 2006 and September 2007, the multivariate NMDS analysis found CWD treatments were different than their respective Control treatments. However, I did not find pronounced changes in community structure such as abundance or species richness. Furthermore, the NMDS analysis did not show that communities in CWD and the Control treatments were consistently different. Therefore, the presence of logs did not greatly alter the invertebrate community structure. This may be due to the same reasons that environmental conditions did not differ between treatments (see above).

However, indicator species analysis showed that abundance and occurrence of some species differed between CWD and Control areas. For example, ptiliid beetles

(Coleoptera) were strongly associated with Added CWD sites. These beetles are known inhabitants of CWD, using it as both habitat and a food resource (Arnett and Thomas

2000). In September 2006, springtails (Collembola) and orobatid mites (Acari) were also associated with CWD. Many of these species are detritivores that feed on fungae that could be growing on the decaying CWD (Arnett and Thomas 2000). Predators such as lycosid spiders (Aranea) and carabid beetles (Coleoptera) were associated with CWD in September 2006, perhaps due to higher numbers of their prey. However, the

Indicator Species Analysis also revealed that associations between the treatments and taxa or trophic groups were not consistent between sampling dates. For example, lycosid spiders were later associated with Control treatments. Also, indicator species that were detritivores, herbivores and predators were found in all treatments.

195

Temporal variation in the invertebrate community in the floodplain may occur if flood events relocate invertebrates. Others have found that floods “reset” the spatial distribution of invertebrate communities in floodplains (Naiman and Decamps 1997,

Tockner et al 1999). I did not collect water levels throughout the study and thus cannot directly test this idea. However, data from a USGS gauging station (USGS 04206000) downstream at the Cuyahoga River show that Mud Brook probably experienced frequent flood events. The frequent flooding may have obscured any impacts of the

CWD on invertebrates at Mud Brook.

My third hypothesis predicted that CWD would have a greater impact than site location on invertebrate communities. Instead, I found clear differences between locations at Mud Brook. For example, the MRPP analysis detected different invertebrate communities at the Natural Control and the Added Control sites. Species richness was also greater in Natural CWD and Natural Control sites than Added CWD and Added Control sites. Others have described spatial differences within floodplains.

For example, habitat heterogeneity caused by flood events plays an important role in biodiversity through microhabitat formation and maintenance (Tockner et al. 2000,

Ward et al. 1999).

A potential reason for the spatial differences at Mud Brook was that environmental conditions were different between locations. In September 2007, soil moisture was lower and organic matter content was higher in the location with Natural

CWD and Natural Control treatments, which might have affected invertebrate

196

biodiversity. For example, many invertebrates are not tolerant of living in anoxic soils.

(Schmitz and Harrison 2004). Therefore, biodiversity may have increased because soils that are not water-logged are oxidized and can provide a better refuge for soil dwelling invertebrates (Schmitz and Harrison 2004). Likewise, soils with higher organic content may provide abundant detritus for detritivores and plant biomass for herbivores

(Brussaard et al. 1997, Wall et al. 1999). Furthermore, I observed plant communities were also different between locations. The location with Natural CWD and Natural

Control treatments was dominated by Leersia oryzoides, Sparganium americanum, and

Polygonum arifolium, which are native wetland species (Reed 1988, Tiner 1991, USDA

Plants). Whereas the Control CWD and Added CWD treatment sites were dominated by

Phalaris arundinacea, an invasive wetland species, and Sagittaria latifolia, a native species. Phalaris arundinacea is a problem in wetlands because it forms dense monocultures, which degrades wildlife habitat and decreases invertebrate biodiversity

(Spyreas et al. 2010).

Perhaps one of the most intriguing results I detected was the high temporal variability in patterns of invertebrate distribution. As discussed above, I found differences between CWD and Control treatments and between Natural and Added locations on some dates but not others. My data did not support the idea that this occurred because it took time for differences to develop (i.e., if it took several months for invertebrates to colonize the Added CWD treatments) because some differences appeared on early dates but disappeared later. There are several potential reasons for

197

the high amount of temporal variability that I observed. In September 2005 before the onset of this study, a storm from Hurricane Katrina caused a 100 year flood event in

Mud Brook. Water levels at MBP were > 1 m above normal and most CWD was relocated within the floodplain (M. Drinkard, pers. observ.). Large storms can alter impact biotic communities for years afterwards (Parsons et al. 2005), and the communities in MBP may still have been reestablishing an equilibrium in 2006 and 2007.

Another explanation for the variability in the results could be the scale of the experiments. Impacts of CWD can be variable on a small scale (individual logs), whereas the presence or absence of CWD might have predictable impacts across the entire floodplain (Ulyshen and Hanula 2009). Bowman et al. (1999) found such results while comparing effects of CWD on small mammals at both the log and landscape scale.

McGill (2010) posited that many environmental factors that researchers report are unimportant, may have been studied at an inappropriate scale. For example, if I tested

CWD presence across the entire floodplain there might still be noticeable impacts in spite of habitat homogenization by floods. Hence, future studies should look at impacts of CWD on a habitat-wide level to fully test for their impact on floodplain invertebrate communities.

198

Acknowledgements

I would like to thank Brendan Morgan, Jennifer Clark, Doug Kapusinski, Justin

Montemarano, Leonard Neil Drinkard 2 & 3 for their help in the field and lab. This grant was support by the Art and Margaret Herrick Endowment fund for the Herrick Aquatic

Ecology Research Facility.

CHAPTER 6

SYNTHESIS AND DISCUSSION

Most of the studies that developed and tested the Flood Pulse Concept (FPC) were conducted in large-order riparian systems such as the Amazon River in Brazil (Junk et al. 1989), Danube River in Europe (Tockner et al. 1999a) and medium-order rivers like those found in the southeastern USA coastal plains (Benke 2001). In these systems, the timing and relative magnitude of river flooding is predictable and results in higher plant and animal diversity, abundance, and biomass in the river and floodplain ecosystems (Junk 1989, Benke 2001, Bunn and Arthington 2002, Malmqvist 2002, Ward et al. 2002). These studies found that flood-pulsing increased community diversity and productivity because it created beneficial impacts such as resource subsidy and habitat heterogeneity (Polis et al. 1997, Pollack et al. 1998, Ward et al. 1999, Robinson et al.

199

200

2002). These studies also found that flood-pulsing caused some stressful impacts such as erosion, sedimentation, submersion, and soil anoxia (Vargo et al. 1998, Quinn et al.

2000, Trondstad et al. 2005). Furthermore, habitat features commonly found in unrestricted floodplains, such as coarse woody debris from riparian trees, moderated the effects of some stressors (Alsfeld et al. 2009). Therefore, the relative importance of the beneficial versus the stressful impacts varied among river systems and among habitat types within the floodplain. However, the current paradigm is that flood- pulsing in large rivers creates a net positive impact because biotic communities obtain more nutrients and potential niches and riparian taxa are adapted to survive the predictable occurrence of stressful conditions (Blom 1999, Middleton 2002, Desilets and

Houle 2005, Jackson and Colmar 2005, Batzer et al. 2006).

Although there are many studies about the FPC in large rivers, the impact of flood pulsing on low-order systems has not been well explored. My studies tested the effects of flooding in headwater ecosystems in both mesocosm and natural wetlands.

My research is of particular importance because these systems make a significant impact on downstream habitats (Tockner et al. 2000b, Gomi et al. 2001, Freeman et al.

2007). I hypothesized that flood-pulsing would drive community composition in headwater creeks just as in large-order systems, but I also expected that the short, stochastic flood regimes would alter the relative importance of positive and negative impacts on the biota. For example, long floods in large rivers may reduce riparian plant productivity when they cause extended periods of soil anoxia (Molles et al. 1998), but

201

short floods in headwaters could create less stressful soil conditions. However, productivity in headwaters might not be affected if the short floods cause little to no resource subsidy. Therefore, my dissertation examined the positive and negative impacts of flood-pulsing caused by the hydrological regime in headwater ecosystems.

My dissertation research tested impacts of flood-pulsing in medium-scale man made basins (i.e., mesocosms) at the Herrick Aquatic Ecology Research Facility (HAERF) and natural wetlands at Mud Brook Preserve (MBP). In this chapter, I will discuss the positive and negative impacts caused by this hydrology, and I will show that that several aspects of the FPC in large rivers may not apply to headwater systems. Furthermore, I will also discuss the limitations and benefits of using a mesocosm approach to test the impact of an ecosystem-level process such as flood-pulsing.

Impacts of flood-pulsing on invertebrates and plants in headwater wetlands

Positive impacts of flood pulsing

I expected that flood-pulsing would benefit species that had adaptations to survive the frequent, unpredictable and short-term inundation in headwater wetlands.

As expected, several plant species had higher abundances in the intermittently flooded

202

zone of flood pulse mesocosms at HAERF. Plant taxa that were more abundant in the

Flood Pulse wetlands were usually Obligate wetland species (OBL). These included a broad taxonomic range of species including sedges (e.g., Carex lurida, Cyperus strigosus,

Scirpus cyperinus), rushes (e.g., Juncus effusus), and forbs (e.g., Polygonum hydropiperoides). Species classified as OBL taxa are found in wetlands <99% of the time and are well-adapted surviving extended periods of submergence or saturated soils.

They are rarely found outside of wetlands because they don’t tolerate desiccation, or they are outcompeted by upland species (Tiner 2006). Battaglia and Collins (2006) also found tight linkage between hydroperiod and plant communities and also found that

OBL taxa abundance was highest in areas with fluctuating hydroregime.

In this study, I observed that there were distinct communities in the elevational zones of the mesocosm banks. For example, OBL plant cover was highest in the low zone and less flood-tolerant species were increasingly abundant as elevation increased.

Plant zonation along elevation gradients is common in wetland systems, and it often caused by changes in hydrology and soil conditions (Sanderson et al. 2008). However, I also found that OBL plant cover was more abundant in the mid and high zones of Flood

Pulse wetlands than Static wetlands. Thus, flood pulses created environmental conditions that benefited OBL taxa at higher elevations in Flood Pulse wetlands .

Because OBL taxa are only found growing in wetlands, this indicates that the areal extent of wetland habitat should be greater in Flood Pulse wetlands than static wetlands that receive equal water inputs.

203

These results have some important implications when developing plans to restore drained or impounded floodplain wetlands. First, I showed that flood-pulsing allowed the wetland area to extend to higher elevations. The treatment also increased plant species that had desirable ecological attributes such as native and non-weedy species. Wetland managers often try to increase these taxa because they provide better habitat conditions for wildlife (Marzluff and Ewing 2001). However, many formerly drained riparian wetlands have been restored by installing water control structures in the dike and inundating the floodplain manually, which creates static water levels during most of the year (France 2003). Instead, creating a permanent opening in dikes to allow a natural flood-pulsing hydrology may be a better strategy to enhance habitat quality in restored riparian wetlands.

In large rivers, long term floods import nutrients into the floodplain and subsequently, enhance riparian plant productivity (Bayley and Guimond 2009). In my mesocosm studies, the Low zone was flooded ~20% of the time in Flood pulse mesocosms, but only briefly (<5% of the time) in Static Wetlands. Thus, the strongest impacts of flood-pulsing on plant growth should have been evident in this zone.

Although biomass was greatest in low elevation zone, it did not differ between Flood

Pulse wetlands and Static wetlands. This suggests that proximity to the permanent pools had a greater impact on plant growth than the short, stochastic floods, and therefore there may be minimal resource subsidies in these systems. An alternate conclusion is that the positive (e.g., resource subsidy) and negative (unpredictable soil

204

anoxia) effects may have cancelled each other out. I did not directly measure soil abiotic conditions, and therefore cannot speculate which of these scenario is correct.

Unlike plant community responses, I did not detect any strong positive impact of flood-pulsing on aquatic invertebrates in HAERF. Some trophic groups (shredders and scrapers) were more abundant in Flood pulse mesocosms, but this only occurred on a single date. There was no difference in emerging insect numbers between Flood pulse and Static wetlands. I had expected that flood-pulsing would provide a resource subsidy for invertebrates that accessed detritus and other food items on the floodplain.

In other studies, invertebrate communities showed a positive response to resource subsidies, but these researchers studied large river floodplains (Freeman et al. 2007,

Lake et al. 2007). However, the narrow banks at HAERF were only accessible during short floods, and therefore there was probably little impact on invertebrate food resources. Furthermore, all mesocosms at HAERF were permanently flooded, which minimized differences to aquatic invertebrate habitat conditions. Complete draw downs are common in flood-pulsing riparian wetlands, and this will restructure aquatic invertebrate communities (Silver et al. 2012). Draw downs oxygenate soil nutrients and make them biologically available to plants that provide food for invertebrates (Olila et al. 1997, Pant and Reddy 2001, Strack et al. 2008). Dewatering also eliminates desiccation intolerant invertebrates (Blom and Voesenek 1996, Blom 1999). If I had designed my experiment at HAERF to compare invertebrate communities in static mesocosms that were permanently flooded and flood pulsing mesocosms that were

205

intermittently dewatered, I probably would have found major differences. For example, I found clear differences at Mud Brook Preserve between the permanently flooded pools and the intermittently flooded wetlands. However, my study was not designed to test the impacts of dewatering as an environmental factor.

Logs that are washed into floodplain wetlands create a physically complex habitat and reduce environmental stresses (e.g. dry, hot soils) (Naiman and Decamps

1997, Spieles and Horn 2009). At Mud Brook Preserve, I observed many fallen trees had been carried in by floods. Therefore, I expected that adding coarse woody debris

(CWD) would increase invertebrate biodiversity because many taxa inhabit moist soils under decaying logs. However, when I added CWD, there was no change in invertebrate diversity. There are several potential explanations why my treatment did not show a beneficial impact. First, perhaps the frequent floods from the headwater creek eliminated any impact of CWD on soil moisture. Second, the CWD was only in place for a few months, and it can take years before an impact is detected (Hansen et al.

1991). Furthermore, I added Decay Class II logs, which are relatively intact (Pyle and

Brown 1998). Later decay stages range from pronounced exterior decay (Class III) to interior decay (Class IV) to complete decay (Class V). If I had used more decayed CWD, I might have produced more noticeable results (Harmon et al. 1986). Finally, some have suggested that the effects of CWD are most predictable at the landscape scale. For example, animal and plant diversity in small patches with different amounts of CWD varies widely, but diversity increases with CWD presence across the entire floodplain

206

(Gurnell et al. 1995, Bowman et al. 2000, Alsfeld et al. 2009, Ulyshen and Hanula

2009). Therefore, it may be better to design an experiment testing the landscape-level impact of CWD in headwater floodplains before discounting its importance.

Negative impacts of flood-pulsing

I hypothesized that flood-pulsing would increase plant and invertebrate diversity in the flood pulse mesocosms. Unexpectedly, I found no impact on aquatic invertebrates

(see above) and plant diversity in the intermittently flooded zone of the mesocosms was lower in the flood pulse than the static wetlands. Although the difference was only statistically significant in 2005, the same pattern was evident in 2003 and 2004.

Therefore, more plant taxa were eliminated by flood-pulsing than benefited from the treatment. As discussed above, many OBL taxa benefited from flood-pulsing.

However, Upland (UPL) and Facultative Upland (FACU), annual, dicotyledons, exotic and weedy species were lower in Flood Pulse wetlands showing that the negative impacts suppressed many species.

Another negative outcome of flood pulsing was the increased amount of bare ground in the Flood Pulsing mesocosms. This response occurred in the study of overall impacts on plant communities (Chapter 2) and the examination of plant zonation

(Chapter 3). For example, bare ground was higher in all elevational zones of the Flood pulse mesocosms than the Static mesocosms. Bare ground is a known sign of stressful

207

environmental conditions (Vartapetian and Jackson 1997), and the flood-pulsing probably increased seedling survival in the newly created wetlands at HAERF.

My zonation study at HAERF measured the length of time required for flooding to negatively impact plants. For example, there was more bare ground in the Mid zone of

Flood Pulse wetlands, and that zone was submerged only ~20% of the time in the Flood pulse treatment. Interestingly, there was more bare ground in High zone of Flood Pulse wetlands, even though this zone was only flooded eight times in the 2005-2006 study.

This finding supports the idea that even short floods can be an important factor in headwater systems.

Although there was more bare ground in flood pulse mesocosms, biomass did not differ between treatments (see above). This indicates that although there were fewer plants in the Flood Pulse wetlands, the plants were larger. Similar trade-offs between stand density and plant vigor have been described in other studies. For example, water hyacinth (Eichhornia crassipes) biomass was the same in impounded lakes and lakes open to floodwaters, but densities were lower and individual plants were larger in impounded lakes (Neiff et al. 2008). In large river ecosystems, riparian plant growth increases when floods bring in large amounts of nutrients from upstream ecosystems (Tockner et al. 1999a; Mettler et al. 2001). However, the lack of a difference in total plant biomass at HAERF suggests that the short floods did not input a significant amount of nutrients. A more likely explanation is that the few plants that survived the stressful conditions in the flood pulsing wetlands, grew more vigorously

208

because they experienced less competition for nutrients, water, or light. It is well established that thinning terrestrial plant stands (e.g., forests) increases vigor of the remaining individuals (Mitchell et al. 1983, Rodriguez et al. 2003). There are few studies of this in headwater systems, but it is possible that plants in Flood Pulse wetlands respond in the same way.

Although plant diversity and cover at HAERF was lower in response to flood pulsing, it is not clear what were the important stressors in these wetlands. Several types of stressful conditions are caused during floods, and these will eliminate plants if they occur frequently (Blom 1999, Gurnell et al. 2007). For example, silt deposits increase seedling mortality (Jurik et al. 1994, Ewing 1996, Gleason et al. 2003), and soil anoxia reduces root cell metabolism and causes a buildup of soil toxins (Justin and Armstrong

1987, Vartapetian and Jackson 1997). The low zone in HAERF mesocosms was inundated over 152 times in 2005 – 2006, and there was probably a short but frequent occurrence of stressful conditions. I did not measure soil redox potential, but I often observed silt deposits on plant leaves after floods subsided. Further research is needed to understand which specific stressors were caused by short, unpredictable floods in headwater wetlands.

Using mesocosms to test large-scale ecological processes

209

The mesocosm approach has been widely used, and it has both limitations and benefits (Carpenter 1996, Drenner and Mazmuder 1999, Ahn and Mitsch 2002, McGill

2010, Spivak et al. 2011). An important advantage of using this approach is the relative ease of creating similar experimental replicates, which helps eliminate potential confounding variables that affect the response to the treatment. At HAERF, I tested the impact of hydrology on plants and invertebrates in 10 relatively similar wetlands.

The mesocosms at HAERF were equal in surface area, depth, age, soils, water source, and the timing and frequency of flooding. Therefore the mesocosm approach allowed me to control for many variables while precisely manipulating hydrology, which would not have been possible if my replicates were natural flood-pulsing and static wetlands.

Another advantage of using the earthen mesocosms at HAERF was that they simulated habitat conditions found in natural wetlands. First, the mesocosms were sized about the same as small natural pools I observed in the Mud Brook floodplain.

HAERF also supported a diverse plant and animal community, and many taxa at HAERF occur at Mud Brook and other natural wetlands (Thompson et al. 2007). Most importantly, the water control structures allowed large storms to flood the mesocosms.

Therefore, the hydrology I tested was realistic, and I found very similar flooding regimes in HAERF and the Mud Brook floodplain.

However, my research also pointed out some of the limitations of using a mesocosm approach to test large-scale ecological processes. My results suggest that the mesocosms did not fully simulate all habitat conditions in natural wetlands. For

210

example, the plant community differed somewhat between the mesocosms and natural wetlands. For example, the four most abundant plant species at HAERF in 2006 were

Daucus carota, Solidago Canadensis, Poa palustris, and Symphyotrichum lanceolata but at Mud Brook they were Leersia oryzoides, Phalaris arundinacea, Sparganium americanum, and Polygonum arifolium. Aquatic invertebrate communities also differed between mesocosms at HEARF and the natural wetlands at Mud Brook. My multivariate analyses showed that the intermittently flooded zone and permanently flooded pool at Mud Brook were both different from either treatment at HAERF.

A key difference between HAERF and natural ecosystems is the scale of the habitat. Although the 10 m X 20 m mesocosms could not be considered “small-scale”, they could not reproduce all of the physical and biological complexity of a natural floodplain wetland. Previous studies have found that the amount of physical complexity is determined by mesocosm size, and this will determine biodiversity (Spivak et al. 2011). For example, smaller habitats tend to be less complex, and therefore, more prone to impacts of stressors and have lower diversity than larger habitats

(Therriault and Kolasa 2000). Not surprisingly, the HAERF mesocosms had lower plant and invertebrate biodiversity than described in natural floodplain wetlands (Tockner and Stanford 2002). Therefore, some biotic interactions were not fully tested with my experimental design.

Mesocosms also cannot recreate several important habitat features of natural floodplain wetlands such as habitat age, presence of overland flow, and the pool to

211

floodplain ratio. The mesocosms at HAERF were constructed one year before I initiated my dissertation research, and I only ran the flood pulse experiment for 2 years.

Seabloom et al. (2001) reported that it took at least 2 years for plant communities to change in response to a hydrology alteration. Therefore, communities at HAERF may still have been changing at the conclusion of my study. A second factor was that my mesocosms lacked overland flow because they were flooded through the water control structures. As rivers over flow their banks, the water picks up substantial amounts of nutrients from the floodplain that are deposited into riparian wetlands (Baxter et al.

2005, Alsfeld et al. 2009). This resource subsidy from floodplains to wetland pools strongly influences the environmental conditions, productivity and diversity (Polis et al.

1997). A related factor was that each mesocosm was surrounded by relatively a narrow intermittently flooded zone. The amount of exchange that could occur between the intermittently flooded zone and the pools during floods was proportional to the ratio of these habitats. I estimated that the ratio of floodplain area : pool area was approximately 1:1 at HAERF, but the wetland pool I studied at Mud Brook Preserve was embedded within a 40 hectare floodplain. Therefore, there was less possibility that resource subsidies could affect aquatic invertebrate or submerged macrophyte communities at HAERF. These design issues are all difficult to overcome, and some important features of natural systems may be impossible to reproduce with a mesocosm approach.

212

Overall Conclusions

In summary, I found that flood-pulsing in headwater wetlands causes a “mixed bag” of positive and negative impact to plant and invertebrate communities. Although there were beneficial impacts to a few plant species (i.e. OBL taxa), the overall plant community showed a net negative impact (lower biodiversity, more bare ground).

Furthermore, the treatment had a limited impact on the aquatic invertebrate community. These results contrast with the current paradigm that flood pulsing causes an overall positive impact on plant and animal communities in large river floodplains.

My dissertation research suggests that some ecological principles developed in large- order systems with predictable and long-term floods are not applicable to small-order floodplain systems with short unpredictable flood regimes. A fruitful area for future research will be to test if the results I found in the mesocosms at HAERF can be widely applied to natural headwater systems.

LITERATURE CITED

Abbe, T.B., and D.R. Montgomery. 1996. Large woody debris jams, channel hydraulics

and habitat formation in large rivers. Regulated Rivers Research and

Management 12: 201-221.

Adis, J. and W. J. Junk. 2002. Terrestrial invertebrates inhabiting lowland river

floodplains of central Amazonia and central Europe: A review. Freshwater

Biology 7: 711-731.

Ahn, C. and W.J. Mitsch. 2002. Scaling considerations of mesocosms wetlands in

simulating large created freshwater marshes. Ecological Engineering 18: 327-

342.

Alsfeld, A.J., J.L. Bowman, and A. Deller-Jacobs. 2009. Effects of woody debris,

microtopography, and organic matter amendments on the biotic community of

constructed depressional wetlands. Biological Conservation 142: 247-255.

213

214

Amoros, C. and G. Bornette. 2002. Connectivity and biocomplexity in waterbodies of

riverine floodplains. Freshwater Biology 47 (4): 761-776.

Anderssen, M., A. Krockenberger, and L. Schwarzkopf. 2010. Experimental

manipulation reveals the importance of refuge habitat temperature selected by

lizards. Australian Ecology 35: 294-299.

Andreas, B.K., J.J. Mack, and J.S. McCormack. 2004. Floristic quality assessment and

index (FQAI) for vascular plants and mosses for the State of Ohio. Ohio

Environmental Protection Agency, Division of Surface Water, Wetland Ecology

Group, Columbus, Ohio.

Arar, E.J. and G.B. Collins. 1997. Method 445.0, In Vitro Determination of Chlorophyll

a and Pheophytin a in Marine and Freshwater Algae by Fluorescence. In

Methods for the Determination of Chemical Substances in Marine and Estuarine

Environmental Matrices, 2nd Edition. National Exposure Research Laboratory,

Office of research and development, USEPA., Cincinnati, Ohio (EPA/600/R-

97/072, Sept. 1997).

Arnett, R.H. and M. C. Thomas. 2000. American Beetles, Volume 1: Archostemata,

Myxophaga, Adephaga, Polyphaga: Staphyliniformia. CRC Press, Boca Raton,

Florida, USA.

215

Barry, M.J., B.K. Andreas, and F.A. de Szalay. 2008. Long-term plant community

changes in managed fens in Ohio, USA. Aquatic Conservation 18: 392-407.

Barry, M.J., R. Bowers, and F.A. de Szalay. 2004. Effects of hydrology, herbivory and

sediment disturbance on plant recruitment in a Lake Erie coastal wetland.

American Midland Naturalist 151: 217-232.

Battaglia, L.L., and B.S. Collins. 2006. Linking hydroperiod and vegetation response in

Carolina bay wetlands. Plant Ecology 184: 173-185.

Batzer, D.P., R.B. Rader, and S. A. Wissinger. 1999. Invertebrates in freshwater

wetlands of North America: Ecology and management. New York: Wiley. 1100

Batzer, D. P., and S. A. Wissinger. 1996. Ecology of insect communities in nontidal

wetlands. Annual Review of Entomology 41: 75-100.

Batzer, D. P., R. Cooper, and S. A. Wissinger. 2006. Wetland animal ecology. p. 242 -

284. In D. P. Batzer and R. R. Sharitz (eds.) Ecology of Freshwater and

Estuarine Wetlands. University of California Press, Berkeley, CA, USA.

Batzer, D.P., S.E. Dietz-Brantley, B.E. Taylor, and A.E. DeBiase. 2005. Evaluating

regional differences in macroinvertebrate communities from forested

depressional wetlands across eastern and central North America. Journal of the

North American Benthological Society 24: 403-414.

216

Baxter, C.V., K.D. Fausch, and W.C. Saunders. 2005. Tangled webs: reciprocal flows of

invertebrate prey link streams and riparian zones. Freshwater Biology 50: 201-

220.

Bayley, P. B. 1995. Understanding Large River: Floodplain Ecosystems. Bioscience 45:

153-158.

Bayley, S.E., and J.K. Guimond. 2009. Aboveground biomass and nutrient limitation in

relation to river connectivity in montane floodplain marshes. Wetlands 29:

1243-1254.

Bazzanti, M., V.D. Bella, and M. Seminara. 2003. Factors affecting macroinvertebrate

communities in astatic ponds in central Italy. Journal of Freshwater Ecology 18:

537-548.

Benigno, G.M., and T.R. Sommer. 2008. Just add water: Sources of chironomid drift in

a large river floodplain. Hydrobiologia 600: 297-305.

Benke, A. C. 2001. Importance of flood regime to invertebrate habitat in an

unregulated river-floodplain ecosystem. Journal of the North American

Benthological Society 20: 225-240.

Benke, A.C., I. Chaubey, G.M. Ward, and E.L. Dunn. 2000. Flood pulse dynamics of an

unregulated river floodplain in the southeastern US coastal plain. Ecology 81:

2730-2741.

217

Blom, C.W.P.M. 1999. Adaptations to flooding stress: From plant community to

molecule. Plant Biology 1: 267-273.

Blom, C.W.P.M., and L.A.C.J. Voesenek. 1996. Flooding: The survival strategies of

plants. Tree 11: 290-295.

Blossey, B., L.C. Skinner, and J. Taylor. 2001. Impact and management of purple

loosestrife (Lythrum salicaria) in North America. Biodiversity and Conservation

10: 1787-1807.

Bornette, G. and C. Amoros. 1996. Disturbance regimes and vegetation dynamics:

Role of floods in riverine wetlands. Journal of Vegetation Science 7: 615-22.

Boudell, J.A. and J. Stromberg. 2008. Flood pulsing and metacommunity dynamics in a

desert riparian ecosystem. Journal of Vegetation Science 19: 373-380.

Boulton, A. J., and L.N. Lloyd. 1992. Flooding frequency and invertebrate emergence

from dry floodplain sediments of the River Murray, Australia. Regulated Rivers

Research and Management 7: 137-151.

Boutin, C., and P. Keddy. 1993. A Functional Classification of Wetland Plants. Journal

of Vegetation Science 4: 591-600.

Bowles, M.L., P.D. Kelsey, and J.L. McBride. 2005. Relationships among environmental

factors, vegetations zones, and species richness in a North American calcareous

prairie fen. Wetlands 25: 685-696.

218

Bowman, J.C., D. Sleep, G.J. Forbes, and M. Edwards. 2000. The association of small

mammals with coarse woody debris at log and stand scales. Forest Ecology and

Management 129: 119-124.

Braccia, A. and D.P. Batzer. 2001. Invertebrates associated with woody debris in a

southeastern U.S. forested floodplain wetland. Wetlands 21: 18-31.

Braccia, A. and D.P. Batzer. 2008. Breakdown and invertebrate colonization of dead

wood in wetland, upland, and river habitats. Canadian Journal for Research 38:

2697-2704.

Bragg, D.C. 2000. Simulating catastrophic and individualistic large woody debris

recruitment for a small riparian system. Ecology 81: 1383-94.

Bright, E. G., D. P. Batzer, and J.A. Garnett. 2010. Variation in invertebrate and fish

communities across floodplain ecotones of the Altamaha and Savannah Rivers.

Wetlands 30: 1117-1128.

Brock, M.A., D.L. Nielsen, R.J. Shiel, J.D. Green and J.D. Langley. 2003. Drought and

aquatic community resilience: the role of eggs and seeds in sediments of

temporary wetlands. Freshwater Biology 48: 1207-1218.

Brose, U. 2001. Relative importance of isolation, area and habitat heterogeneity for

vascular plant species richness of temporary wetlands in east-German farmland.

Ecography 24: 722-730.

219

Brussaard, L., V.M. Behan-Pelletier, D.E. Bignell, V.K. Brown, W. Didden, P. Folgarait,

C. Fragoso, D. Wall, V.V.S.R. Gupta, Tsutomu Hattori, D.L. Hawksworth, C.

Klopatek, P. Lavelle, D.W. Malloch, J. Rusek, B. Soderstrom, J.M. Tiedje, and

R.A. Virginia. 1997. Biodiversity and ecosystem functioning in soil. Ambio 26:

563-570.

Bunn, S. E., and A. H. Arthington. 2002. Basic principles and ecological consequences

of altered flow regimes for aquatic biodiversity. Environmental Management 30:

492-507.

Byrd, K.B. and M. Kelly. 2006. Salt marsh vegetation response to edaphic and

topographic changes from upland sedimentation in a pacific estuary. Wetlands

26: 813-829.

Capenter, S.R. 1996. Microcosm experiments have limited relevance for community

and ecosystem ecology. Ecology 77: 677-680.

Casanova, M.T. and M.A. Brock. 2000. How do depth, duration, and frequency of

flooding influence the establishment of wetland plant communities? Plant

Ecology 147: 237-250.

Cavalcanti, G.G. and B.G. Lockaby. 2005. Effects of sediment deposition on fine root

dynamics in riparian forests. Soil Science Society of America Journal 69: 729–

737.

220

Chadde, S. 2002. A Great Lakes Wetland Flora: A Complete Guide to the Aquatic and

Wetland Plants of the Upper Midwest. PocketFlora Press.

Chase, J.M. and M.A. Leibold. 2002. Spatial scale dictates the productivity-biodiversity

relationship. Nature 416: 427-430.

Chase, J.M. 2007. Drought mediates the importance of stochastic community

assembly. Proceedings of the National Academy of Science 104: 17430-17434.

Connell, J.H. 1983. On the prevalence and relative importance of interspecific

competition: evidence from field experiments. American Naturalist 122: 661-

696.

Corti, D., S.L. Kohler, and R.E. Sparks. 1997. Effects of hydroperiod and predation on a

Mississippi river floodplain invertebrate community. Oecologia 109:154-65.

Costanza, R., R. d’Arge, R. de Groot, S. Farber, M. Grasso, B. Hannon, K. Limburg, S.

Naeem, R.V. O’Neill, J. Paruelo, R.G. Raskin, P. Sutton, and M. van den Belt.

1998. Special Section: Forum on valuation of ecosystem services: The value of

ecosystem services: putting the issues in perspective. Ecological Economics 25:

67-72.

Covich, A.P., M.A. Palmer, and T.A. Crowl. 1999. The role of benthic invertebrate

species in freshwater ecosystems. Bioscience 49: 119-127.

221

Craft, C., S. Broome, and C. Campbell. 2002. Fifteen years of vegetation and soil

development after brackish-water marsh creation. Restoration Ecology 10: 248-

258.

Crow, G.E. and C.B. Hellquist. 2000. Aquatic and Wetland Plants of Northeastern

North America. Vol. 1: Pteridophytes, Gymnosperms, and Angiosperms:

Dicotyledons, 480 pp.; Vol. 2: Angiosperms: Monocotyledons, 400 pp.

University of Wisconsin Press, Madison. de Szalay, F.A. and V.H. Resh. 2000. Factors influencing macroinvertebrate

colonization of seasonal wetlands: Responses to emergent plant cover.

Freshwater Biology 45: 295-308. de Szalay, F.A., L. C. Carroll, J. A. Beam, and V. H. Resh. 2003. Temporal overlap of

nesting duck and aquatic invertebrate abundances in the Grasslands Ecological

Area, California. Wetlands 23: 739-749.

Dechene, A.D. and C.M. Buddle. 2010. Decomposing logs increase orbatid mite

assemblage diversity in mixed wood boreal forest. Biodiversity Conservation 19:

237-256.

Delnicki, D. and K.J. Reinecke. 1986. Winter food use and body weights of mallards

and wood ducks in Mississippi. Journal of Wildlife Management 50: 43-51.

222

Desilets, P. and G. Houle. 2005. Effects of resource availability and heterogeneity on

the slope of the species-area curve along a floodplain-upland gradient. Journal

of Vegetation Science 16: 487-496.

Dietz-Brantley, S.E., B.E. Tylor, D.P. Batzer, and A.E. DeBiase. 2002. Invertebrates that

aestivate in dry basins of Carolina Bay wetlands. Wetlands 22: 767-775.

Dittmar, L.A. and R.K. Neeley. 1999. Wetland seed bank response to sedimentation

varying in loading rate and texture. Wetlands 19: 341-351.

Drenner, R.W. and A. Mazmuder. 1999. Microcosm experiments have limited

relevance for community and ecosystem ecology: Comment. Ecology 80: 1081-

1085.

Drinkard, M.K., M.W. Kershner, A. Romito, J. Neiset, and F.A. de Szalay. 2011.

Responses of plant and invertebrate assemblages to water-level fluctuation in

headwater wetlands. Journal of North American Benthological Society 30: 981-

996.

Dufrene, M. and P. Legendre. 1997. Species assemblages and indicator species: The

need for a flexible asymmetrical approach. Ecological Monographs 67: 345-366.

Dwire, K.A., J.B .Kauffman, E.N.J. Brookshire, and J.E. Baham. 2004. Plant biomass and

species composition along an environmental gradient in montane riparian

meadows. Oceologia 139: 309-317.

223

Dynesius, M. and C. Nilsson. 1994. Fragmentation and flow regulation of river systems

in the northern third of the world. Science 266: 753-62.

Englehart, K.A.M. and M.E. Ritchie. 2002. The effect of aquatic plant species richness

on wetland ecosystem processes. Ecology 83: 2911-2924.

Euliss, N.H. and D.M. Mushet. 1999. Influence of agriculture on aquatic invertebrate

communities of temporary wetlands in the Prairie Pothole Region of North

Dakota, USA. Wetlands 19: 578-583.

Ewing, K. 1996. Tolerance of four wetland plant species to flooding and sediment

deposition. Environmental and Experimental Botany 36: 131-146.

Facelli, J.M. and S.T.A. Pickett. 1991. Plant litter – It’s dymanics and effects on plant

community structure. Botanical Review 57: 1-32.

Fairchild, G.W., J. Cruz, A.M. Faulds, A.E.Z. Short, and J.F. Matta. 2003. Microhabitat

and landscape influences on aquatic beetle assemblages in a cluster of

temporary and permanent ponds. Journal of North American Benthological

Society 22: 224-240.

Finkenbine, J.K, J.W. Atwater, and D.S. Mavinic. 2000. Stream health after

urbanization. Journal of the American Water Resources Association 36: 1149-

1160.

224

France, R.L. 2003. Wetland Design: Principles and Practices for Landscape Architects

and Land-Use Planners. W.W. Norton and Company, Inc. New York, New York.

Fraser, L.H. and J.P. Karnezis. 2005. A comparative assessment of seedling survival and

biomass accumulation for fourteen wetland species grown under minor water

depth differences. Wetlands 25: 520-530.

Fraser, L.H. and T.E. Miletti. 2007. Effect of minor water depth treatments on

competitive effect and response of eight wetland plants. Plant Ecology 195: 33-

43.

Fredrickson, L.H. and T.H. Taylor. 1982. Management of seasonally flooded

impoundments for wildlife. Fish and Wildlife Service Resource Publ. 148. US

Dept. of Interior, Washington D.C.

Freeman, M.C., C.M. Pringle, and C.R. Jackson. 2007. Hydrologic connectivity and the

contribution of stream headwaters to ecological integrity at regional scales.

Journal of the American Water Resources Association 43: 5-14.

Galatowitsch, S.M. and A.G. van der Valk. 1995. Natural revegetation during

restoration of wetlands in the southern prairie pothole region of North America.

Symposium proceedings: Restoration of Temperate Wetlands: 129-142.

Galatowitsch, S.M. and A.G. van der Valk. 1996. The vegetation of restored and

natural prairie wetlands. Ecological Applications 6: 102-112.

225

Gippel, C.J. 1995. Environmental hydraulics of large woody debris in streams and rivers.

Journal of Environmental Engineering 121: 388-395.

Gladden, J.E. and L.A. Smock. 1990. Macroinvertebrate distribution and production on

the floodplains of two lowland headwater streams. Freshwater Biology 24: 533-

545.

Gleason, R. A., N. H. Euliss, D. E. Hubbard, and W. G. Duffy. 2003. Effects of

sediment load on emergence of aquatic invertebrates and plants from wetland

soil egg and seed banks. Wetlands 23: 26-34.

Gomi, T., R.C. Sidle, and J.S. Richrdson. 2002. Understanding processes and

downstream linkages of headwater streams. Bioscience 52: 905-916.

Goodson, J.M., A.M. Gurnell, P.G. Agnold, and I.P. Morrissey. 2001. Riparian seed

banks: Structure, process and implications for management. Progress in Physical

Geography 25: 301-325.

Grove, S.J. 2002. Saproxylic insect ecology and the sustainable management of forests.

Annual Review of Ecology and Systematics 33: 1-23.

Gurnell, A., J. Goodson, K. Thompson, O. Mountford, and N. Clifford. 2007. Three

seedling emergence methods in soil seed bank studies: Implications for

interpretation of propagule deposition in riparian zones. Seed Science Research

17: 183-199.

226

Gurnell, A.M., H. Piegay, F.J. Swanson, and S.V. Gregory. 2002. Large wood and fluvial

processes. Freshwater Biology 47: 601-19

Gurnell, A.M., K.J. Gregory, and G.E. Petts. 1995. The role of coarse woody debris in

forest aquatic habitats – Implications for management. Aquatic Conservation:

Marine and Freshwater Ecosystems 5: 143-166.

Hannah, D.M., P. J. Wood, and J.P. Sadler. 2004. Ecohydrology and hydroecology: A

new paradigm? Hydrological Processes 18: 3439-3445.

Hansen, A.J., T.A. Spies, F.J. Swanson, and J.L. Ohmann. 1991. Conserving biodiversity

in managed forests: Lessons from natural forests. Bioscience 41: 382-392.

Harmon, M.E., J.F. Franklin, F.J. Swanson, P. Sollins, S.V. Gregory, J.D. Lattin, N.H.

Anderson, S.P. Cline, N.G. Aumen, J.R.Sedell, G.W. Lienkaemper, K. Cromack,

Jr., and K.W. Cummins. 1986. Ecology of Coarse Woody Debris in Temperate

Ecosystems. Academic Press, Inc. Orlando, Florida.

Harper, D., J. Mekotova, S. Hulme, J. White, J. Hall. 1997. Habitat heterogeneity and

aquatic invertebrate diversity in floodplain forests. Global Ecology and

Biogeography Letters 6: 275-85.

Hefting, M., Clement, J.C., Dowrick, D., Cosandey, A.C., Bernal, S., Cimpian, C., Tatur, A.,

Burt, T.P. and Pinay, G. 2004. Water table elevation controls on soil nitrogen

227

cycling in riparian wetlands along a European climatic gradient. Biogeochemistry

67: 113-134.

Helmus, M.R., and G.G. Sass. 2008. The rapid effects of a whole-lake reduction of

coarse woody debris of fish and benthic macroinvertebrates. Freshwater Biology

53: 1423-1433.

Hillman, T.P and G.P Quinn. 2002. Temporal changes in macroinvertebrate

assemblages following experimental flooding in permanent and temporary

wetlands in an Australian floodplain forest. River Research and Applications 18:

137-54.

Houlahan, J.E. and C.S. Findlay. 2004. Estiamting the ‘critical’ distance at which

adjacent land-use degrades wetland water and sediment quality. Landscape

Ecology 19: 677-690.

Huryn, A. D., and K. E. Gibbs. 1996. Riparian Sedge Meadows in Maine: A

Macroinvertebrate Community Structured by River-Floodplain Interaction.

Pages 363-382 in D. P. Batzer, R. B. Rader and S. A. Wissinger (editors).

Invertebrates in Freshwater Wetlands of North America: Ecology and

Management. John Wiley and Sons.

Huryn, A.D. and J.B. Wallace. 2000. Life history and production of stream insects.

Annual Review of Entomology 45: 83-110.

228

Jackson, M.B. and T.D. Colmer. 2005. Response and adaptation by plants to flooding

stress. Annals of Botany 96: 501-505.

Jeffries, M. 1994. Invertebrate communities and turnover in wetland ponds affected by

drought. Freshwater Biology 32:603-12.

Jenkins, K. M. and A. J. Boulton. 2003. Connectivity in a Dryland River: Short-Term

Aquatic Microinvertebrate Recruitment Following Floodplain Inundation.

Ecology 84: 2708-2723.

Johansson, M. E., and C. Nilsson. 2002. Responses of Riparian Plants to Flooding in

Free-Flowing and Regulated Boreal Rivers: An Experimental Study. The Journal

of Applied Ecology 39: 971-986.

Johnson, N.F. and C. A. Triplehorn. 2004. Borror and DeLong’s Introduction to the

Study of Insects. Brooks Cole, New York.

Jones, T.A., and L.D. Daniels. 2008. Dynamics of large woody debris in small streams

disturbed by the 2001 Dogrib fire in the Alberta foothills. Forest Ecology and

Management 256: 1751-1759.

Junk, W.J., M.Brown, I.C. Campell, M. Finlayson, B. Gopal, L. Ramsburg, and B.G.

Warner. 2006. The comparative biodiversity of seven globally important

wetlands: a synthesis. Wetlands 68: 400-414.

229

Junk, W. J., P. B. Bayley, and R. E. Sparks. 1989. The flood pulse concept in river-

floodplain systems. Canadian Special Publication of Fisheries and Aquatic

Science 106: 110-127.

Junk, W.J. and K.M Wantzen. 2004. The flood pulse concept: new aspects,

approaches, and applications – an update. pp. 117-140 In: R. Welcomme and

T. Petr (eds.). Proceedings of the second international symposium on the

management of large rivers for fisheries. FAO regional office for Asia and the

Pacific , Bangkok, Thailand.

Jurik, T.W., S.C. Wang, and A.G. van der Valk. 1994. Effects of sediment load on

seedling emergence from wetland seed banks. Wetlands 14: 159-165.

Justin, S.H.F.W., and W. Armstrong. 1987. The anatomical characteristics of roots and

plant response to soil flooding. New Phytology 106: 465-495.

Kail, J., H. Hering, S. Muhar, M. Gerhard, and S. Preis. 2007. The use of large wood in

stream restoration: Experiences from 50 projects in Germany and Austria.

Journal of Applied Ecology 44: 1145-1155.

Kennedy, R.A., M.E. Rumpho and T.C. Fox. 1992. Anaerobic metabolism in plants.

Plant Physiology 100: 1-6.

230

Kercher, S.M., Herr-Turoff, A., and J.B. Zedler. 2007. Understanding invasion as a

process: The case of Phalaris arundinacea in wet prairies. Biological Invasions 9:

657-665.

Kolasa, J., L.L. Manne, and S.N. Pandit. 2012. Species-area relationships arise from

interaction of habitat heterogeneity and species pool. Hydrobiologia 685: 135-

144.

Koning, C.A. 2004. Impacts of small amounts of sandy sediment on wetland soils and

vegetation: Results from field and greenhouse studies. Wetlands 24: 295-308.

Kozlowski TT. 2002. Physiological-ecological impacts of flooding on riparian forest

ecosystems. Wetlands 22:550-61.

Lake, P.S., N. Bond, and P. Reich. 2007. Linking ecological theory with stream

restoration. Freshwater Biology 52: 597-615.

Lamberti, G.A., D.T. Chaloner, and A.E. Hershey. 2010. Linkages among aquatic

ecosystems. Journal of North American Benthological Society 29: 245-263.

Laser, M., J. Jordan, and K. Nislow. 2009. Riparian forest and instream large wood

characteristics, West Branch Sheepscot River, Maine, USA. Forest Ecology and

Management 257: 1558-1565.

231

Lenssen, J. P. M., G.E. ten Dolle, and C.WP.M. Blom 1998. The effect of flooding on

the recruitment of reed marsh and tall forb plant species. Plant Ecology 139: 13-

23.

Lenssen, J.P.M . and de Kroon, H. 2005. Abiotic constraints at the upper boundaries of

two Rumex species on a freshwater flooding gradient. Journal of Ecology 93:

138-147.

Lenssen, J.P.M, van de Steeg, H.M. and de Kroon, H. 2004. Does disturbance favor

weak competitors? Mechanisms of changing plant abundance after flooding.

Journal of Vegetation Science 15: 305-314.

Lenssen,J.P.M., F.B.J. Menting, W.H. van der Putten, and C.W.P.M. Blom. 1999.

Effects of sediment type and water level on biomass production of wetland plant

species. Aquatic Botany 64: 151-165.

Lester, R.E., W.Wright, and M. Jones-Lennon. 2007. Does adding wood to agricultural

streams enhance biodiversity? An experimental approach. Marine and

Freshwater Research 58: 687-698.

Mackenzie, R.A. and J.L. Kaster. 2004. Temporal and spatial patterns of insect

emergence from a Lake Michigan coastal wetland. Wetlands 24: 688-700.

232

Mahaney, W. M., D. H.Wardrop, and R. P. Brooks. 2004. Impacts of stressors on the

emergence and growth of wetland plant species in Pennsylvania, USA. Wetlands

24: 538-549.

Malard, F., U. Uehlinger, and R. Zar. 2006. Flood pulse and riverscape dynamics in a

braided glacial river. Ecology 87: 704-716.

Malmqvist, B. 2002. Aquatic invertebrates in riverine landscapes. Freshwater Biology

47: 679-694.

Marzluff, J.M. and K. Ewing. 2001. Restoration of fragmented landscapes for the

conservation of birds: A general framework and specific recommendations for

urbanizing landscapes. Restoration Ecology 9(3): 280-292.

McBride, J.R. and J. Strahan. 1984. Establishment and survival of woody riparian

species on gravel bars of an intermittent steam. American Midland Naturalist

112: 235-245.

McCune, B. and M. J. Mefford. 2006. PC-ORD. Multivariate Analysis of Ecological

Data. Version 5. MjM Software, Gleneden Beach, Oregon, U.S.A.

McGill, Brian J.2010. Matters of scale. Science 328: 575-576.

233

McIllroy, S.K. C. Montage, C.A. Jones, and B.L. McGlynn. 2008. Identifying linkages

between land use, geomorphology, and aquatic habitat in a mixed-use

watershed. Environmental Management 42: 867-876.

Mensing, D.M., S.M. Galatowitsch, and J.R. Tester. 1998. Anthropogenic effects on

the biodiversity of riparian wetlands of a northern temperate landscape. Journal

of Environmental Management 53: 349-77.

Merritt, R. W., K. W. Cummins and M. B. Berg (editors). 2008. An introduction to the

aquatic insects of North America. Fifth edition. Kendall/Hunt Publishing

Company.

Middleton, B. 1999. Wetland restoration, flood pulsing, and disturbance dynamics.

John Wiley and Sons. New York.

Middleton, B. 2002. Flood-pulsing in wetlands: Restoring the natural hydrological

balance. Wiley, New York, NY.

Mitsch, W.J., and J.G. Gosselink. 2002. The value of wetlands: Importance of scale and

landscape setting. Ecological Economics 35: 25-33.

Mitsch, W.J., and J.G. Gosselink. 2007. Wetlands. 4th edition. John Wiley and Sons,

Inc., New York.

Molles, Jr., M.C., C.S. Crawford, L.M. Ellis, H.M. Valett, and C.N. Dahm. 1998.

234

Managed flooding for riparian ecosystem restoration. Bioscience 48: 749-56.

Moog, P.R. and P. Janiesch. 1990. Root growth and morphology of Carex species as

influenced by oxygen deficiency. Functional Ecology 4: 201-208.

Moon, M., M.R. Rattry, F.E. Putz, and G. Bowes. 1993. Acclimatization to flooding of

the herbaceous vine, Minkania scandens. Functional Ecology 7: 610-615.

Naiman, R. J., and H. Decamps. 1997. The ecology of interfaces: Riparian zones.

Annual Review of Ecology and Systematics 28: 621-658.

Nakamura, F., and F.J. Swanson. 1993. Effects of coarse woody debris on morphology

and sediment storage of a mountain stream in Western Oregon. Earth Surface

Processes and Landforms 18: 43-61.

Nilsson, C., and M. Svedmark. 2002. Basic principles and ecological consequences of

changing water regimes: Riparian plant communities. Environmental

Management 30: 468-480.

Odum, W. E., E. P. Odum, and H. T. Odum. 1995. Natures Pulsing Paradigm.

Estuaries 18: 547-555.

Ohio Department of Resources. 2010. First year monitoring report for the Mud Brook

Preserve wetland mitigation area, City of Hudson, Summit County, Ohio.

Prepared by URS Corporation, Cleveland, Ohio. Last accessed: 3 march 2012.

235

http://www.dot.state.oh.us/Divisions/Planning/Environment/Ecological_Resourc

es_Permits/MitigationInventory/Documents/summit_8_Mud_Brook/mud_brook

_year_1_comp.pdf

Olila, O.G., K.R. Reddy, and D.L. Stites. 1997. Influence of draining on soil phosphorus

forms and distribution in a constructed wetland. Ecological Engineering 9(3):

157-169.

Palmer, M.A., H.L. Menninger, and E. Bernhardt. 2010. River restoration, habitat

heterogeneity and biodiversity: A failure of theory or practice? Freshwater

Biology 55: 205-222.

Pant, H.K. and K.R. Reddy. 2001. Hydrologic influence on stability of organic

phosphorus in wetland detritus. Journal of Environmental Quality 30 (2): 668-

674.

Parsons, M., C.A. Loughlin, K.A. Kotschy, K.H. Rogers and M.W. Rountree. 2005. The

effects of extreme floods on the biophysical heterogeneity of river landscapes.

Frontiers in Ecology and the Environment 3: 487-494.

Pennings, S.C., and R.M.Callaway. 1992. Salt-marsh plant zonation: The relative

importance of competition and physical factors. Ecology 73: 681-690.

236

Pettit, N.E., and R.J. Naiman. 2006. Flood-deposited wood creates regeneration niches

for riparian vegetation on semi-arid South African river. Journal of Vegetation

Science 17: 615-624.

Piegay, H. 1997. Interactions between floodplain forests and overbank flows: Data

from three piedmont rivers of southeastern France. Global Ecology and

Biogeography Letters 6: 187-96.

Pinay, G., J.C. Clement, and R.J. Naiman. 2002. Basic principles and ecological

consequences of changing water regimes on nitrogen cycling in fluvial systems.

Environmental Management 30: 481-91.

Pittman, S. E., T. King, S. Faurby, and M. E. Dorcas. 2011. Genetic and demographic

status of an isolated bog turtle (Glyptemys muhlenbergii) population:

implications for the conservation of small populations of long-lived animals.

Conservation Genetics 4: 781-790.

Plum, N. 2005. Terrestrial invertebrates in flooded grassland: A literature review.

Wetlands 25: 721-37.

Plum, N.M. and J. Filser. 2005. Floods and drought: Response of earthworms and

potworms (Oligochaeta : Lumbricidae, enchytraeidae) to hydrological extremes

in wet grassland. Pedobiologia 49: 443-53.

237

Poff, N.L., Allen, J.D., Bain, M.B., Karr, J.R., Prestegaard, K.L., Richter, B.D., Sparks, R.E.

and Stromberg, J.C. 1997. The natural flow regime: A paradigm for river

conservation and restoration. Bioscience 47: 769-784.

Polis, G. A., W. B. Anderson, and R. D. Holt. 1997. Toward an integration of

landscape and food web ecology: The dynamics of spatially subsidized food

webs. Annual Review of Ecology and Systematics 28: 289-316.

Pollock, M.M., R.J. Naiman, and T. A. Hanley. 1998. Plant species richness in riparian

wetlands - A test of biodiversity theory. Ecology 79: 94-105.

Pringle, C.M. 2001. Hydrologic connectivity and the management of biological reserves:

A global perspective. Ecological Applications 11: 981-98.

Pyle, C., and M.M. Brown. 1998. A rapid system of decay classification for hardwood

logs of the eastern deciduous forest floor. Journal of the Torrey Botanical

Society 125 : 237-45.

Quinn, G.P., T.J. Hillman, and R. Cook. 2000. The response of macroinvertebrates to

inundation in floodplain wetlands: A possible effect of river regulation?

Regulated Rivers Research and Management 16: 469-77.

Reed, P. 1988. National list of plant species that occur in wetlands. National Wetland

Inventory. U.S. Department of the Interior, Fish and Wildlife Service, St.

Petersburg, Fl.

238

Reed, P. B., Jr. 1997. 1997 revision of the national list of plant species that occur in

wetlands. US Department of the Interior, Fish and Wildlife Service. Washington

DC. 470 p. http://www.nwi.fws.gov/ecology/htm

Reese, E. G., and D. P. Batzer. 2007. Do invertebrate communities in floodplains

change predictably along a river's length? Freshwater Biology 52: 226-239.

Robertson, A. I., P. Bacon, and G. Heagney. 2001. The responses of floodplain primary

production to flood frequency and timing. Journal of Applied Ecology 38: 126-

136.

Robinson, C. T., K. Tockner, and J. V. Ward. 2002. The fauna of dynamic riverine

landscapes. Freshwater Biology 47: 661-677.

Roni, P., K. Hanson, and T. Beechie. 2008. Global review of the physical and biological

effectiveness of stream habitat rehabilitation techniques. North American

Journal of Fisheries Management 28: 856-890.

Roumet, C., C. Urcelay, and S. Diaz. 2006. Suites of root traits differ between annual

and perennial species growing in the field. New Phytologist 170: 357-368.

Rundio, D.E., and S.T. Lindley. 2012. Reciprocal fluxes of stream and riparian

invertebrates in a coastal California basin with Mediterranean climate.

Ecological Restoration.DOI: 10.1007/s11284-011-0920-6.

239

Sanderson, J.S., N.B. Kotilar, and D.A. Steingraeber. 2008. Opposing environmental

gradients govern vegetation zonation in an intermountain playa. Wetlands 28:

1060-1070.

Schmitz, A. and J.F. Harrison. 2004. Hypoxic tolerance in air-breathing invertebrates.

Respiratory Physiology and Neurobiology 141: 229-242.

Schneider, D.C. 2001. The rise of the concept of scale in ecology. Bioscience 51(7): 545-

553.

Seabloom, E.W. and A.G. van der Valk. 2003. The development of vegetative zonation

patterns in restored prairie pothole wetlands. Journal of Applied Ecology 40: 92-

100.

Seabloom, E.W., A.G. van der Valk, and K.A. Moloney. 1998. The role of water depth

and soil temperature in determining initial composition of prairie wetland

coenoclines. Plant Ecology 138: 203-216.

Sheeley, D.G. and L.M. Smith. 1989. Tests of diet and condition bias in hunter killed

northern pintails. Journal of Wildlife Management 53: 765-769

Sheldon, F., A.J. Boulton, and J.T. Puckridge. 2002. Conservation value of variable

connectivity: Aquatic invertebrate assemblages of channel and floodplain

habitats of a central Australian arid-zone river, Cooper Creek. Biological

Conservation 103: 13-31.

240

Shipley, B., P.A. Keddy, and L.P. Lefkovitch. 1991. Mechanisms producing plant

zonation along a water depth gradient: A comparison with the exposure

gradient. Canadian Journal of Botany 69: 1420-1424.

Sih, A., P. Crowley, M. McPeek, and K. Stohmier. 1985. Predation, competition, and

prey communities: A review of field experiments. Annual Review of Ecology and

Systematics 16: 269-311.

Silvestri, S., A. Defina, and M. Marani. 2005. Tidal regime, salinity and salt marsh plant

zonation. Estuarine Coastal and Shelf Science 62: 119-130.

Smock, L. A. 1994. Movements of invertebrates between stream channels and

forested floodplains. Journal of the North American Benthological Society 13:

524-531.

Smock, L. A. 1999. Riverine floodplain forests of the southeastern United States:

invertebrates in an aquatic-terrestrial ecotone. Pages 137-166 in D. P. Batzer,

R. B. Rader and S. A. Wissinger (editors). Invertebrates in Freshwater

Wetlands of North America. John Wiley and Sons, Inc., New York.

Sparks, R.E. 1995. Need for ecosystem management of large rivers and their floodplain.

Bioscience 45: 168-183.

Spieles, D.J. and W.J. Mitsch. 2003. A model of macroinvertebrate trophic structure

and oxygen demand in freshwater wetlands. Ecological Modeling 161: 183-194.

241

Spieles, D.J., and J.D. Horn. 2009. Macroinvertebrate community structure in created

wetlands of different successional stage. Aquatic Ecosystem Health and

Management 12: 320-329.

Spies, T.A., J.F. Franklin, and T.B. Thomas. 1988. Coarse woody debris in Douglas-Fir

forests of western Oregon and Washington. Ecology 69: 1689-1702.

Spivak, A.C., M.J. Vanni, and E.M. Mette. 2011. Moving on up: Can results from simple

aquatic mesocosm experiments be applied across broad spatial scales?

Freshwater Biology 56: 279-291.

Spyeas, B.W.W., A.E. Plocher, D.M. Ketzner, J.W. Matthews, J.L. Ellis, and E.J. Heske.

2010. Biological consequences of invasion by reed canary grass (Phalaris

arundinacea). Biological Invasions 12: 1253-1267.

Stagliano, D.M., A.C. Benke, and D.H. Anderson. 1998. Emergence of aquatic insects

from 2 habitats in a small wetland of the southeastern USA: Temporal patterns

of numbers and biomass. Journal of the North American Benthological Society

17: 37-53.

Stanczak, M. and J.B. Keiper. 2004. Benthic invertebrates in adjacent created and

natural wetlands in northeastern Ohio, USA. Wetlands 24: 212-218.

Steiger, J., E. Tabacchi, S. DuFour, D. Corenblit, and J.L. Reiry. 2005.

Hydrogeomorphic processes affecting riparian habitat within alluvial channel-

242

floodplain river systems: A review for the temperate zone. River Research and

Applications 21: 719-737.

Steiger, J., A. M. Gurnell, and J. M. Goodson. 2003. Quantifying and characterizing

contemporary riparian sedimentation. River Research and Applications 19: 335-

352.

Stewart, T.W. and J.A. Downing. 2004. Macroinvertebrate communities and

environmental conditions in recently constructed wetlands. Wetlands 28: 141-

150.

Strack, M., J.M. Waddington, R.A. Bourbonniere, E.L. Buckton, K. Shaw, P.

Whittington, and J.S. Price. 2008. Effect of water table drawdown on peatland

dissolved carbon export and dynamics. Hydrological Processes DOI:

10.1002/hyp.6931.

Strader R.W. and P.H. Stinson. 2005. Moist-Soil Management Guidelines for the U.S.

Fish and Wildlife Service Southeast Region Migratory Bird Field Office Division of

Migratory Birds. Southeast Region, U.S. Fish and Wildlife Service. Jackson, MS.

Tabacchi, E. 1995. Structural variability and invasions of pioneer plant communities in

riparian habitats of the middle Ardour River (SW France). Canadian Journal of

Botany 73: 33–44.

243

Tabacchi, E., L. Lambs, H. Guilloy, A.M. Planty-Tabacchi, E. Muller, and H. Decamps.

2000. Impacts of riparian vegetation on hydrological processes. Hydrological

Processes 14: 2959-2976.

Therriault, T.W. and J. Kolasa. 2000. Explicit links among physical stress, habitat

heterogeneity and biodiversity. Oikos 89: 387-391.

Thomaz, S.M., and L.M. Bini. 2007. Floods increase similarity among aquatic habitats in

river-floodplain systems. Hydrobiologia 579: 1-13.

Thompson, K., M.C. Miller, and T.M. Culley. 2007. Comparison of plant species

richness, diversity, and biomass in Ohio wetlands. Ohio Journal of Science 107

(3): 32-38.

Thompson, R.M., N.R. Phillips, and C.R. Townsend. 2009. Biological consequences of

clear-cut logging around streams – Moderating effects of management. Forest

Ecology and Management 257: 931-940.

Thoms MC. 2003. Floodplain-river ecosystems: Lateral connections and the

implications of human interference. Geomorphology 56: 335-349.

Thorp, J. H., and A. P. Covich (editors). 2001. Ecology and Classification of North

American Freshwater Invertebrates. 2nd edition. Academic Press.

Timms, B.V. and A.J. Boulton. 2001. Typology of arid-zone floodplain wetlands of the

244

Paroo river (inland Australia) and the influence of water regime, turbidity, and

salinity on their aquatic invertebrate assemblages. Archiv fuer

Hydrobiologie 153: 1-27.

Tiner, R. 1991. The concept of a hydrophyte for wetland identification. Bioscience 41:

236-246.

Tiner, R.W. 2006. Note: Lists of potential hydrophytes for the United States: A regional

review and their use in wetland identification. Wetlands 26: 624-634.

Tockner, K., F. Malard, U. Uehlinger, and J.V. Ward. 2002. Nutrients and organic

matter in a glacial river-floodplain system (Val Roseg, Switzerland). Limnology

and Oceanography 47: 266-77.

Tockner, K., D. Pennetzdorfer, N. Reiner, F. Schiemer, and J. V. Ward. 1999a.

Hydrological connectivity, and the exchange of organic matter and nutrients in a

dynamic river-floodplain system (Danube, Austria). Freshwater Biology 41: 521-

535.

Tockner, K., F. Malard, and J. V. Ward. 2000. An extension of the flood pulse concept.

Hydrological Processes 14: 2861-2883.

Tockner, K.F. and J.A. Stanford. 2002. Riverine flood plains: Present state and future

trends. Environmental Conservation 29(3): 308-330.

245

Tockner, K., F. Scheimer, C. Baumgartner, G. Kum, E. Weigand, I. Zweimuller, and J.V.

Ward. 1999. The Danube restoration project: Species diversity across

connectivity gradients in the floodplain system. Regulated Rivers Research and

Management 15: 245-258.

Tockner, K., M.S. Lorang, and J.A. Stanford. 2010. River flood plains are model

ecosystems to test general hydrogeomorphic and ecological concepts. River

Research and Applications 26: 76-86.

Tolonen, K.T., H. Hamalainen, I.J. Holopainen, K. Mikkonen and J. Karjalainen. 2003.

Body size and substrate association of littoral insects in relation to vegetation

structure. Hydrobiologia 499: 179-190.

Toogood, S. and C.B. Joyce. 2009. Effects of raised water levels on wet grassland plant

communities. Applied Vegetation Science 12: 283-294.

Trebitz, A.S., J.C. Brazner, V.J. Brady, R. Axler, and D.K. Tanner. 2007. Turbidity

tolerances of great lakes coastal wetland fishes. North American Journal of

Fisheries Management 27: 619-633.

Tronstad, L.M., B.P. Tronstad, and A.C. Benke. 2005a. Invertebrate responses to

decreasing water levels in a subtropical river floodplain wetland. Wetlands 25:

583-593.

Tronstad, L.M., B.P. Tronstad, and A.C. Benke. 2005b. Invertebrate seedbanks:

246

Rehydration of soil from an unregulated river floodplain in the south-eastern US.

Freshwater Biology 50: 646-655.

Uetz, G.W.,K.L.V.D Laan, G.F. Summers, P.A.K. Gibson, and L.L. Getz. 1979. The

effects of flooding on floodplain arthropod distribution, abundance and

community structure. American Midland Naturalist 101: 286-299.

Ulyshen, M.D. and J.L. Hanula. 2009. Responses of arthropods to large-scale

manipulations of dead wood in Loblolly Pine stands of the southeastern United

States. Environmental Entomology 38: 1005-1012.

USDA. 2011. United States Department of Agriculture, Natural Resources Conservation

Services. Plants Database. http://plants.usda.gov/ (http://plants.usda.gov, 18

July 2011). National Plant Data Team, Greensboro, NC 27401-4901 USA.

USGS. 2009. "USGS Real-Time Water Data for USGS 04206000 Cuyahoga River at Old

Portage OH." USGS Water Data for the Nation.

/oh/nwis/uv?site_no=04206000>.

Van de Meutter, F., L. de Meester, and R. Stoks. 2005. Water turbidity affects

predator-prey interactions in a fish-damselfly system. Oecologia 144: 327-336.

247

van der Sman, A.J.M., N.N. Joosten, and C.W.P.M. Blom. 1993. Flooding regimes and

life-history characteristics of short-lived species in river forelands. Journal of

Ecology 81: 121-130.

Vannote, R.L., G.W. Minshall, K.W. Cummins, J.R. Sedell, and C.E. Cushing. 1980. The

river continuum concept. Canadian Journal of Fisheries and Aquatic Sciences 37:

130-137.

Vargo, S.M., R.K. Neely, and S.M. Kirkwood. 1998. Emergent plant decomposition and

sedimentation: Response to sediments varying in texture, phosphorus content

and frequency of deposition. Environmental and Experimental Botany 40: 43-58.

Vartapetian, B. B., and M. B. Jackson. 1997. Plant adaptations to anaerobic stress.

Annals of Botany 79: 3-20.

Verschuren, D., J. Tibby, K. Sabbe, and N. Roberts. 2000. Effects of depth, salinity, and

substrate on the invertebrate community of a fluctuating tropical lake. Ecology

81: 164-182.

Vivian-Smith, G. 1997. Microtopographic heterogeneity and floristic diversity in

experimental wetland communities. The Journal of Ecology. 85: 71-82.

Voss, E. G. (editor). 2001. Michigan Flora. Cranbrook Institute of Science, Bloomfield

Hills, Michigan.

248

Vretere, V., S.E.B. Weisner, J.A. Strand, and W. Graneli. 2001. Phenotypic plasticity in

Phragmites australis as a functional response to water depth. Aquatic Botany

69: 127-145.

Wall, D.H. and J.C. Moore. 1999. Interactions underground: Soil biodiversity,

mutualism, and ecosystem processes. Bioscience 49: 109-117.

Walsh, C.J., K.A. Waller, J. Gehling, and R. Mac Nally. 2007. Riverine invertebrate

assemblages degraded more by catchment urbanization than by riparian

deforestation. Freshwater Biology 52: 574-587.

Ward, J.V., K. Tockner, and F. Schiemer. 1999. Biodiversity of floodplain river

ecosystems: Ecotones and connectivity. Regulated Rivers Research and

Management 15: 125-39.

Ward, J. V., K. Tockner, D. B. Arscott, and C. Claret. 2002. Riverine landscape

diversity. Freshwater Biology 47: 517-539.

Ward, J.V. and J.A. Stanford. 1995. An ecosystem perspective of alluvial rivers:

Connectivity and the hyporheic corridor. Journal of the North American

Benthological Society 12: 105-119.

249

Webster, C.R., C.J.F. Huckins, and J.M. Shields. 2008. Spatial distribution of riparian

zone coarse woody debris in a managed northern temperate watershed.

American Midland Naturalist 159: 225-237.

Weiher, E., and P.A. Keddy. 1995. The assembly of experimental wetland communities.

Oikos 73:323-335.

Weiher, E., P.G.D. Clarke, and P.A. Keddy. 1998. Community assembly rules,

morphological dispersion, and the coexistence of plant species. Oikos 81: 309-

322.

Whiles, M.R. and B.S. Goldowitz. 2001. Hydrologic influences on insect emergence

production from Central Platte River wetlands. Ecological Applications 11: 1829-

1842.

Whiles, M.R. and B.S. Goldowitz. 2005. Macroinvertebrate communities in Central

Platte River wetlands: Patterns across a hydrologic gradient. Wetlands 25: 462-

72.

Williams, D.D. 1997. Temporary ponds and their invertebrate communities. Aquatic

Conservation Marine and Freshwater Ecosystems 7: 105-117.

Williams, D.D., N. Heeg, and A.K. Magnusson. 2007. Habitat background selection by

colonizing intermittent pond invertebrates. Hydrobiologia 592: 487-498.

Wissinger, S., J. Steinmetz, J.S. Alexander, and W. Brown. 2004. Larval cannibalism,

250

time constraints, and adult fitness in caddisflies that inhabit temporary wetlands.

Oecologia 138: 39-47.

Zar, J.H. 1999. Biostatistical analysis. 4th edition. Prentice Hall, Upper Saddle River,

New Jersey.

Zelnik, I., and A. Carni. 2008. Distribution of plant communities, ecological strategy

types and diversity along a moisture gradient. Community Ecology 9: 1-9.