OLIGOTROPHICATION EFFECTS FOLLOWING DIVERSION OF WASTE EFFLUENT FROM AN EMBAYMENT OF LAKE MARTIN, ALABAMA Except where reference is made to the work of others, the work described in this thesis is my own or was done in collaboration with my advisory committee. This thesis does not include proprietary or classified information. __________________________________ Richard Jason Dickey Certificate of Approval: Michael J. Maceina David R. Bayne, Chair Professor Professor Fisheries and Allied Aquaculture Fisheries and Allied Aquaculture Elise R. Irwin Joe L. Pittman Professor Interim Dean Fisheries and Allied Aquaculture Graduate School OLIGOTROPHICATION EFFECTS FOLLOWING DIVERSION OF WASTE EFFLUENT FROM AN EMBAYMENT OF LAKE MARTIN, ALABAMA Richard Jason Dickey A Thesis Submitted to the Graduate Facility of Auburn University in Partial Fulfillment of the Requirements for the Degree of Master of Science Auburn, Alabama December 15, 2006 iii OLIGOTROPHICATION EFFECTS FOLLOWING DIVERSION OF WASTE EFFLUENT FROM AN EMBAYMENT OF LAKE MARTIN, ALABAMA Richard Jason Dickey Permission is granted to Auburn University to make copies of this thesis at its discretion, upon the request of individuals or institutions and at their expense. The author reserves all publication rights. Signature of Author Date of Graduation iv VITA Richard Jason Dickey, son of Wayne and Linda (Pappas) Dickey, was born Friday, January 13 th , 1978 in Oak Ridge, Tennessee. He graduated from Oak Ridge High School in 1996. In 2002, he received his Bachelors of Science in Wildlife and Fisheries Science from the University of Tennessee, Knoxville. During his undergraduate career, he gained internship experience with the Environmental Sciences Department at Oak Ridge National Laboratories and Cadmus Group, Inc. After graduation he spent one year working with the U.S. Fish and Wildlife Service in Panama City, Florida as a volunteer with Americorps/ Student Conservation Association. In August, 2003, he entered the Graduate School at Auburn University. v THESIS ABSTRACT OLIGOTROPHICATION EFFECTS FOLLOWING DIVERSION OF WASTE EFFLUENT FROM AN EMBAYMENT OF LAKE MARTIN, ALABAMA Richard Jason Dickey Master of Science, December 15, 2006 (B.S., University of Tennessee, 2002) 116 Typed Pages Directed by David R. Bayne I examined the effects of nutrient diversion on water quality, phytoplankton, and macroinvertebrate and fish community structures over a 10-year period in Lake Martin, Alabama. In 2001, a pipeline began diverting treated effluent from Alexander City, Alabama out of Elkahatchee Creek Embayment (ECE) and into the mainstem of Lake Martin reservoir. Oligotrophication in ECE produced measurable responses within the biotic community, including: 1) decreased algal biomass; 2) a shifting black bass community; 3) decreased catch and condition of sunfish and crappie; and 4) altered structure and function of littoral macroinvertebrate communities. Elevated conductivity at the mouth of Dennis Creek Embayment (DCE), the reference site, suggested the encroachment of diverted effluent. No changes conclusive to nutrient enrichment were found with respect to: key nutrients, algal biomass, or macroinvertebrate assemblages. vi However, some Centrarchid fish responded in a manner consistent with eutrophication. Members of Lepomis expressed higher condition following diversion. In addition, the black bass dynamic shifted to favor longer largemouth bass (Micropterus salmoides) with higher relative weights, while spotted bass (Micropterus punctulatus) were fewer, shorter, and less robust. The aquatic communities in this study provided a measurable response to subtle changes in trophic status which underscores the sensitivity of the aquatic community. The ability to detect and predict these subtle changes enables reservoir managers to make more confident predictions and ultimately lead to improved design and management of reservoirs. vii ACKNOWLEDGEMENTS I wish to thank Dr. David Bayne, Wendy Seesock, Eric Reutebuch, and Dr. Cliff Webber for their continued support and guidance. I also wish to thank Dr. Mike Maceina and Dr. Elise Irwin for their counsel and statistical guidance. Peer support is also essential for graduate success and I am particularly gracious for the outstanding friendships I?ve developed. Finally, a sincere thanks to my parents for providing me the support and encouragement to realize my goals. viii Style manual or journal used: Transactions of the American Fisheries Society Computer software used: Microsoft Word, Excel 2002 for Windows, SAS version 9.1.3 ix TABLE OF CONTENTS LIST OF TABLES .................................................................................................... x LIST OF FIGURES ................................................................................................... xii I. INTRODUCTION ......................................................................................... 1 II. LITERATURE REVIEW .............................................................................. 3 III. MATERIALS AND METHODS................................................................... 10 Study Sites ..................................................................................................... 10 Water Quality Sampling ................................................................................ 11 Phytoplankton Sampling................................................................................ 12 Fish Sampling ................................................................................................ 12 Macroinvertebrate Sampling.......................................................................... 14 IV. RESULTS ...................................................................................................... 17 Water Quality................................................................................................. 17 Phytoplankton ................................................................................................ 18 Fish................................................................................................................. 19 Benthic Invertebrates from Hester-Dendy Multiplate Samplers ....... 24 Benthic Invertebrates from Bottom Dredge Samples ........................ 26 V. DISCUSSION................................................................................................ 29 Elkahatchee Creek Embayment ..................................................................... 29 Dennis Creek Embayment ............................................................................. 33 Summary........................................................................................................ 40 LITERATURE CITED .............................................................................................. 66 APPENDICES ........................................................................................................... 77 x LIST OF TABLES Table 1 Analytical methods used for assessing water quality variables in ECE and DCE in Lake Martin, Alabama ....................... 39 Table 2 Sampling schedule for water quality, phytoplankton, invertebrates, and fish during the 10?year study period in ECE and DCE in Lake Martin, Alabama....................................... 40 Table 3 Summary of Pearson Correlation Coefficients for chlorophyll a concentrations in Elkahatchee Creek Embayment before nutrient removal and the water quality variables that were significantly correlated ...................................................................... 41 Table 4 Summary of Pearson Correlation Coefficients for chlorophyll a concentrations in Elkahatchee Creek Embayment following nutrient removal and the water quality variables that were significantly correlated ...................................................................... 42 Table 5 Summary of Pearson Correlation Coefficients for chlorophyll a concentrations in Dennis Creek Embayment and the water quality variables that were significantly correlated........................... 42 Table 6 Mean values (? SD) of benthic invertebrate assemblage variables from Hester-Dendy multiplate samples taken during the fall season in Elkahatchee Creek Embayment in Lake Martin, Alabama............ 43 Table 7 Mean values (? SD) of benthic invertebrate assemblage variables from Hester-Dendy multiplate samples taken during the summer season in Elkahatchee Creek Embayment in Lake Martin, Alabama 44 Table 8 Mean values (? SD) of benthic invertebrate assemblage variables from Hester-Dendy multiplate samples taken during the fall season in Dennis Creek Embayment in Lake Martin, Alabama.................... 45 xi Table 9 Mean values (? SD) of benthic invertebrate assemblage variables from Hester-Dendy multiplate samples taken during the summer season in Dennis Creek Embayment in Lake Martin, Alabama ........ 46 Table 10 Mean values (? SD) of benthic invertebrate assemblage variables from dredge samples taken during the fall season in Elkahatchee Creek Embayment in Lake Martin, Alabama .................................... 47 Table 11 Mean values (? SD) of benthic invertebrate assemblage variables from dredge samples taken during the summer season in Elkahatchee Creek Embayment in Lake Martin, Alabama .................................... 48 Table 12 Mean values (? SD) of benthic invertebrate assemblage variables from dredge samples taken during the fall season in Dennis Creek Embayment in Lake Martin, Alabama............................................... 49 Table 13 Mean values (? SD) of benthic invertebrate assemblage variables from dredge samples taken during the summer season in Dennis Creek Embayment in Lake Martin, Alabama .................................... 50 xii LIST OF FIGURES Figure 1 Map of Lake Martin, Alabama and study area ........................................ 51 Figure 2 A map of the upstream portion of Lake Martin, Alabama, also containing the study sites, Elkahatchee Creek Embayment (ECE) and Dennis Creek Embayment (DCE) ..................... 52 Figure 3 Mean (? SE) growing season soluble reactive phosphorous (SRP) concentrations measured from the photic zone at the mid-embayment sites in Lake Martin at Stations E1 and D1 of Elkahatchee Creek Embayment and Dennis Creek Embayment, respectively, from 1994?2004................................................................................................ 53 Figure 4 Mean (? SE) growing season soluble reactive phosphorous (SRP) concentrations measured from the photic zone at the downstream sites in Lake Martin at Stations E2 and D2 of Elkahatchee Creek Embayment and Dennis Creek Embayment, respectively, from 1994?2004....................................................................................... 53 Figure 5 Mean (? SE) growing season total phosphorous (TP) concentrations measured from the photic zone at the mid-embayment sites in Lake Martin at Stations E1 and D1 of Elkahatchee Creek Embayment and Dennis Creek Embayment, respectively, from 1994?2004 ..................... 54 Figure 6 Mean (? SE) growing season total phosphorous (TP) concentrations measured from the photic zone sampled at the downstream sites in Lake Martin at Stations E2 and D2 of Elkahatchee Creek Embayment and Dennis Creek Embayment, respectively, from 1994?2004 .............. 54 Figure 7 Mean (? SE) growing season total alkalinity measured from the photic zone sampled at the mid-embayment sites in Lake Martin at Stations E1 and D1 of Elkahatchee Creek Embayment and Dennis Creek Embayment, respectively, from 1994?2004.................................. 55 xiii Figure 8 Mean (? SE) growing season total alkalinity measured from the photic zone sampled at the downstream sites in Lake Martin at Stations E2 and D2 of Elkahatchee Creek Embayment and Dennis Creek Embayment, respectively, from 1994?2004.................................. 55 Figure 9 Mean (? SE) growing season specific conductance measured from the photic zone at the mid-embayment sites in Lake Martin at Stations E1 and D1 of Elkahatchee Creek Embayment and Dennis Creek Embayment, respectively, from 1994? 2004 ........................................... 56 Figure 10 Mean (? SE) growing season specific conductance measured from the photic zone sampled at the downstream sites in Lake Martin at Stations E2 and D2 of Elkahatchee Creek Embayment and Dennis Creek Embayment, respectively, from 1994?2004.................................. 56 Figure 11 Mean (? SE) growing season pH measured from the photic zone in Lake Martin at Stations E1 and E2 in Elkahatchee Creek Embayment and Stations D1 and D2 in Dennis Creek Embayment from 1994?2004................................................................................................ 57 Figure 12 Mean (? SE) growing season corrected chlorophyll a values measured from the photic zone in Lake Martin at Stations E1 and E2 in Elkahatchee Creek Embayment and Stations D1 and D2 in Dennis Creek Embayment from 1994?2004........................................................ 57 Figure 13 Mean (? SD) growing season TSI values based on corrected chlorophyll a measured from the photic zone in Lake Martin at Stations E1 and E2 in Elkahatchee Creek Embayment and Stations D1 and D2 in Dennis Creek Embayment from 1994?2004..................... 58 Figure 14 Percent black bass composition represented by spotted and largemouth bass in Lake Martin, Alabama at stations E1 and E2 in Elkahatchee Creek Embayment from 1994?2004 ................................... 59 Figure 15 Percent black bass composition represented by spotted and largemouth bass in Lake Martin, Alabama at stations D1 and D2 in Dennis Creek Embayment from 1994?2004 ....................................... 59 Figure 16 Plot of Weight:Length relations of largemouth bass sampled during the fall in DCE from 1994?2004. Solid and dashed lines represent the regression lines for Pre and Post treatments, respectively....................... 60 xiv Figure 17. Plot of Weight:Length relations of bluegill sunfish sampled during the fall in ECE from 1994?2004. Solid and dashed lines represent the regression lines for Pre and Post treatments, respectively................. 60 Figure 18. Plot of Weight:Length relations of bluegill sunfish sampled in DCE from 1994?2004. Solid and dashed lines represent the regression lines for Pre and Post treatments, respectively ........................................ 61 Figure 19. Plot of Weight:Length relations of redear sunfish sampled in DCE from 1994?2004. Solid and dashed lines represent the regression lines for Pre and Post treatments, respectively ........................................ 62 Figure 20. Plot of Weight:Length relations of blacktail redhorse sampled in ECE from 1994 ? 2004. Solid and dashed lines represent the regression lines for Pre and Post treatments, respectively....................... 63 1 I. INTRODUCTION The Federal Water Pollution Control Act of 1972 addressed mounting concern in America over the widespread degradation of rivers, lakes, and estuaries due to cultural eutrophication. Decades of unabated nutrient enrichment had resulted in undrinkable, foul smelling water prone to summer fish kills and little commercial or recreational value (Carpenter et al. 1998). The obvious solution was to limit the inflow of key nutrients into these systems, namely phosphorous and nitrogen. This movement, termed cultural oligotrophication, aided in the recovery of nuisance conditions by limiting undesirable algal growth and greatly increasing water clarity (Ney 1996). However, the effect was bimodal. Fewer nutrients translated into lower productivity, particularly in reservoirs. By the mid 1990?s, it had become apparent that aesthetically clearer water and fisheries productivity were not complementary (Schupp and Wilson 1993). Establishing a healthy balance between clearer water and fisheries productivity has been complicated by the range of factors influencing trophic dynamics across the continent; these include, hydrology, nutrient input, regional geomorphology and climate (Ney, 1989; Axler et al. 1988.). As stewards, our challenge is to develop strategies for maintaining healthy and sustainable water resources. A sound understanding of reservoir nutrient dynamics is paramount for achieving this goal, both at the worldwide and regional scale. 2 In 1994, Auburn University conducted a limnological study of the Elkahatchee Creek embayment (ECE) on Lake Martin reservoir near Alexander City, Alabama. This study continued through 1997 and 1998. During this time, 8.5 million gallons/day (permitted discharge) of both treated domestic sewage and industrial (primarily textile) effluent were being discharged into Sugar Creek, a tributary of ECE, from the Alexander City Wastewater Treatment Plant (ACWTP). The purpose of the study that began in 1994 was to address rising concern over water quality and environmental condition in ECE. A nearby tributary, Dennis Creek Embayment (DCE), located directly across Lake Martin from ECE, was chosen as a reference site due its proximity, similarity in size and the lack of point source discharge. The objectives of the study were to: 1) determine water quality conditions in ECE and DCE; (2) compare water quality conditions of ECE and DCE with conditions existing in the mainstem of Lake Martin and; (3) document aquatic community (phytoplankton, benthos, and fish) responses to the prevailing environmental conditions during the growing season (Bayne et al. 1995). In June 2001, a pipeline constructed for the ACWTP began diverting effluent out of ECE and into the thalweg on the mainstem of Lake Martin. The purpose of this thesis is to assess the effects of nutrient removal in ECE on water quality, phytoplankton biomass, and fish and benthic invertebrate communities by comparing results obtained in 2002 through 2004 with those reported in 1994, 1997, and 1998. Specifically, I propose to describe: 1) water quality conditions in ECE and DCE prior to and after nutrient diversion; 2) the effects of nutrient availability on phytoplankton biomass; 3) trends in 3 the size, condition, and abundance of fish species, and; 4) changes in the density and diversity of benthic invertebrates. 4 II. LITERATURE REVIEW Political and public perceptions of water quality often focus on water clarity with less emphasis on aquatic communities. As a result, water quality initiatives target nutrient reduction without considering the numerous impacts on reservoir biota. Recent studies have linked nutrient reductions with changes in: zooplankton and algal biomass, fish size and growth; and macroinvertebrate richness and diversity (Chambers et al. 2006; McCormick et al. 2004; Maceina and Bayne 2001; Allen et al. 1999, Hoxmeier and DeVries 1998, DiCenzo et al. 1995, Hendricks et al. 1995, Bayne et al. 1994; Blancher 1984). Eutrophication can describe either the process of nutrient addition or the effects of added nutrients in the aquatic environment (Larkin and Northcote 1969). Cultural eutrophication describes changes resulting from anthropogenic influences. Increased levels of mainly phosphorous and nitrogen, stemming from agricultural, urban, domestic, and industrial effluent can bolster trophic status, sometimes to undesirable levels. Adverse effects of eutrophication in freshwater systems can range from benignly unpleasant tastes or odors to more serious issues such as water treatment inefficiencies, oxygen depletion, and fish kills (Carpenter et al. 1998). Opposite of cultural eutrophication is the anthropogenic abatement and/or removal of nutrients, coined ?cultural oligotrophication? (Ney 1996). Nutrient reductions 5 most commonly occur from construction of upstream impoundments or point-source abatement programs, such as advanced wastewater treatment facilities or sewage diffusion (Bayne et al. 2003, Axler et al. 1988; Gloss et al. 1980). By decreasing nutrient loads, these efforts limit phytoplankton growth and, in turn, limit the primary productivity in the system. Expected outcomes of oligotrophication should include not only clearer water, but decreases in fish biomass and angler satisfaction as well (Yurk and Ney 1989). Phosphorous (P) and nitrogen (N) are often the primary nutrients influencing productivity in freshwater ecosystems. Not all forms of P are readily available to plants and those that are get quickly utilized or bound in lake sediments. Moreover, sediment- bound P does not share the internal recycling efficiency of N (Boyd 2000). Therefore, P is often the least abundant and most limiting nutrient for productivity (Wetzel 1983). An assessment of 19 Alabama reservoirs determined that total phosphorous (TP) explained 49% of the variation in algal biomass (Maceina et al. 1996). The same study concluded that TP was nearly four times as influential as total nitrogen in predicting algal biomass. Canfield and Bachman (1981) reported total phosphorous (TP) to have accounted for 57% of the variation in chlorophyll a among 433 artificial lakes spread across the United States. Miller et al. (1974) found that P was limiting in 35 of 49 lakes with N limiting in the remaining eight. Nitrogen limitation is usually predominant in marine systems but can also exist in eutrophic freshwater systems where P is abundant (Welch 1980). The ratio of total nitrogen to total phosphorous (TN:TP) has been a useful index comparing the relative 6 influence of each nutrient across the trophic spectrum. For example, Welch (1980) reported that oligotrophic lakes had a TN:TP ? 16/1, indicating P limitation. Conversely, eutrophic lakes had a TN:TP < 16/1, indicating N limitation. Smith (1986) used this index to explain the relative dominance of nitrogen fixing blue-green algae to other phytoplankton types. Downing and McCauley (1992) proposed that the TN:TP ratio reflected the source of nutrient input into a system. Nutrient cycling plays a critical role in the recovery time following nutrient abatement. The nitrogen budget in freshwater systems can be maintained at equilibrium by a variety of natural processes, i.e. atmospheric contribution, nitrogen fixation, mineralization, nitrification, and denitrification (Boyd 2000). Following a major nutrient reduction, this equilibrium can be reached in 0?5 years (Anderson et al. 2005). Phosphorous however, lacks the recycling efficiency of N. Periods of anoxia or low dissolved oxygen trigger the release of P by lowering the oxidation-reduction potential at the sediment-water interface. Ultimately, these P inputs are either quickly utilized by plants or re-adsorbed and trapped again in lake sediments (Boyd 2000). S?ndergaard et al. (1999) detected P release at sediment depths down to 25 cm among 265 shallow Danish lakes. Internal P loading can delay improvements in water quality for 10?25 years following nutrient abatement and is greatly influenced by initial loading rate, flushing schedule, depth, and the chemical characteristics of the sediment (Anderson et al. 2005; Jeppesen et al. 2005, S?ndergaard et al 2003; S?ndergaard et al. 1999; Van Der Molen et al. 1998). Edmondson and Lehman (1981) reported that P loading accounted for about 57% of the total P income during a 27-year period in Lake Washington, WA. 7 The relationship between productivity and the nutrient budget is of central importance to limnology. Classical theory asserts a trophic hierarchy whereby energy inputs are first expressed by producers and then flow upward along the foodchain (Elton 1927). Lindemann (1947) further theorized that the efficiency of energy flow followed a predictable pattern throughout natural lake succession; i.e. productivity increased from oligotrophy to eutrophy before eventually declining with lake senescence. Carlson (1977) later proposed the trophic state index (TSI) to quantitatively classify trophic status based on Secchi visibility, chlorophyll a, or TP measurements. The effects of nutrient enrichment on plankton communities have been well documented. Generally, nutrient enrichment leads to: 1) an increasing dominance of Chlorophyta (green algae) and especially Cyanophyta (blue-green algae), and; 2) greater density and lower diversity of algal flora (Bold and Wynne 1985; Edmondson and Lehman 1981; Welch 1980; Schindler et al. 1973; Edmondson 1969). Bayne et al. (1990) recorded a shift in the phytoplankton community from a diatomaceous to a green and blue-green dominance during a 10-year period in West Point Lake, GA. In turn, zooplankton communities reflect this shift from more to less desirable algal species (Porter 1977). General community trends have shown that: 1) zooplankton abundance and biomass increase with trophy; 2) mean length of zooplankton declines with trophy, and; 3) dominance shifts from macrozooplankton to microzooplankton as systems become more productive (Harman et al. 1995; Bayne et al. 1992; Pace 1986; Blancher 1984; Canfield and Watkins 1984; Bays and Crisman 1982; Pace and Orcutt 1981). 8 Presence and distribution of benthic macroinvertebrates along the trophic gradient are regulated by the range of physical and chemical characteristics of each waterbody (Merritt and Cummins 1996). General trends of increased productivity, driven by increased organic enrichment, include: 1) increased standing stock overall; 2) increased densities of Chironomidae, Oligochaeta, Chaoboridae, Gastropoda, and Sphaeriidae; 3) fewer Hexagenia spp., and; 4) a greater composition of Oligochaeta to Chironomidae (Popp and Hoagland 1995; Welch 1980; Dermott et al. 1977; Peterka 1972; J?nasson 1969; Carr and Hiltunen 1965). To date, several researchers have demonstrated a direct relationship between productivity and fish species richness, growth, biomass, and recruitment. McKenna (1995) associated greater diversity with more eutrophic conditions among three New York Finger Lakes. Phosphorous reductions in Lake Erie (1969?1996) led to both: 1) reduced species richness in the shallower and clearer West Basin and; 2) increased species richness in the deeper and more turbid Central Basin (Ludsin et al. 2001). A study of 65 Florida lakes, ranging from oligotrophic to hypereutrophic, found no correlation between species richness and trophic status (Bachman et al. 1996). In 27 Alabama reservoirs, Hendricks et al (1995) found that increased trophic state corresponded to greater average weights of angler-caught black bass (Micropterus spp.). A later study of 32 Alabama reservoirs detected faster growth (mm TL) among largemouth bass and spotted bass as reservoir trophy increased (chlorophyll a ranged from 2 mg/m 3 to 27 mg/m 3 ) (Maceina et al. 1996). Allen et al. (1999) documented a 9 positive linear relationship between trophic status and age-0 largemouth bass among nine Alabama reservoirs (chlorophyll a ranged from 1 mg/m 3 to 27 mg/m 3 ). Crappie (Pomoxis spp.) have also been reported to exhibit faster growth (mm TL) at higher trophic levels. A comparison of three Alabama reservoirs, ranging in trophic status, documented faster crappie growth in the eutrophic reservoir than in the oligomesotrophic (Bayne et al. 1994). This trend was later confirmed in a study of 32 Alabama reservoirs; although, crappie growth did not increase when chlorophyll a levels exceeded 8 mg/m 3 , the lower threshold for being considered eutrophic (Maceina et al. 1996). These trends are not exclusive to piscivores. In 10 Alabama reservoirs, DiCenzo et al. (1996) found that gizzard shad (Dorosoma cepedianum) were more abundant in eutrophic systems; whereas, growth rates and condition were higher in the oligomesotrophic systems. These results confirmed an earlier study of four Alabama reservoirs where gizzard shad abundances increased with trophic status while average total length decreased (Bayne et al. 1994). This trend was also confirmed by Clayton and Maceina (2002) where gizzard shad exhibited faster growth and greater body condition but lower abundances in a reservoir with a lower trophic status. Thus for gizzard shad, increased abundances at higher trophic states may initiate density-dependent regulation which is less pronounced in more oligotrophic systems (Clayton and Maceina 2002). Moreover, with higher abundances, slower growth and smaller sizes gizzard shad make a more ideal forage fish in eutrophic, rather than oligotrophic systems (DiCenzo et al. 1996). 10 Fewer studies have addressed the direct impact of oligotrophication. Nutrient abatement programs in Lake Ontario reduced phosphorous by 25% which coincided with a 50% decrease in forage fish biomass, as well as decreased weights and poorer conditions of stocked salmonids (Great Lakes Fishery Commission 1992). Likewise, a 50% reduction in phosphorous load into Beaver Lake, Arkansas corresponded to a 55% decline in planktivorous fish and a 9% decline in piscivirous fish standing stocks (Ney 1996). During a 12 year study of Smith Mountain Lake, Virginia, Ney (1996) found that as TP declined 79%, annual growth rates of black bass and striped bass (Morone saxatilis) also declined by 12%. Maceina and Bayne (2001) reported that nutrient reductions in West Point Reservoir, Alabama/Georgia resulted in fewer and smaller largemouth bass and smaller, yet more abundant spotted bass (chlorophyll a values declined from 40 ?g/l to 9?17 ?g/l). Greene and Maceina (2000) reported greater densities of age-0 largemouth bass than for spotted bass in eutrophic reservoirs, but greater densities and biomass of the latter in meso- and oligotrophic systems. Buynak et al. (1989) recorded greater catch of largemouth bass in the more eutrophic portions of Cave Run Lake, Kentucky and greater catch of spotted bass in the less productive portions. Reservoir managers make decisions that affect a variety of stakeholders and multi-recreational users. A sound understanding of the nutrient-fishery dynamic is essential for striking the balance between achieving aesthetically clearer water and maximizing fish production. This thesis enhances this understanding by examining nutrient availability, phytoplankton biomass, fisheries productivity, and 11 macroinvertebrate density and diversity as they directly pertain to oligotrophication in a southeastern reservoir. 12 III. MATERIALS AND METHODS Study Sites Lake Martin (Figure 1), a 15,850 ha reservoir was created on the Tallapoosa River system in 1926 to provide power generation, flood control, and recreation for east- central Alabama. The lake has a mean hydraulic retention time of 190 days and has been classified as low-ranged mesotrophic (TSI = 39) by calculating the trophic state index (TSI) based on chlorophyll a concentration (Bayne 1995). Elkahatchee Creek Embayment (ECE) is located south of Alexander City (Figure 2). Until 2001, ECE had received treated municipal waste from Alexander City from a smaller tributary, Sugar Creek. The shoreline is predominantly forested with some shoreline residential development. Station E1 is located at mid-embayment downstream of the Sugar Creek confluence. Station E2 is located at the mouth near the mainstem of Lake Martin (Figure 2). Dennis Creek Embayment (DCE) is located directly across the reservoir (Figure 2) from ECE. Dennis Creek receives no known point-source discharge though does contain some shoreline residential development. Station D1 is located in the upstream portion of DCE and Station D2 is at the mouth near the mainstem of Lake Martin (Figure 2). 13 Water Quality Sampling Temperature and dissolved oxygen, photic zone, Secchi visibility and specific conductance were measured in situ in the water column at each station (Table 1). Water samples were collected using a submersible pump and hose apparatus. Composite samples were pumped from throughout the photic zone and stored in coolers on ice in 2- L Nalgene bottles until arrival at the laboratory for analysis. The following water quality variables were measured in the laboratory: pH, total alkalinity, total hardness, total phosphorous, soluble reactive phosphorous, total nitrogen, nitrate-nitrogen, nitrite- nitrogen, total ammonia nitrogen, turbidity, and total suspended solids (Table 1). Standard analytical methods were followed for all variables and holding times were well within recommended limits (APHA et al. 1998). Water samples were collected monthly during the growing season (April? October) during 1994 and 2004 (Table 2). Only May and June were sampled during 1997. In 1998, June, July, and August were sampled. During 2002, samples were collected monthly from June through August. In 2003, the growing season was represented by April, June, and August. Growing season values were assessed for each water quality variable, divided into either a pre (1994?1998) or post (2002?2004) treatment category, and compared among sites and through time using a split-plot repeated measures analysis of variance (RMANOVA) (Maceina et al. 1994). In some cases the models indicated a significant station/treatment interaction (P < 0.05) signifying that the upstream and mouth stations had responded differently to nutrient 14 diversion over time. For these instances t-tests were used to address the effects at each station separately. Phytoplankton Sampling Composite phytoplankton samples were collected throughout the photic zone as described above (Table 2). Chlorophyll a concentrations (corrected for phaeophyton) were measured in the laboratory according to protocol of the American Public Health Association (APHA 1998). Algal biomass was estimated from chlorophyll a concentrations and comparisons were made between sites and throughout time using RMANOVA. Chlorophyll a data were used to generate trophic scores using the trophic state index (TSI) (Carlson, 1977). Trophic scores were compared between sites and throughout time using RMANOVA. Stepwise regression and correlation analyses identified the variables that were significantly (P < 0.05) related with chlorophyll a concentrations. In some instances, log 10 transformations were calculated to homogenize variances and linearize relationships among variables. Simple and multiple linear regressions were used to describe the nature and magnitude of these relationships. Partial squared correlation coefficients (pr 2 ) addressed the effects of each variable independently, relative to the effects of all others. Multicollinearity was considered problematic when condition index values exceeded 3.00 and when variance inflation factors (VIF) exceeded 1/(1-R 2 ). Fish Sampling Fish were collected twice from the upstream sites in ECE (E1) and DCE (D1). Fish were sampled during July and October 1994, 2002, and 2004 (Table 2). Only July 15 was sampled during 2003. Fish were collected diurnally via boat electrofishing and sinking gill nets. Each site received 60 minutes pedal time of pulsed DC current using a Smith Root? boat electrofisher. Pedal time represented direct fishing effort distributed equally among available habitat types, excluding travel time. Transects were not performed. Sampling also involved a 4 hr soak period using three sinking gill nets (two 1.5 inch and one 2.5 inch bar mesh) checked at 2 hour intervals. Fishes were kept in a livewell until transported to a central location where they were inspected for lesions/abnormalities, identified to species, and measured to the nearest. All fish were weighed to the nearest 0.1 gram with the exception of threadfin shad which were measured only to the nearest mm (TL). Total length was measured to the nearest mm. Fishes were released live when possible near the vicinity of capture. Unidentified species were preserved on ice and transported to the laboratory for identification. Catch per effort (CPE) for electrofishing and gill netting effort was calculated as fish/hour and fish/net, respectively. Student?s t-tests were used to compare CPE between treatments at each site. Relative weight (Wr) assessed fish condition for species having a standard equation (Anderson and Neumann 1996). Data were then pooled for sampling method and Student?s t-tests were used to make comparisons between treatments at each site. For some species, a standard weight equation did not exist. Analysis of covariance (ANCOVA) was then used to detect elevation differences among log 10 transformed weight-length regressions. Significant differences (P < 0.05) between y-intercepts signified that one population was heavier, or better conditioned, at a given length than the 16 other. Condition was also assessed using ANCOVA by assessing differences in slope, rather than elevation, between treatments at each site. Finally, total length (TL) was compared between treatments at each site using Student?s t-tests. Statistical analyses were considered significant at ? = 0.05 and were performed using SAS 9.1.3 (SAS Institute 2005). Macroinvertebrate Sampling Macroinvertebrates were collected during: summer (August) and fall (October) in 1994, summer (July) in 2002, and summer (August) and fall (November) in 2004 (Table 2). Summer samples were compared among all three years. Fall samples were compared only between 1994 and 2004. Macroinvertebrates were sampled using Hester-Dendy multiplate samplers and a petite ponar dredge. One Hester-Dendy sampler placed in DCE during October, 2004, was never recovered. Hester-Dendy multiplate samplers (HD) consisted of nine masonite plates (7.6 cm x 7.6 cm) stacked and separated by 3 mm spacers. These plates imitated the natural reservoir substrate (e.g., rocks and woody debris) that reservoir invertebrates prefer while insuring a standardized colonization medium among stations. The total surface area (excluding edges) of each sampler was about 0.09 m 2 . Three samplers were deployed in each embayment in the vicinity of stations E1 and D1. Each sampler was suspended above the substrate at depths of 1 to 3 m. This depth allowed adequate dissolved oxygen for supporting diverse communities. Samplers were given a 30-day colonization period, after which they were removed and transported on ice in water filled containers. In the laboratory, macroinvertebrates were: removed from the plates, sieved through a U.S. 17 Standard No.30 screen (0.6 mm mesh), preserved in 80% ethyl alcohol, and identified to the lowest practical taxon and counted. Densities were calculated as the mean number of organisms/m 2 . A petite ponar dredge (1.5 m x 1.5 m) was used to collect macroinvertebrates inhabiting the bottom sediment. Three dredge samples were taken from each embayment during July and October in the vicinity of stations E1 and D1. The samples were taken at depths greater than 3 m to ensure representation of anoxic sediments of the profundal zone, if present. Sieving (U.S. No. 30 sieve) and elutriation removed excess silt and debris from macroinvertebrate fauna. Samples were preserved in 80% ethyl alcohol, sorted, and identified to the lowest practical taxon. Densities were calculated as the mean number of organisms/m 2 . Community structure and function were assessed separately for HD and bottom dredge samples. Comparisons were made by season and between treatments using estimates of total density, species richness, diversity, evenness, and dominance by trophic level. Trophic level designations were classified based on feeding behaviors outlined by Merritt and Cummins (1996). Functional feeding groups (FFG) included: filter-collectors (FC), gather-collectors (GC), piercers (PI), predators (P), scrapers (S), and shredders (SH). Densities were compared using Student?s t-tests and Analysis of Variance (ANOVA). Heterogeneity was assessed using the Shannon Diversity Index (H?) and evenness (J?) (Brower and Zar 1984). Student?s t-tests assessed differences between treatments at each site. Variance was computed using the formula: S 2 = ? [ fi(log e fi) 2 ? [(? fi(log e fi) 2 ] / N] / N 2 ; 18 with degrees of freedom calculated as: d.f. = (S 2 1 + S 2 2 ) 2 / [(S 2 1 ) 2 / N 1 ] + [(S 2 2 ) 2 / N 2 ]. Statistical analyses were considered significant at ? = 0.05 and were performed using SAS 9.1.3 (SAS Institute 2005). 19 IV. RESULTS Water Quality Soluble reactive phosphorous (SRP) declined (t 11.1 = -4.96; P < 0.0004) from a mean of 43 ?g?L (? 8; ? SE) prior to nutrient diversion to 2 ?g?L (? 0) following diversion at station E1, the mid-embayment portion of ECE (Figure 3). Station E2, located near the mouth of ECE, had a substantial, though less dramatic and statistically insignificant (t 11.1 = -1.94; P < 0.078), decline in mean SRP from 6 ?g?L (? 3) during the pre to 1 ?g?L (? 0) during the post (Figure 4). Soluble reactive phosphorous did not differ (F 1,1 = 0.01; P < 0.943) among treatment periods in DCE. Total phosphorous (TP) also declined (t 11.2 = -7.59; P < 0.0001) at station E1 from a mean of 116 ?g?L (? 12) during the pre to 27 ?g?L (? 1) following nutrient diversion (Figure 5). At station E2, mean TP declined (t 12.6 = -5.72; P < 0.001) from 47 ?g?L (? 4) during the pre to 21 ?g?L (? 1) during the post (Figure 6) but did not differ (F 1,1 = 0.14; P < 0.769) between treatments in DCE. In ECE, the mean TN:TP ratio rose (t 46 = 4.73; P < 0.0001) from 7.5 during the pre to 15.1 during the post. In DCE, this ratio remained similar (t 45.7 = 1.00; P < 0.325) between treatments and averaged 14.6 during the pre and 16.6 during the post. Total alkalinity (TA) declined (t 13.9 = -5.26; P < 0.0001) at station E1 from a mean of 23 mg/L CaCO 3 (? 1) during the pre to 15 mg/L CaCO 3 (? 1) during the post 20 (Figure 7). At station E2, mean TA declined (t 30 = -2.99; P < 0.0056) from 16 mg/L CaCO 3 (? 1) during the pre to 13 mg/L CaCO 3 (? 1) during the post (Figure 8) but did not differ (F 1,1 = 6.04; P < 0.246) between treatments in DCE. Specific conductance declined at only the mid-embayment ECE station but increased at both DCE stations following nutrient diversion. At E1, mean conductance declined (t 11.4 = -8.37; P < 0.0001) from 119 ?mhos/cm (? 8) to 52 ?mhos/cm (? 1) following diversion (Figure 9). At station D1, mean conductance increased (t 25 = 2.29; P < 0.0309) from 40 ?mhos/cm (? 1) to 43 ?mhos/cm (? 1) following diversion. At station D2, mean conductance increased (t 28.3 = 4.38; P < 0.0001) from 41 ?mhos/cm (? 1.3) to 52 ?mhos/cm (? 2.2) following diversion (Figure 10). Mean pH decreased in ECE (F 1,1 = 164.10; P < 0.0496) from 7.39 (? 0.16) to 6.65 (? 0.07) following diversion but did not differ in DCE (F 1,1 = 17.74; P < 0.1484) (Figure 11). Hardness, total suspended solids, turbidity, compensation point, total nitrogen, nitrite-nitrogen, nitrate-nitrogen, ammonia-nitrogen, and Secchi disk visibility showed no significant (P > 0.05) changes following nutrient diversion in either embayment. Phytoplankton Chlorophyll a (CHLA) concentrations were higher (F 1,1 = 404.69; P < 0.0316) in ECE before effluent diversion and averaged 13 mg?m 3 (? 2; ? SE), as compared to the post average of 9 mg?m 3 (? 2) (Figure 12). Trophic state declined from pre (TSI = 54) to post (TSI = 51) but remained statistically similar (F 1,1 = 20.77; P < 0.1375) (Figure 13). Prior to nutrient diversion in ECE, pH, total nitrogen, turbidity, total suspended solids, compensation point, and Secchi visibility were all significant (P < 0.05) correlates 21 with CHLA (Table 3). Of these, total nitrogen (TN) was the best predictor of CHLA and explained 78% of the variability in the model: log 10 CHLA = 0.189 + 5.22 (log 10 TN). Following nutrient diversion in ECE, TP, total suspended solids, compensation point, and Secchi visibility were all significant (P < 0.05) correlates with CHLA (Table 4). Of these, TP was the best predictor and explained 18% of the variability in the model: log 10 CHLA =0.860 +6.32(log 10 TP). In DCE, CHLA concentrations remained similar (F 1,1 = 0.42; P < 0.6336) between treatments and averaged 5 mg?m 3 (? 1) before nutrient diversion compared with 5 mg?m 3 (? 1) after diversion (Figure 12). Trophic state rose slightly between pre (Mean TSI = 46) and post treatments (Mean TSI = 47), but remained statistically similar (F 15,41 = 0.79; P < 0.6845) (Figure 13). Soluble reactive phosphorous, TIN, total suspended solids, and compensation point were all significant correlates with CHLA (Table 5). However, none of these variables were effective predictors of CHLA. This was most likely attributable to the low degree of variation in CHLA in DCE throughout the study period (Figure 12). Fish A total of 5,271 fishes weighing 759.7 kg representing 7 families and 25 species were collected at both sites from 1994 to 2004 (Appendix I and II). Dominant taxa included: Micropterus punctulatus, M. salmoides, Lepomis macrochirus, L. microlophus, and Pomoxis nigromaculatus (Centrarchidae); Dorosoma cepedianum and D. petenense (Clupeidae); Moxostoma poecilurum (Catostomidae); Cyprinella venusta and Notropis 22 texanus (Cyprinidae); and Morone chrysops (Moronidae) which accounted for 92% of total catch and 70% of total biomass. Black bass accounted for 11% of total catch and 22% of total biomass overall (gill net and electrofishing). The percentage of black bass in ECE represented by spotted bass (Micropterus punctulatus) increased from 40% to 60% (Figure 14) following nutrient diversion; whereas, spotted bass comprised the majority of black bass throughout the study in DCE (Figure 15). In ECE, no changes in condition or CPE (gill net or electrofishing) of largemouth bass (Micropterus salmoides) were detected (P > 0.49). Relative weight decreased (t 130 = -4.08; P < 0.0001) from a mean of 91.8 (? 8.1; ? SD) to 83.1 (? 8.6) following oligotrophication. Mean TL increased (t 36.5 = 3.88; P < 0.004) from 220 mm (? 137; ? SD) prior to diversion to 318 mm (? 76) following diversion. In DCE, largemouth bass expressed higher condition (ANCOVA; t = 2.96; P < 0.0056) during the post for fish collected in the fall (Figure 16), but not for summer caught fish (P < 0.85). Mean TL increased (t 66 = 2.79; P < 0.0069) from 222 mm (? 103) during the pre to 297 mm (? 117) during the post. No differences (P > 0.05) in CPE or Wr were detected. In ECE, mean CPE of spotted bass collected by electrofishing increased, though statistically similar (t 5 = 2.51; P > 0.054), from 7 fish/hr (? 2) during the pre to 24 fish/hr (? 2) during the post. No differences (P > 0.05) in spotted bass length or condition were detected. However, mean Wr declined (t 133 = -4.90; P < 0.0001) from 96.8 (? 9.3) during the pre to 86.3 (? 8.9) during the post. 23 In DCE, mean CPE of spotted bass collected by gill nets declined (t 5 = -3.70; P < 0.014) following nutrient diversion from 18 fish/net during the pre to 4 fish/net during the post. In addition, mean Wr declined (t 59 = -2.30; P < 0.025) from 91.5 (+ 5.9) during the pre to 87.0 (+ 8.5) during the post. Mean TL also declined (t 181 = -3.62; P <0.0004) from 280 mm (? 109) during the pre to 212 mm (? 120) during the post. Bluegill sunfish (Lepomis machrochirus) (79%) and redear sunfish (Lepomis microlophus) (11%) comprised the majority of sunfish collected. Less dominant species included green sunfish (Lepomis cyanellus) (5%), longear sunfish (Lepomis megalotis) (3%), warmouth sunfish (Lepomis gulosus) (1%), and redbreast sunfish (Lepomis auritus) (1%). Differences in CPE among treatments were not evident at either site for any sunfish species regardless of season or gear type, but low sample sizes likely hampered these analyses (t <1.61; df range: 1-5; P > 0.06). In ECE, bluegill sunfish were expressed poorer condition (ANCOVA; t = -2.21; P < 0.028) following nutrient removal for fish collected during the fall (Figure 17), but not for summer caught fish (P > 0.90). Relative weight also declined (t 817 = -14.37; P < 0.0001) from a mean of 87.9 (? 9.1) during the pre to 77.9 (? 9.6) during the post. Mean TL also declined (t 645 = -9.72; P < 0.0001) from 121 mm (? 25) to 104 mm (? 31) following nutrient removal and the percentage of fish longer than 89 mm TL decreased from 95% to 67%, respectively. In DCE, bluegill sunfish were better conditioned (ANCOVA; t = 3.92; P < 0.0001) following diversion (Figure 18). Mean Wr also increased (t 284 = 6.00; P < 0.0001) from 78.6 (? 5.7) during the pre to 82.7 (? 8.7) during the post. For fish collected 24 in the fall, mean TL increased (t 442 = 2.22; P < 0.027) from 94 mm (? 32) during the pre to 102 mm (? 34) during the post. Redear sunfish expressed higher condition (ANCOVA; t = 2.45; P < 0.0155) in DCE during the post (Figure 19); however, no differences (P > 0.05) in TL or Wr were detected. In ECE, only three individuals were recorded prior to nutrient diversion preventing statistical comparisons across treatments. Black crappie (Pomoxis nigromaculatus) represented the majority of crappie with only one white crappie (Pomoxis annularis) sampled in DCE during fall of 2002. No differences (P > 0.07) in CPE or condition were detected at either site. In ECE, mean Wr declined (t 170 = -3.69; P < 0.0003) from 85.4 (? 6.5) during the pre to 78.3 (? 6.9) during the post. Total length remained similar (P > 0.40). In DCE, mean TL decreased (t 150 = - 12.37; P < 0.0001) from 255 mm (? 23) during the pre to 189 mm (? 51) during the post. No differences (P > 0.12) in Wr were detected. Gizzard shad (Dorosoma cepedianum) and threadfin shad (Dorosoma petenense) accounted for 18% of total catch and 5% of total biomass (gill net and electrofishing). Threadfin shad comprised 87% of shad abundance, but gizzard shad accounted for 95% of shad biomass. Gizzard shad Wr increased (t 62 = 2.25; P < 0.0283) in DCE following nutrient diversion from a mean of 80.8 (? 13.6) to 87.4 (? 11.3), but remained similar (P > 0.57) in ECE between treatments. No differences (P > 0.45) in CPE (gill net or electrofishing), condition, or TL were detected among treatments at either site. This was likely caused by low sample sizes during the pre, particularly in ECE where only nine individuals were recorded. 25 In ECE, mean length of threadfin shad increased (t 475 = 9.78; P < 0.0001) from 56 mm (? 6) during the pre to 63 mm (? 6) during the post. In DCE, mean length of threadfin shad decreased (t 321 = -6.99; P < 0.0001) from 63 mm (? 8) during the pre to 55 mm (? 13) during the post. No differences (P > 0.05) in CPE (gill net or electrofishing), condition, or weight were detected. Blacktail redhorse (Moxostoma poecilurum) comprised 90% of suckers with the remaining 10% comprised of spotted suckers (Minytrema melanops). Blacktail redhorse condition remained similar (P > 0.18) at both sites. Though in ECE, mean total weight decreased (ANCOVA; t = - 5.16; P < 0.0001) from 272.7 g during the pre to 251.7 g during the post (Figure 20). In addition, mean TL declined (t 45.3 = -11.12; P < 0.0001) from 368 mm (? 27) during the pre to 293 mm (? 59) during the post. In DCE, blacktail redhorse CPE (gill net) declined (t 4 = -5.58; P< 0.0051) from a mean of 17 fish/net (? 1) during the pre to 6 fish/net (? 1) during the post. Mean TL also declined (t 248 = -6.80; P < 0.0001) from 341 mm (? 43) during the pre to 286 mm (? 51) during the post. Total weight remained similar (P > 0.42) between treatments. Blacktail shiner (Cyprinella venusta) (52%) and weed shiner (Notropis texanus) (44%) comprised the majority of minnow abundance. Less dominant species included bullhead minnow (Pimephales vigilax) (3%) and pugnose minnow (Opsopoeodus emiliae) (1%). No differences (P > 0.05) in CPE (gill net or electrofishing), condition, or total weight were detected for these species. In DCE, mean TL of blacktail shiners increased (t 48.6 = 5.87; P < 0.0001) from 59 mm (? 5) during the pre to 78 mm (? 23) during the post. In ECE conversely, mean TL 26 decreased (t 91 = -2.31; P < 0.023) from 89 mm (? 17) during the pre to 76 mm (? 19) during the post. The same trend was detected in weed shiners. Mean TL in DCE increased (t 80 = 3.40; P < 0.001) from 40 mm (? 4) during the pre to 58 mm (? 7) during the post and decreased (t 34.9 = -5.17; P < 0.0001) in ECE from 58 mm (? 2) during the pre to 53 mm (? 5) during the post. White bass (Morone chyrsops) accounted for 88% of Moronidae catch (gill net and electrofishing) with hybrid bass (Morone chrysops x saxatalis) comprising the rest. No differences (P > 0.05) in CPE (gill net or electrofishing), condition, Wr, or TL were detected for either species at either site. Notably however, white bass CPE (gill net) in ECE decreased from 30 fish/net, during fall of 1994, to an average of 6 fish/net during the post. Hybrid bass were only sampled during fall, 2002, but occurred in both embayments. Twelve hybrid bass were collected in ECE compared with only three in DCE. Channel catfish (Ictalurus punctatus) represented 80% of the catfish sampled with flathead catfish (Pylodictus olivaris) comprising the remainder. Low sample sizes during the pre treatment prevented pre versus post comparisons among sites (Appendix I,II). Low sample sizes also prevented statistical analysis of mobile logperch (Percina kathae), yellow perch (Perca flavescens), and common carp (Cyrpinus carpio) (Appendix I,II). Mobile logperch were collected at low densities but occurred at both sites and during both treatments. Yellow perch was only collected once, during summer, 2004 in DCE. Common carp was only collected during the post treatment but occurred in both embayments equally. 27 Benthic Invertebrates from Hester-Dendy Multiplate Samplers A total of 10,022 benthic invertebrates representing 13 Orders, 21 Families, and 46 taxa were collected from Hester-Dendy multiplate samplers at both sites during the study period. About 77% of the benthic invertebrates sampled consisted of chironomid midges (Diptera), caddisfly larvae (Trichoptera), and aquatic worms (Oligochaeta). Mean density for samples collected during the fall season in ECE remained similar following diversion (t 4 = -0.19; P > 0.85). However, diversity dramatically declined (t 1,272 = 18.2; P < 0.0001), as did evenness (Table 6). Following diversion, mean densities of chironomid midges and trichopteran larvae, mainly Cyrnellus fraternus, declined 78% and 98%, respectively. Samples collected during the post treamtent were dominated by cladoceran zooplankton, which accounted for 91% of total density. Filter- collectors remained the predominant trophic group following diversion but data collected in 2004 represented five trophic levels, as opposed to only three collected in 1994. Densities collected during the summer season also remained similar (ANOVA: F 2,6 = 1.5; P > 0.29) following diversion; although total density in 2002 appeared much higher than in 1994 or 2004 (Table 7). Increased densities of Glyptotendipes spp., Ablabesmyia spp., and Naidid oligochaetes were responsible for the elevated mean. Most notable overall was a 96% decrease in mean density of Cyrnellus fraternus between 1994 and 2004. Species richness, diversity, and evenness increased steadily over the ten-year period. In 1994, only three trophic levels were represented and were dominated by filter- collectors and gather-collectors. All five trophic levels were collected in 2004. 28 Only two of three multiplate samplers were recovered from DCE during the fall season in 2004. As a result, any disparity between years, such as the declines in species richness, diversity, and evenness, may have been directly attributable to the lost sampling effort and not due to environmental factors (Table 8). Nonetheless, total density significantly increased (t 685 = 24.6; P < 0.0001) following diversion, primarily due to a 94% increase in cladoceran zooplankton. Mean density of Chironomidae declined about 82%. Trophic levels collected during both years were dominated by filter-collectors. In DCE, mean density for data collected during the summer season remained similar (ANOVA: F 2,6 = 0.74; P > 0.51); while; species richness, diversity, and evenness increased steadily over the ten year period (Table 9). Diversity values in 2004 were significantly higher (t 871 = - 5.45; P < 0.0001) than those recorded in 1994. Insects of the Family Ceratopogonidae and Polycentropodidae declined 18% and 82%, respectively, while those belonging to Naididae increased 96%. Three trophic levels were collected in 1994 and were dominated by filter-collectors and shredders. All five trophic levels were represented in 2004 and were dominated by filter-collectors and gather-collectors. Benthic Invertebrates from Bottom Dredge Samples A total of 2,357 benthic invertebrates representing 9 Orders, 15 Families, and 30 taxa were collected from bottom dredge samples from both sites during the study. The phantom midge, Chaoborus spp. (Chaoboridae), comprised about 79% of benthic invertebrates sampled. Other dominant taxa groups consisted of chironomid midges (Diptera), and aquatic worms (Oligochaeta), which made up an additional 15% of abundance. 29 In ECE, mean density for data collected during the fall season remained similar (t 4 = 1.56; P < 0.195) following diversion (Table 10). Species richness also remained similar but diversity and evenness significantly increased (t 360 = -4.13; P < 0.0001) during the post. Mean densities of Chaoboridae, Chironomidae, Copepoda, and Sphaeriidae all decreased following diversion while those of Tubificidae and cladoceran zooplankton increased. Predators, mainly Chaoborus spp., were the dominant trophic group in all samples but substantial increases in shredders and gather-collectors were apparent in the 2004 samples. In ECE, mean density was significantly higher (ANOVA: F 2,6 = 19.68; P < 0.0023) during 2002 for data collected during the summer season, but remained similar between 1994 and 2004 (Table 11). Data collected in 2002 had the highest densities of most taxa, particularly Chaoboridae and Sphaeriidae. While taxa richness was highest during 2002, diversity and evenness significantly declined (t 191 = 7.72; P < 0.0001) about 86% between 1994 and 2004. Insects of the Family Chironomidae were most abundant in 1994 but had decreased by 90% in 2004. Predators remained the predominant trophic level following diversion. In DCE, mean density for data collected during the fall season remained similar (t 4 = -1.18; P < 0.303) following diversion (Table 12). Five fewer taxa were collected during the post but diversity remained similar (t 213 = 0.50; P < 0.618), as did evenness. Most notable were the increased densities of Chaoboridae and Tubificidae following diversion. Predators remained the predominant trophic group; although gather-collectors and shredders were substantially higher during the post. 30 In DCE, mean density was higher following diversion for data collected during the summer season (Table 13), though the trend was not statistically significant (ANOVA: F 2,6 = 3.51; P < 0.303). Taxa richness was highest during 1994 and decreased by almost half in 2004. Diversity and evenness steadily declined following diversion. Mean densities of Chaoboridae recorded in 2002 (5,541/m 2 ) and 2004 (5,733/m 2 ) were twice those recorded in 1994 (2,326/m 2 ). This was also reflected by the increased dominance of predators during the post (Table 13). 31 V. DISCUSSION Nutrient removal resulted in sharp declines in TP and SRP in ECE. Consonant with these declines was a 31% decrease in phytoplankton biomass as well as significant changes in condition and abundance of fish species and density and diversity of macroinvertebrates. However, CHLA values by the end of the study remained nearly twice those recorded in DCE, perhaps indicating the persistence of key nutrients resulting from decades of internal loading and continued non-point source (urban) influence. Meanwhile, elevated hypolimnetic conductivity at the mouth of DCE following diversion suggested the possible encroachment of diverted sewage. However, water quality measurements and phytoplankton biomass remained static following diversion. Biotic communities in DCE displayed mixed results to the potential enrichment. Elkahatchee Creek Embayment At E1, TP and SRP declined 77% and 96% following nutrient diversion resulting in a 31% decline in phytoplankton biomass. The influence of TN on CHLA during the pre (78% of the variation in CHLA) and the lack of change in TN levels between treatments suggested that ECE was nitrogen limited during the pre due to high phosphorous loading. This was also supported by the low TN:TP ratio (7.5). Downing and McCauley (1992) reported similarly low ratios associated for: cattle manure seepage (8.9), feedlot runoff (6.4) and sewage (5.3). Following diversion, the TN:TP ratio (15.1) 32 increased to levels more similar to those recorded in DCE (14.6?16.6) and similar to average river water (18.9) (Ebise and Inoue 1991). Moreover, the strong relationship between TN and CHLA decoupled leaving TP as the most significant regressor. Despite the decrease in phytoplankton biomass following nutrient diversion, CHLA in ECE during the post was still slightly higher than that recorded in DCE. I believe that urban runoff and internal P loading from hypolimnetic sediments were primarily responsible for the elevated CHLA during the post-diversion. Lung and Larson (1995) attributed internal P loading for supplying up to one-third of daily TP loads in Lake Pepin, MS. Edmondson and Lehman (1981) reported over half of the total P income into Lake Washington, WA over a 27-year period was attributed to internal P loading from the sediment. Although sediment P was not sampled, specific conductance values (used as a means of tracking effluent) collected from the bottom were highly correlated with SRP (R = 0.727; P < 0.0001) during the pre-diversion. Mean conductivity in ECE remained slightly higher than detected at D1 following diversion which suggested that decades of point-source discharge, along with urban runoff, were still affecting the embayment. Consonant with the decline in CHLA were significant changes in the fish and invertebrate communities. Black bass dominance shifted to favor spotted bass following nutrient removal. Buynak et al. (1989) reported a longitudinal shift in black bass dominance as spotted bass comprised the majority of black bass in the oligotrophic portions of Cave Run Lake, Kentucky. Greene and Maceina (2000) found greater densities of age-0 spotted bass in mesotrophic and oligotrophic reservoirs and greater 33 densities of age-0 largemouth bass in eutrophic reservoirs. Maceina and Bayne (2001) found fewer largemouth bass with lower relative weights and more abundant, yet smaller, spotted bass following oligotrophication in West Point Reservoir, GA. Spotted bass became more abundant, but less robust following oligotrophication. A study of 10 Alabama reservoirs detected lower Wr in spotted bass sampled from oligo- mesotrophic systems as compared with eutrophic systems (DiCenzo et al. 1995). A later study of 32 Alabama impoundments reported poorer body conditions of spotted bass in oligomesotrophic reservoirs, compared with fish sampled from eutrophic systems (Maceina et al. 1996). Several studies have also documented slower growth of spotted bass in lakes of lower trophic status, though this trend was not detected in our results (Maceina and Bayne 2001; DiCenzo et al. 1995). Largemouth bass were longer yet less robust following oligotrophication. Maceina and Bayne (2001) reported that crappie expressed higher relative weights and exhibited faster growth in more eutrophic reservoirs. Bayne et al. (1994) reported smaller largemouth bass in the mesotrophic Lake Martin (TSI = 39) compared with the moderately eutrophic Lake Eufaula (TSI = 57), Alabama. However, growth rates (mm TL) in Lake Martin remained similar to those sampled in the highly eutrophic Weiss Lake (TSI = 65) and moderately eutrophic Jones Bluff Lake (TSI = 55). Black crappie comprised 19% of the fish community by abundance prior to diversion, compared with < 2% following oligotrophication (gill net and electrofishing). Hoxmeier and DeVries (1998) found that crappie contributed a larger proportion to the fish community in oligotrophic reservoirs. In addition, black crappie were less robust 34 following oligotrophication. Maceina et al. (1996) reported that crappie were generally better conditioned and exhibited faster crappie growth in more eutrophic reservoirs. Bayne et al. (1994) also detected faster crappie growth in eutrophic systems. Bluegill sunfish were significantly shorter and poorer conditioned during the post. A study of four Alabama impoundments found that sunfish were less abundant in a mid- range mesotrophic system (TSI: 39) than in two eutrophic systems (TSI: 55, 57) (Bayne et al. 1994). Low sample sizes prevented any statistical comparisons of gizzard shad. However, based on previous studies, I would have expected less abundant, yet larger, more robust, and faster growing fish with decreasing productivity (Clayton and Maceina, 2002; Bremigan and Stein 2001; Allen et al. 1999; Hoxmeier and DeVries, 1998; DiCenzo et al. 1996; McQueen et al. 1998; and Bayne et al. 1994). This was similar to the response I observed for threadfin shad, which were longer and had higher average weights following oligotrophication. Macroinvertebrate community structure shifted after diversion, but densities remained static following oligotrophication in ECE, except for the spike observed in summer, 2002 (Table 7). Dramatic declines in Cyrnellus fraternus, Glytpotendipes spp., and Dictrotendipes spp. corresponded with significantly higher diversity and evenness in Hester-Dendy summer collections following diversion in ECE. The dominance of several tolerant species is a key signature of degraded systems (Karr and Chu 1999). Improved water quality allows broader partitioning of the resource. The presence of all five trophic 35 levels following diversion as opposed to only three in 1994 was consistent with this aspect of recovery. Bottom dredge samples from ECE revealed more variable results. Density remained static between treatments, except for the summer 2002 spike (Table 11). Summer diversity steadily declined throughout the study periods and was significantly lower in 2004. This was likely due to the overwhelming dominance (97% of total density) of Chaoborus spp. in 2004. The ability of a taxon to grossly out compete all others is usually indicative of a stressed system and is not suggestive of a stable recovery. Fall collections from ECE revealed that Chaoborus spp. and Coelotanypus spp. comprised 90% of total density prior to nutrient diversion (see Appendix Table 15). Popp and Hoagland (1995) also reported a 90% composition of these two taxa following increased sediment enrichment in Pawnee Reservoir, NE. Following oligotrophication, diversity and evenness during the fall significantly increased in ECE, which corresponded with: 1) lower densities of Chaoborus spp.; 2) a more even distribution of chironomid taxa; 3) increased densities of Tubificidae, and; 3) the addition of a fourth trophic level (shredders) (Table 10). Notably however, the presence of shredders was due solely to the appearance of Chironomus spp., which together with Tubificidae was also a pollution tolerant species and also reported at high densities by Popp and Hoagland (1995). Bottom dredge diversity steadily declined during the summer. These results suggest an unstable recovery in the profundal sediments of ECE. 36 Dennis Creek Embayment Phytoplankton biomass and TSI in DCE remained similar between treatment periods. Moreover, no changes in key nutrients, such as TP, SRP, TN, or the TN:TP ratio were detected. Specific conductance however increased significantly at both stations in DCE following nutrient diversion. These values were particularly high at D1, located nearest the point of sewage diffusion immediately between the mouths of DCE and ECE (Figure 2). During periods of chemical stratification, the plume of effluent-enriched water would have entered into DCE via the hypolimnion. I speculate that the substantial increase in conductivity recorded in DCE following effluent diversion, particularly in the mouth, may be evidence of sewage encroachment. However, water quality results provided no evidence of nutrient transfer to the photic zone. Although trophic state and nutrient levels remained static following diversion, certain fish species displayed the characteristics indicative of eutrophication. For example, as system productivity increases, the black bass dynamic tends to favor largemouth bass (Buynak et al. 1989). Certainly, both species have been shown to exhibit faster growth and better condition in eutrophic systems (DiCenzo et al, 1995; Greene and Maceina, 2000; Allen et al., 1999), but abundance and biomass of largemouth bass can be much higher than for spotted bass (Greene and Maceina, 2000; Buynak et al., 1989). At the end of this study, spotted bass remained the dominant black bass. However, they became less abundant, shorter, and expressed lower relative weights following nutrient diversion; whereas, largemouth bass were longer and had higher condition. These data 37 are inconsistent with existing research with respect to Wr of spotted bass. However, my data suggest that the black bass dynamic may be shifting to favor the production of largemouth bass. Lepomis spp. also responded in a manner consistent with eutrophication. Bluegill sunfish were longer and more robust following diversion. No differences in abundance were evident, but bluegill sunfish have been shown to increase in numbers at higher trophic levels (McKenna 2001; Bayne et al. 1994). Redear sunfish also expressed higher condition following diversion. While no references were found pertaining to redear condition or growth, a study of 65 Florida lakes revealed increased standing crop with increasing trophic status (Bachmann et al. 1996). Not all species responded in a manner consistent with nutrient enrichment. Black crappie were significantly shorter following diversion, though no differences in condition or catch were recorded. Shorter and less robust crappie have been associated with lower productivity in several southeastern reservoirs (Maceina et al. 1996; Bayne et al. 1994). However, size difference among treatments was likely more related to variable recruitment than trophic status. Blacktail shiners and weed shiners were longer following oligotrophication. No studies were found comparing cyprinid metrics across the trophic gradient in southeastern reservoirs. Persson et al. (1991) reported increased size of cyprinids among 13 Swedish lakes as trophic status increased from low to medium productivity. The larger sizes were reported as a size-refuge response to the increased importance of piscivory in medium productive lakes. 38 Gizzard shad normally respond to eutrophication with increased abundance, slower growth, and lower Wr (Clayton and Maceina, 2002; DiCenzo et al. 1996). I found no differences in abundance, but these fish were significantly more robust following diversion. Threadfin shad were shorter and average weight declined following diversion; a reversal of the trend detected in ECE. Both shad species are omnivorous filter feeders relying primarily on zooplankton, phytoplankton, and organic detritus (Baker and Schmitz 1971; Cramer and Marzolf 1970; Kutkhun 1957). Also, both species typically show increases in abundance and biomass as trophic status increases in southeastern reservoirs (Allen et al. 2000; DiCenzo et al. 1996; Harman et al. 1995). However, my data suggest that the relative abundance of these shad species to one another may shift along the trophic gradient. Although no differences in CPE were detected, the percentage of total shad represented by threadfin shad increased 69% during the post treatment. I found no studies that explicitly compared the relative abundance of these shad species to one another across the trophic gradient. However, data taken from Bayne et al. (1994) corroborated this trend and showed the percentage of total shad represented by threadfin shad increased 45% as trophic level increased among four Alabama reservoirs. Indeed, a number of mechanisms might explain such a trend: 1) increased vulnerability of gizzard shad at higher trophic status; 2) capture efficiency in more turbid waters; or; 3) shifts in food and/or habitat availability. If this is a legitimate response, rather than a spurious coincidence, then the relative proportion of threadfin shad could provide a useful and sensitive index for assessing trophic changes in southeastern reservoirs. 39 Macroinvertebrates inhabiting the littoral zone of DCE did not respond in a manner consistent with nutrient enrichment. Density from Hester-Dendy samplers remained static, richness and evenness increased, and diversity rose by 16% following nutrient diversion. In 1994, Polycentripodidae larvae, mostly Cyrnellus fraternus, comprised about 44% of total density. By 2004, density declined to less than 2%. These trends were more indicative of stress recovery and were especially similar to those recorded in ECE. Generally, nutrient enrichment results in greater densities, lower diversity and evenness, and increasing dominance of select taxa (Popp and Hoagland 1995; Welch 1980; Dermott et al. 1977; Peterka 1972; J?nasson 1969; Carr and Hiltunen 1965). Macroinvertebrates inhabiting the bottom sediments displayed variable responses between seasons that were both consistent and inconsistent with nutrient enrichment. Density appeared higher during both seasons following diversion but the increases were not statistically significant (P > 0.05). The ability of ANOVA and t-tests to detect these changes was likely undermined by high variability among replicates. Seasonable variability is often pronounced in lentic macroinvertebrate sampling (Hergenrader and Lessig 1980; Parrish and Wilhm 1978). Fall diversity remained similar among treatments, but structural changes between pre and post displayed characteristics indicative of nutrient enrichment. For example, five taxa present in 1994 were absent in 2004, including the burrowing mayfly (Hexagenia spp.). Carr and Hiltunen (1965) reported a 99% decline in Hexagenia spp. resulting from organic pollution during a 31-year study of western Lake Erie. Also, densities of Tubificidae and Chaoboridae were dramatically 40 higher in DCE following nutrient diversion. These species are generally associated with organic enrichment during eutrophication (Merritt and Cummins 1996). Following diversion, summer diversity declined in response to the increasing dominance of Chaoborus spp., which comprised 98% of total density in 2004. However, Tubificidae and Chironomidae densities sharply declined during the post. These taxa are also associated with organic enrichment and would not be expected to decline with increasing eutrophication. Moreover, sharp increases in Chaoborus spp. were also recorded in 2004 in ECE. Summary Oligotrophication in ECE produced measurable responses within the biotic community. Nutrient reduction resulted in lower algal biomass, a shift in the black bass community, decreased catch and condition of sunfish and crappie, and altered structure and function of littoral macroinvertebrate communities. Despite a 31% decrease following diversion, phytoplankton biomass remained nearly twice that recorded in DCE. This delayed recovery is likely attributed to non-point source (urban) discharge and internal phosphorous loading. Results from other studies suggest that these effects may persist for many years. Circumstantial evidence suggested that diverted sewage may have infiltrated the DCE. However, despite increased conductivity in DCE, no changes conclusive to nutrient enrichment were found with respect to: nutrients, algal biomass, or macroinvertebrate 41 assemblages. Though, Centrarchid fish (black bass and sunfish) responded in a manner consistent with eutrophication. Most research assessing faunal changes across the trophic gradient compare systems of disparate trophic status. However, ECE and DCE displayed only mild fluctuations in trophic status between treatments. As provided in this thesis, both pronounced and subtle changes in the nutrient regime can produce a measurable response within the aquatic community. Understanding the nature and magnitude of this response is key for managing the balance between aesthetically clearer water and fish production. 42 Table 1. Analytical methods used for assessing water quality variables in ECE and DCE in Lake Martin, Alabama. Variable Method Reference In Situ Temperature thermistor APHA, 1998 Dissolved oxygen membrane electrode APHA, 1998 pH glass electrode APHA, 1998 Specific conductance conductivity cell APHA, 1998 Visibility Secchi disk APHA, 1998 Euphotic zone determination submarine photometer Lind, 1985 Laboratory Analyses Total suspended solids vacuum filtration APHA, 1998 Turbidity nephelometric APHA, 1998 Alkalinity potentiometric titration APHA, 1998 Total ammonia (NH 3 -N) phenate method APHA, 1998 Nitrite (NO 2 -N) diazotizing method APHA, 1998 Nitrate (NO 3 -N) cadmium reduction APHA, 1998 Total nitrogen persulfate method APHA, 1998 Total phosphorous persulfate method, ascorbic acid APHA, 1998 Soluble reactive phosphorous ascorbic acid APHA, 1998 Total hardness EDTA titrimetric method APHA, 1998 Chlorophyll a Spectrophotometric APHA, 1998 43 Table 2. Sampling schedule for water quality, phytoplankton, invertebrates, and fish during the 10?year study period in ECE and DCE in Lake Martin, Alabama. Year Water Quality/ Phytoplankton Invertebrate Fish 1994 April ? October August; October July; October 1997 May ? June ? ? Pre 1998 June ? August ? ? 2002 June ? August July July; October 2003 April; June; August ... July Post 2004 April - October August; November July; October Table 3. Summary of Pearson Correlation Coefficients for chlorophyll a concentrations in ECE before nutrient removal. The water quality variables that were significantly correlated are reported in the table along with sample size (n) and P values. Variable r n P < pH 0.881 24 0.0001 Total Nitrogen 0.884 14 0.0001 Turbidity 0.770 24 0.0001 Total Suspended Solids 0.770 24 0.0001 Secchi Disk Visibility -0.859 12 0.0003 Compensation Point -0.726 22 0.0001 Secchi Disk Visibility -0.774 24 0.0001 44 Table 4. Summary of Pearson Correlation Coefficients for chlorophyll a concentrations in ECE following nutrient removal. The water quality variables that were significantly correlated are reported in the table along with sample size (n) and P values. Variable r n P < Total Phosphorous 0.488 34 0.0034 Total Suspended Solids 0.526 34 0.0014 Compensation Point -0.372 34 0.0302 Secchi Disk Visibility -0.424 34 0.0123 Table 5. Summary of Pearson Correlation Coefficients for chlorophyll a concentrations in DCE. The water quality variables that were significantly correlated are reported in the table along with sample size (n) and P values. Variable r n P < Total Nitrogen -0.401 44 0.0070 Total Ammonia-Nitrogen -0.403 54 0.0025 Nitrate-Nitrogen -0.298 54 0.0285 45 Table 6. Mean values (? SE) of benthic invertebrate assemblage variables from Hester- Dendy multiplate samples taken during the fall season in ECE in Lake Martin, Alabama. Community Variables 1994 2004 Species Richness 16 13 Shannon Diversity 1.61 (? 0.002) 0.45 (? 0.002) Evenness 0.581 0.17 Density (# / m 2 ) Total Density 2674 (? 225) 2822 (? 751) Baetidae 7 ? Caenidae ? 4 Ceratopogonidae 15 ? Chironomidae 896 200 Cladocera 44 2570 Coenagrionidae 7 ? Copepoda Hydroptilidae ? 4 Naididae 7 Oligochaeta 37 ? Pleuroceridae ? 4 Polycentropodidae 1274 26 Filter-Collector 6000 (? 1155) 2707 (? 774) Gather-Collector 455 (? 58) 18 (? 3) Shredder ? 26 (? 15) Predator 107 (? 10) ? Piercer 11 (0) Scraper ? 17 (? 5) 46 Table 7. Mean values (? SE) of benthic invertebrate assemblage variables from Hester- Dendy multiplate samples taken during the summer season in ECE in Lake Martin, Alabama. Community Variables 1994 2002 2004 Species Richness 12 15 22 Shannon Diversity 1.04 (? 0.001) 1.88 (? 0.0006) 2.17 (? 0.001) Evenness 0.42 0.69 0.70 Density (# / m 2 ) Total Density 7593 (? 4568) 10633 (? 1173) 3519 (? 1785) Caenidae ? 30 26 Chironomidae 4267 7433 2111 Cladocera 22 356 748 Coenagrionidae 7 30 7 Copepoda 44 ? 15 Heptageniidae ? 15 Hydroptilidae 30 4 Naididae 74 1156 400 Nematoda ? 207 ? Ostracoda 7 59 44 Polycentropodidae 3140 1304 126 Sphaeriidae ? ? 22 Turbellaria .. 30 ? Filter-Collector 4163 (? 2737) 6352 (? 255) 1881 (? 1199) Gather-Collector 3052 (? 1779) 1863 (? 425) 570 (? 144) Shredder ? ? 133 (? 86) Predator 296 (? 80) 922 (? 695) 237 (? 132) Piercer ? 89 (0) 11 (0) Scraper ? 15 (2) 47 Table 8. Mean values (? SE) of benthic invertebrate assemblage variables from Hester- Dendy multiplate samples taken during the fall season in DCE in Lake Martin, Alabama. Community Variables 1994 2004** Species Richness 17 7 Shannon Diversity 1.78 (? 0.003) 0.216 (? 0.001) Evenness 0.630 0.111 Density (# / m 2 ) Total Density 1648 (? 460) 3611 (? 1900) Caenidae 4 ? Chironomidae 630 112 Cladocera 200 3472 Harpacticoida 7 ? Naididae 59 6 Oligochaeta 7 ? Ostracoda 15 Polycentropodidae ? 17 Filter-Collector 1140 (? 268) 3594 (? 946) Gather-Collector 293 (? 98) 11 (0) Shredder 100 (? 83) 11 (0) ** Only 2/3 multiplate samplers were recovered from study site 48 Table 9. Mean values (? SE) of benthic invertebrate assemblage variables from Hester- Dendy multiplate samples taken during the summer season in DCE in Lake Martin, Alabama. Community Variables 1994 2002 2004 Species Richness 18 19 21 Shannon Diversity 1.86 (? 0.002) 1.99 (? 0.003) 2.21 (? 0.001) Evenness 0.64 0.69 0.73 Density (# / m 2 ) Total Density 2200 (? 178) 1563 (? 520) 2059 (? 388) Caenidae ? ? 63 Ceratopogonidae 22 4 Chironomidae 1007 352 767 Cladocera 44 78 222 Coenagrionidae 7 7 22 Corduliidae ? 4 ? Copepoda 11 11 Heptageniidae ? 4 4 Hydroptilidae 59 48 Naididae 22 537 541 Oligochaeta 74 ? ? Ostracoda 22 93 337 Polycentropodidae 963 81 30 Sphaeriidae ? ? 4 Turbellaria 311 ? Filter-Collector 1189 (? 151) 311 (? 33) 559 (? 70) Gather-Collector 622 (? 57) 641 (? 263) 660 (? 85) Shredder 1561 (-) 333 (? 185) 37 (? 20) Predator 256 (? 32) ? 44 (? 22) Piercer ? 59 (? 43) 22 (-) Scraper ? 11 (-) 52 (? 37) 49 Table 10. Mean values (? SE) of benthic invertebrate assemblage variables from dredge samples taken during the fall season in ECE in Lake Martin, Alabama. Community Variables 1994 2004 Species Richness 9 9 Shannon Diversity 1.02 (? 0.004) 1.42 (? 0.006) Evenness 0.46 0.65 Density (# / m 2 ) Total Density 3585 (? 488) 2593 (? 411) Chaoboridae 2044 1289 Chironomidae 1274 963 Cladocera ? 44 Copepoda 178 Corbiculidae 15 ? Oligochaeta Sphaeriidae 59 15 Tubificidae ? 237 Filter-Collector 111 (? 54) 59 (? 15) Gather-Collector 74 (? 29) 237 (? 115) Shredder ? 667 (? 245) Predator 3259 (? 565) 1585 (? 660) 50 Table 11. Mean values (? SE) of benthic invertebrate assemblage variables from dredge samples taken during the summer season in ECE in Lake Martin, Alabama. Community Variables 1994 2002 2004 Species Richness 6 8 4 Shannon Diversity 1.04 (? 0.009) 0.42 (? 0.003) 0.15 (? 0.005) Evenness 0.58 0.20 0.11 Density (# / m 2 ) Total Density 1778 (? 92) 5852 (? 752) 2622 (? 362) Chaoboridae 1259 5304 2548 Chironomidae 311 15 30 Cladocera ? 44 ? Copepoda 74 15 Leptoceridae ? ? Nematoda 15 Sphaeriidae 133 3111 44 Tubificidae ? 134 ? Filter-Collector 133 (? 51) 356 (? 205) 44 (? 0) Gather-Collector ? 163 (? 78) ? Shredder ? 11 (-) 11 (-) Predator 1526 (? 150) 5304 (? 417) 2563 (? 376) 51 Table 12. Mean values (? SE) of benthic invertebrate assemblage variables from dredge samples taken during the fall season in DCE in Lake Martin, Alabama. Community Variables 1994 2004 Species Richness 12 7 Shannon Diversity 1.26 (? 0.015) 1.16 (? 0.012) Evenness 0.51 0.60 Density (# / m 2 ) Total Density 1318 (? 449) 2074 (? 456) Chaoboridae 756 1244 Chironomidae 430 370 Cladocera 15 30 Copepoda 44 Corbiculidae 15 ? Ephemeridae Tubificidae 74 385 Filter-Collector 89 (? 0) 89 (-) Gather-Collector 133 (? 36) 385 (? 97) Shredder 67 (? 18) 267 (? 77) Predator 1096 (? 365) 1348 (? 418) 52 Table 13. Mean values (? SE) of benthic invertebrate assemblage variables from dredge samples taken during the summer season in DCE in Lake Martin, Alabama. Community Variables 1994 2002 2004 Species Richness 11 4 6 Shannon Diversity 1.34 (? 0.009) 0.38 (? 0.002) 0.11 (? 0.002) Evenness 0.58 0.27 0.06 Density (# / m 2 ) Total Density 3111 (? 353) 6133 (? 975) 5852 (? 1143) Ceratopogonidae 30 ? ? Chaoboridae 2030 5541 5733 Chironomidae 370 474 74 Copepoda 341 400 15 Naididae 119 ? ? Nematoda 30 Sphaeriidae 15 ? ? Tubificidae 178 118 30 Filter-Collector 11 (-) ? 11 (-) Gather-Collector 296 (? 29) 518 (? 121) 89 (-) Shredder 156 (? 54) ? 11 (-) Predator 2326 (? 602) 5541 (? 1033) 5733 (? 1153) Figure 1. Map of Lake Martin, Alabama and study area. 53 54 Alexander City Sugar Creek E1 E2 Elkahatchee Creek D2 D1 Dennis Creek Tallapoosa River Figure 2. A map of the upstream portion of Lake Martin, Alabama, also containing the study sites, Elkahatchee Cree, Embayment (ECE) and Dennis Creek Embayment (DCE). 55 Figure 3. Mean (? SE) growing season soluble reactive phosphorous (SRP) concentrations measured from the photic zone at the mid- embayment sites in Lake Martin at Stations E1 and D1 of ECE and DCE, respectively, from 1994 ? 2004. Year 1994 1997 1998 2002 2003 2004 SR P ( ? g/ L) 0 30 60 90 ECE DCE Year 1994 1997 1998 2002 2003 2004 SR P ( ? g/ L) 0 5 10 15 20 ECE DCE Figure 4. Mean (? SE) growing season soluble reactive phosphorous (SRP) concentrations measured from the photic zone at the downstream sites in Lake Martin at Stations E2 and D2 of ECE and DCE, respectively, from 1994 ? 2004. Year 1994 1997 1998 2002 2003 2004 TP ( ? g/ L) 0 36 72 108 144 180 ECE DCE Figure 5. Mean (? SE) growing season total phosphorous (TP) concentrations measured from the photic zone at the mid-embayment sites in Lake Martin at Stations E1 and D1 of ECE and DCE, respectively, from 1994 ? 2004. Year 1994 1997 1998 2002 2003 2004 TP ( ? g/L ) 0 20 40 60 80 ECE DCE Figure 6. Mean (? SE) growing season total phosphorous (TP) concentrations measured from the photic zone sampled at the downstream sites in Lake Martin at Stations E2 and D2 of ECE and DCE, from 1994 ? 2004. 56 Year 1994 1997 1998 2002 2003 2004 To tal Alk a li n ity (mg /L CaCO 3 ) 0 4 8 12 16 20 24 28 32 ECE DCE Figure 7. Mean (? SE) growing season total alkalinity measured from the photic zone sampled at the mid-embayment sites in Lake Martin at Stations E1 and D1 of ECE and DCE, respectively, from 1994 ? 2004. Figure 8. Mean (? SE) growing season total alkalinity measured from the photic zone sampled at the downstream sites in Lake Martin at Stations E2 and D2 of ECE and DCE, respectively, from 1994 ? 2004. Year 1994 1997 1998 2002 2003 2004 To tal Alk a li n ity (mg /L CaCO 3 ) 0 4 8 12 16 20 ECE DCE 57 Year 1994 1997 1998 2002 2003 2004 Con ductan ce (? mh os/c m) 0 30 60 90 120 150 180 ECE DCE Figure 9. Mean (? SE) growing season specific conductance measured from the photic zone at the mid-embayment sites in Lake Martin at Stations E1 and D1 of ECE and DCE, respectively, from 1994 ? 2004. ECE 58 Figure 10. Mean (? SE) growing season specific conductance measured from the photic zone sampled at the downstream sites in Lake Martin at Stations E2 and D2 of ECE and DCE, respectively, from 1994 ? 2004. Year 1994 1997 1998 2002 2003 2004 Cond uctan ce (? mh os/ c m ) 0 20 40 60 80 100 DCE 59 Figure 11. Mean (? SE) growing season pH measured from the photic zone in Lake Martin at Stations E1 and E2 in ECE and Stations D1 and D2 in DCE from 1994?2004. Treatment Pre Post pH 6.0 6.5 7.0 7.5 8.0 8.5 ECE DCE Treatment Pre Post Corrected Chloro phy ll ? ( m g? m3 ) 0 4 8 12 16 ECE DCE Figure 12. Mean (? SE) growing season corrected chlorophyll a values measured from the photic zone in Lake Martin at Stations E1 and E2 in ECE and Stations D1 and D2 in DCE from 1994?2004. Year 1994 1997 1998 2002 2003 2004 Troph ic S tat e Ind ex 30 40 50 60 70 ECE DCE Hypereutrophic Eutrophic Mesotrophic Oligotrophic Figure 13. Mean (? SE) growing season TSI values based on corrected chlorophyll a measured from the photic zone in Lake Martin at Stations E1 and E2 in ECE and Stations D1 and D2 in DCE from 1994?2004. 60 61 DCE 0% 20% 40% 60% 80% 100% 1994 2002 2003 2004 Year B l a c k B a s s C o mp o s it io n Largemouth Bass Spotted Bass Figure 14. Percent black bass composition represented by spotted and largemouth bass in Lake Martin, Alabama at station E2 in ECE from 1994?2004. ECE 0% 20% 40% 60% 80% 100% 1994 2002 2003 2004 Year Bl a c k Ba s s Co m p o s i t i o n Largemouth Bass Spotted Bass Figure 15. Percent black bass composition represented by spotted and largemouth bass in Lake Martin, Alabama at station D2 in DCE from 1994?2004. Log 10 TL 2.5 2.6 2.7 2.8 Log 10 WT 3 4 Post Pre Pre: Log10 Y = -5.30 + 3.16 * Log10 WT r2 = 0.99 Post: Log10 Y = -5.91 + 3.40 * Log10 WT r2 = 0.99 Log 10 TL 2.5 2.6 2.7 2.8 Log 10 WT 3 4 Post Pre Pre: Log10 Y = -5.30 + 3.16 * Log10 WT r2 = 0.99 Post: Log10 Y = -5.91 + 3.40 * Log10 WT r2 = 0.99 Figure 16. Plot of Weight:Length relations of largemouth bass sampled during the fall in DCE from 1994?2004. Solid and dashed lines represent the regression lines for Pre and Post treatments, respectively. 62 Log 10 TL 1.8 2.0 2.2 2.4 Lo g 10 WT 0 1 2 3 Pre Post Pre: Log10 Y = -5.04 + 3.12 * Log10 WT r2 = 0.98 Post: Log10 Y = -4.87 + 3.01 * Log10 WT r2 = 0.98 Figure 17. Plot of Weight:Length relations of bluegill sunfish sampled during the fall in ECE from 1994?2004. Solid and dashed lines represent the regression lines for Pre and Post treatments, respectively. Log 10 TL 2.0 2.1 2.2 2.3 2.4 2.5 2.6 Log 10 WT 2 3 Pre Post Pre: Log10 Y = -4.94 + 3.06 * Log10 WT r2 = 0.99 Post: Log10 Y = -5.23 + 3.21 * Log10 WT r2 = 0.98 Figure 18. Plot of Weight:Length relations of bluegill sunfish sampled in DCE from 1994?2004. Solid and dashed lines represent the regression lines for Pre and Post treatments, respectively. 63 Log 10 TL 2.20 2.25 2.30 2.35 2.40 2.45 2.50 Lo g 10 WT 2 3 Post Pre Pre: Log10 Y = -4.89 + 3.03 * Log10 WT r2 = 0.99 Post: Log10 Y = -5.20 + 3.17 * Log10 WT r2 = 0.99 Figure 19. Plot of Weight:Length relations of redear sunfish sampled in DCE from 1994?2004. Solid and dashed lines represent the regression lines for Pre and Post treatments, respectively. 64 Log 10 TL 2.50 2.52 2.54 2.56 2.58 2.60 2.62 Lo g 10 WT 2.5 2.6 2.7 2.8 2.9 3.0 Post Pre Pre: Log10 Y = -5.69 + 3.29 * Log10 WT r2 = 0.94 Post: Log10 Y = -5.14 + 3.06 * Log10 WT r2 = 0.99 Figure 20. Plot of Weight:Length relations of blacktail redhorse sampled in ECE from 1994?2004. 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Family Species Dennis Creek Embayment Elkahatchee Creek Embayment Pre Post Pre Post Clupeidae Gizzard Shad (Dorosoma cepedianum) ? 17 4 16 Threadfin Shad (Dorosoma petenense) 171 190 95 382 Cyprinidae Blacktail Shiner (Cyprinella venusta) 8 77 13 80 Common Carp (Cyprinus carpio) ? 10 ? 8 Weed Shiner (Notropis texanus) 2 80 11 58 Pugnose Minnow (Opsopoeodus emiliae) ? 1 ? 3 Bullhead Minnow (Pimephales vigilax) ? 3 ? 2 Catostomidae Spotted Sucker (Minytrema melanops) ? 3 1 ? Blacktail Redhorse (Moxostoma poecilurum) 12 179 16 212 Ictaluridae Channel Catfish (Ictalurus punctatus) ? 33 1 11 Flathead Catfish (Pylodictis olivaris) ? 2 2 13 Moronidae White Bass (Morone chrysops) ? ? ? ? Hybrid Striped Bass (M. chrysops x saxatilis) ? 1 ? ? Centrarchidae Redbreast Sunfish (Lepomis auritus) 7 13 1 6 Green Sunfish (Lepomis cyanellus) 5 17 4 73 Warmouth (Lepomis gulosus) 3 3 3 33 Bluegill (Lepomis macrochirus) 166 553 297 707 Longear Sunfish (Lepomis megalotis) ? 1 5 52 Redear Sunfish (Lepomis microlophus) 19 115 3 86 Spotted Bass (Micropterus punctulatus) 19 107 16 136 Largemouth Bass (Micropterus salmoides) 18 35 24 85 White Crappie (Pomoxis annularis) ? 1 ? ? Black Crappie (Pomoxis nigromaculatus) ? 150 68 30 Percidae Mobile Logperch (Percina kathae) 4 9 2 5 Yellow Perch (Perca flavescens) ? 1 ? ? 79 Appendix II. Summary of fish species and sample sizes collected by sinking gill nets in ECE and DCE in Lake Martin, Alabama during the sampling period 1994?2004. Dennis Creek Elkahatchee Creek Family Species Embayment Embayment Pre Post Pre Post Clupeidae Gizzard Shad (Dorosoma cepedianum) 11 36 5 39 Threadfin Shad (Dorosoma petenense) ? ? ? ? Cyprinidae ? ? ? ? Blacktail Shiner (Cyprinella venusta) ? ? ? ? Common Carp (Cyprinus carpio) ? ? ? 1 Weed Shiner (Notropis texanus) ? ? ? ? Pugnose Minnow (Opsopoeodus emiliae) ? ? ? ? Bullhead Minnow (Pimephales vigilax) ? ? ? ? Catostomidae Spotted Sucker (Minytrema melanops) ? 8 5 20 Blacktail Redhorse (Moxostoma poecilurum) 34 25 7 40 Ictaluridae Channel Catfish (Ictalurus punctatus) 4 10 ? 15 Flathead Catfish (Pylodictis olivaris) 1 ? ? 2 Moronidae White Bass (Morone chrysops) 26 25 30 25 Hybrid Striped Bass (M. chrysops x saxatilis) ? 2 ? 12 Centrarchidae Redbreast Sunfish (Lepomis auritus) ? ? ? ? Green Sunfish (Lepomis cyanellus) ? ? ? ? Warmouth (Lepomis gulosus) ? ? ? ? Bluegill (Lepomis macrochirus) 1 ? ? ? Longear Sunfish (Lepomis megalotis) ? ? ? ? Redear Sunfish (Lepomis microlophus) 1 5 ? 5 Spotted Bass (Micropterus punctulatus) 36 21 6 19 Largemouth Bass (Micropterus salmoides) 11 4 8 29 White Crappie (Pomoxis annularis) ? ? ? ? Black Crappie (Pomoxis nigromaculatus) 43 17 64 10 Percidae Mobile Logperch (Percina kathae) ? ? ? ? Yellow Perch (Perca flavescens) ? ? ? ? 80 Appendix III. Benthic invertebrates collected from Hester-Dendy multiplate samplers during the summer season in 1994 in the ECE in Lake Martin, Alabama. Replicates Mean Taxon FFG A B C Total / 0.09 m 2 AQUATIC INSECTS Diptera Chironomidae Ablabesmyia spp. P 27 14 37 78 289 Dicrotendipes spp. GC 541 98 88 727 2693 Dicrotendipes tritomus FC 8 0 0 8 30 Glyptotendipes spp. FC 162 19 81 262 970 Nanocladius spp. GC 0 0 7 7 26 Nilothauma spp. GC 27 0 0 27 100 Tribelos spp. GC 27 9 7 43 159 Odonata Coenagrionidae Argia spp. P 0 0 2 2 7 Trichoptera Polycentropodidae Cyrnellus fraternus FC 664 56 72 792 2933 Unidentified larvae FC 32 16 8 56 207 Trichoptera Unidentified pupae 8 0 0 8 30 OTHER AQUATIC INVERTEBRATES Cladocera FC 0 6 0 6 22 Copepoda Cyclopoida 8 4 0 12 44 Oligochaeta Naididae GC 0 4 16 20 74 Ostracoda 0 2 0 2 7 Total No. of Organisms 1504 228 318 2050 Total No. Taxa / Sampler 7 9 8 Density (No./m 2 ) / Sampler 16711 2533 3533 7593 81 Appendix IV. Benthic invertebrates collected from Hester-Dendy multiplate samplers during the fall season in 1994 in the ECE in Lake Martin, Alabama. Replicate Mean Taxon FFG A B C Total / 0.09 m 2 AQUATIC INSECTS Diptera Ceratopogonidae P 0 4 0 4 15 Chironomidae Ablabesmyia annulata gp. P 3 0 0 3 11 Ablabesmyia spp. P 5 7 8 20 74 Asheum beckae GC 8 17 0 25 93 Cladopelma spp. GC 3 0 0 3 11 Dicrotendipes spp. GC 5 3 6 14 52 Glyptotendipes spp. FC 33 41 26 100 370 Nilothauma spp. GC 3 0 0 3 11 Pseudochironomus spp. GC 3 14 14 31 115 Tribelos spp. GC 15 14 6 35 130 Unidentified pupae 0 2 2 4 15 Ephemeroptera Baetidae Centroptilum spp. GC 2 0 0 2 7 Odonata Coenagrionidae Argia spp. P 0 2 2 7 Trichoptera Polycentropodidae Cyrnellus fraternus FC 50 48 78 176 652 Polycentropus spp. FC 0 2 0 2 7 Unidentified larvae FC 42 34 66 142 526 Trichoptera Unidentified larvae 8 2 14 24 89 OTHER AQUATIC INVERTEBRATES Cladocera FC 90 14 16 120 444 Copepoda Cyclopoida 2 0 0 2 7 Oligochaeta Oligochaeta GC 2 2 6 10 37 Total No. of Organisms 274 204 244 722 Total No. Taxa / Sampler 14 11 10 Density (No./m 2 ) / Sampler 3044 2267 2711 2674 82 Appendix V. Benthic invertebrates collected from Hester-Dendy multiplate samplers during the summer season of 1994 in the DCE in Lake Martin, Alabama. Replicates Mean Taxon FFG A B C Total / 0.09 m 2 AQUATIC INSECTS Diptera Ceratopogonidae P 0 4 2 6 22 Chironomidae Ablabesmyia spp. P 17 11 24 52 193 Asheum beckae GC 14 8 12 34 126 Chironomini GC 3 0 0 3 11 Dicrotendipes spp. GC 20 46 33 99 367 Glyptotendipes spp. FC 20 8 21 49 181 Labrundinia spp. P 3 0 0 3 11 Nanocladius spp. GC 3 0 0 3 11 Parachironomus arcuatus gp. P 3 3 0 6 22 Polypedilum spp. SH 14 0 0 14 52 Pseudochironomus spp. GC 3 0 0 3 11 Unidentified pupae 4 0 2 6 22 Odonata Coenagrionidae Argia spp. P 0 0 2 2 7 Trichoptera Polycentropodidae Cyrnellus fraternus FC 66 32 62 160 593 Unidentified larvae FC 28 32 40 100 370 Trichoptera Unidentified pupae 0 2 4 6 22 Unidentified larvae 0 0 4 4 15 (table continues) 83 Appendix V. (continued) Replicates Mean Taxon FFG A B C Total / 0.09 m 2 OTHER AQUATIC INVERTEBRATES Cladocera FC 4 8 0 12 44 Oligochaeta Naididae Pristina spp. GC 6 0 0 6 22 Oligochaeta GC 4 12 4 20 74 Ostracoda 4 0 2 6 22 Total No. of Organisms 216 166 212 594 Total No. Taxa / Sampler 15 9 10 Density (No./m 2 ) / Sampler 2400 1844 2356 2200 84 Appendix VI. Benthic invertebrates collected from Hester-Dendy multiplate samplers during the fall season in 1994 in the DCE in Lake Martin, Alabama. Replicate Mean Taxon FFG A B C Total / 0.09 m 2 AQUATIC INSECTS Diptera Chironomidae Ablabesmyia spp. P 15 1 0 16 59 Asheum beckae GC 15 3 4 22 81 Chironomini GC 0 0 1 1 4 Dicrotendipes spp. GC 12 10 7 29 107 Glyptotendipes spp. FC 41 26 21 88 326 Labrundinia spp. P 3 0 0 3 11 Nilothauma spp. GC 0 1 0 1 4 Pseudochironomus spp. GC 0 0 1 1 4 Tribelos spp. GC 3 0 1 4 15 Unidentified pupae 2 2 1 5 19 Ephemeroptera Caenidae Caenis spp. GC 0 1 0 1 4 Odonata Coenagrionidae Argia spp. P 6 1 1 8 30 Trichoptera Polycentropodidae Cyrnellus fraternus FC 40 37 36 113 419 Unidentified larvae FC 16 17 20 53 196 Trichoptera Unidentified larvae 10 7 5 22 81 OTHER AQUATIC INVERTEBRATES Cladocera FC 54 0 0 54 200 Copepoda GC 2 0 0 2 7 Oligochaeta Naididae Pristina spp. GC 12 2 2 16 59 Oligochaeta GC 0 0 2 2 7 Ostracoda 0 2 2 4 15 Total No. of Organisms 231 110 104 445 Total No. Taxa / Sampler 11 10 11 Density (No./m 2 ) / Sampler 2567 1222 1156 1648 85 Appendix VII. Benthic invertebrates collected from Hester-Dendy multiplate samplers during the summer season in 2002 in the ECE in Lake Martin, Alabama. Replicates Mean Taxon FFG A B C Total / 0.09 m 2 AQUATIC INSECTS Diptera Chironomidae Ablabesmyia spp. P 16 200 17 233 863 Dicrotendipes simpsoni FC 114 67 51 232 859 Dicrotendipes spp. GC 0 100 34 134 496 Glyptotendipes spp. FC 244 467 324 1035 3833 Nanocladius alternanthera GC 16 0 0 16 59 Polypedilum beckae 81 167 68 316 1170 Tribelos fuscicorne GC 16 0 17 33 122 Unidentified pupae 8 0 0 8 30 Ephemeroptera Caenidae Caenis spp. GC 0 8 0 8 30 Odonata Coenagrionidae Argia spp. P 8 0 0 8 30 Trichoptera Hydroptilidae Unidentified larvae PI 8 0 0 8 30 Polycentropodidae Cyrnellus fraternus FC 120 40 192 352 1304 OTHER AQUATIC INVERTEBRATES Cladocera FC 48 16 32 96 356 Nematoda 40 0 16 56 207 Oligochaeta Naididae GC 64 72 176 312 1156 Ostracoda 8 8 0 16 59 Turbellaria P 0 8 0 8 30 Total No. of Organisms 791 1153 927 Total No. Taxa / Sampler 13 10 9 Density (No./m 2 ) / Sampler 8789 12811 10300 10633 86 Appendix VIII. Benthic invertebrates collected from Hester-Dendy multiplate samplers during the summer season in 2002 in the DCE in Lake Martin, Alabama. Replicates Mean Taxon FFG A B C Total / 0.09 m 2 AQUATIC INSECTS Diptera Chironomidae Ablabesmyia annulata gp. P 0 0 1 1 4 Ablabesmyia spp. P 2 0 0 2 7 Dicrotendipes simpsoni FC 2 3 4 9 33 Dicrotendipes spp. GC 4 5 0 9 33 Glyptotendipes spp. FC 19 7 4 30 111 Nanocladius alternanthera GC 2 0 2 4 15 Polypedilum beckae 17 1 8 26 96 Tanytarsus sp. C FC 2 0 0 2 7 Tribelos fuscicorne GC 9 3 0 12 44 Ephemeroptera Heptageniidae Unidentified nymph SC 1 0 0 1 4 Odonata Coenagrionidae Argia spp. P 1 1 0 2 7 Corduliidae Unidentified nymph P 0 1 0 1 4 Trichoptera Hydroptilidae Ochrotrichia spp. PI 0 4 0 4 15 Unidentified larvae PI 1 9 2 12 44 Polycentropodidae Cyrnellus fraternus FC 6 7 9 22 81 OTHER AQUATIC INVERTEBRATES Cladocera FC 5 8 8 21 78 Copepoda GC 0 3 0 3 11 Nematoda 7 0 0 7 26 Oligochaeta Naididae GC 90 24 31 145 537 Ostracoda 4 21 0 25 93 Turbellaria P 59 20 5 84 311 Total No. of Organisms 231 117 74 422 Total No. Taxa / Sampler 16 13 10 Density (No./m 2 ) / Sampler 2567 1300 822 1563 87 Appendix IX. Benthic invertebrates collected from Hester-Dendy multiplate samplers during the summer season in 2004 in the ECE in Lake Martin, Alabama. Replicates Mean Taxon FFG A B C Total / 0.09 m 2 AQUATIC INSECTS Diptera Chironomidae Ablabesmyia annulata gp. P 13 1 40 54 200 Ablabesmyia spp. P 4 0 0 4 15 Chironomus spp. SH 8 0 27 35 130 Cricotopus/Orthocladius GC 4 0 0 4 15 Dicrotendipes neomodestus FC 4 1 0 5 19 Dicrotendipes simpsoni FC 21 1 54 76 281 Dicrotendipes spp. GC 4 0 13 17 63 Endochironomus spp. SH 0 1 0 1 4 Glyptotendipes spp. FC 25 22 134 181 670 Labrundinia spp. P 4 0 0 4 15 Polypedilum beckae 29 6 121 156 578 Pseudochironomus spp. GC 4 1 0 5 19 Tanytarsus sp. P FC 4 0 0 4 15 Tribelos fuscicorne GC 0 0 13 13 48 Unidentified pupae 5 2 4 11 41 Ephemeroptera Caenidae Caenis spp. GC 1 0 6 7 26 Heptageniidae Stenacron spp. SC 2 1 1 4 15 Odonata Coenagrionidae Argia spp. P 0 0 2 2 7 Trichoptera Hydroptilidae Unidentified larvae PI 0 1 0 1 4 Polycentropodidae Cyrnellus fraternus FC 10 5 19 34 126 (appendix continues) 88 Appendix IX. Benthic invertebrates, continued. Replicates Mean Taxon FFG A B C Total / 0.09 m 2 OTHER AQUATIC INVERTEBRATES Cladocera FC 26 2 174 202 748 Copepoda 0 0 4 4 15 Oligochaeta Naididae GC 22 76 10 108 400 Ostracoda 0 0 12 12 44 Pelecypoda Sphaeriidae FC 5 0 1 6 22 Total No. of Organisms 195 120 635 950 Total No. Taxa / Sampler 16 12 15 Density (No./m 2 ) / Sampler 2167 1333 7056 3519 89 Appendix X. Benthic invertebrates collected from Hester-Dendy multiplate samplers during the fall season in 2004 in the ECE in Lake Martin, Alabama. Replicates Mean Taxon FFG A B C Total /0.09 m 2 AQUATIC INSECTS Diptera Chironomidae Chironomus spp. SH 1 1 4 6 22 Dicrotendipes neomodestus FC 1 0 1 2 7 Dicrotendipes simpsoni FC 0 0 2 2 7 Dicrotendipes spp. GC 0 1 1 2 7 Endochironomus spp. SH 0 0 1 1 4 Glyptotendipes spp. FC 11 3 12 26 96 Polypedilum beckae 5 1 8 14 52 Unidentified pupae 1 0 0 1 4 Ephemeroptera Caenidae Brachycercus spp. GC 1 0 0 1 4 Heptageniidae Stenacron spp. SC 0 1 1 2 7 Trichoptera Hydroptilidae Unidentified larvae PI 1 0 0 1 4 Polycentropodidae Cyrnellus fraternus FC 2 2 3 7 26 OTHER AQUATIC INVERTEBRATES Cladocera FC 347 245 102 694 2570 Gastropoda Pleuroceridae Elimia spp. SC 0 0 1 1 4 Oligochaeta Naididae GC 1 0 1 2 7 Total No. of Organisms 371 254 137 762 Total No. Taxa / Sampler 9 7 11 Density (No./m 2 ) / Sampler 4122 2822 1522 2822 90 Appendix XI. Benthic invertebrates collected from Hester-Dendy multiplate samplers during the summer season in 2004 in the DCE in Lake Martin, Alabama. Replicates Mean Taxon FFG A B C Total / 0.09 m 2 AQUATIC INSECTS Diptera Ceratopogonidae Unidentified larvae P 1 0 0 1 4 Chironomidae Ablabesmyia annulata gp. P 0 1 4 5 19 Chironomus spp. SH 2 2 0 4 15 Dicrotendipes neomodestus FC 0 2 4 6 22 Dicrotendipes simpsoni FC 11 3 4 18 67 Endochironomus spp. SH 2 2 0 4 15 Glyptotendipes spp. FC 10 9 39 58 215 Nanocladius (N.) alternanthera 8 0 0 8 30 Parachironomus carinatus GC 0 1 12 13 48 Polypedilum beckae 19 7 54 80 296 Polypedilum spp. SH 2 0 0 2 7 Tribelos fuscicorne GC 2 0 0 2 7 Unidentified pupae 1 3 3 7 26 Ephemeroptera Caenidae Caenis spp. GC 2 14 1 17 63 Heptageniidae Stenacron spp. SC 0 0 1 1 4 Odonata Coenagrionidae Argia spp. P 1 1 4 6 22 Trichoptera Hydroptilidae Ochrotrichia spp. SC 11 0 2 13 48 Unidentified larvae PI 2 0 0 2 7 Polycentropodidae Cyrnellus fraternus FC 2 2 4 8 30 91 Appendix XI. Benthic invertebrates, continued. Replicates Mean Taxon FFG A B C Total / 0.09 m 2 OTHER AQUATIC INVERTEBRATES Cladocera FC 16 35 9 60 222 Copepoda 0 0 3 3 11 Oligochaeta Naididae GC 70 33 43 146 541 Ostracoda 45 2 44 91 337 Pelecypoda Sphaeriidae FC 0 0 1 1 4 Total No. of Organisms 207 117 232 556 Total No. Taxa / Sampler 15 15 16 Density (No./m 2 ) / Sampler 2300 1300 2578 2059 92 Appendix XII. Benthic invertebrates collected from Hester-Dendy multiplate samplers during the fall season in 2004 in the DCE in Lake Martin, Alabama. Replicates Mean Taxon FFG A B Total / 0.09 m 2 AQUATIC INSECTS Diptera Chironomidae Dicrotendipes neomodestus FC 10 0 10 56 Endochironomus spp. SH 0 1 1 6 Glyptotendipes spp. FC 7 2 9 50 Polypedilum beckae 1 0 1 6 Trichoptera Polycentropodidae Cyrnellus fraternus FC 2 1 3 17 OTHER AQUATIC INVERTEBRATES Cladocera Daphnidae FC 475 150 625 3472 Oligochaeta Naididae GC 1 0 1 6 Total No. of Organisms 496 154 650 Total No. Taxa / Sampler 6 4 7 Density (No./m 2 ) / Sampler 5511 1711 7222 4815 93 Appendix XIII. Benthic invertebrates collected from dredge samples during the summer season in 1994 in the ECE in Lake Martin, Alabama. Replicates Mean Taxon FFG A B C Total / 0.0225 m 2 AQUATIC INSECTS Diptera Chaoboridae Chaoborus spp. P 26 26 30 82 1215 Chaoborus spp. pupae 1 2 0 3 44 Chironomidae Coelotanypus spp. P 4 3 5 12 178 Parachironomus arcuatus gp. P 0 1 0 1 15 Tanypus spp. P 2 0 6 8 119 OTHER AQUATIC INVERTEBRATES Copepoda Cyclopoida 3 2 0 5 74 Pelecypoda Sphaeriidae Sphaerium spp. FC 1 5 3 9 133 Total No. of Organisms 37 39 44 120 Total No. Taxa / Sampler 5 5 4 Density (No./m 2 ) / Sampler 1644 1733 1956 1778 94 Appendix XIV. Benthic invertebrates collected from dredge samples during the fall season in 1994 in the ECE in Lake Martin, Alabama. Replicate Mean Taxon FFG A B C Total / 0.0225 m 2 AQUATIC INSECTS Diptera Chaoboridae Chaoborus spp. P 36 58 44 138 2044 Chironomidae Cladopelma spp. GC 0 0 1 1 15 Coelotanypus spp. P 15 37 29 81 1200 Microchironomus spp. GC 0 1 0 1 15 Procladius sp. P 0 0 1 1 15 Unidentified larvae GC 2 0 0 2 30 OTHER AQUATIC INVERTEBRATES Copepoda Cyclopoida 8 0 4 12 178 Oligochaeta Oligochaeta GC 1 0 0 1 15 Pelecypoda Corbiculidae Corbicula fluminea FC 0 0 1 1 15 Sphaeriidae Musculium spp. FC 0 4 0 4 59 Total No. of Organisms 62 100 80 242 Total No. Taxa / Sampler 4 4 6 Density (No./m 2 ) / Sampler 2756 4444 3556 3585 95 Appendix XV. Benthic invertebrates collected from dredge samples during the summer season of 1994 in the DCE in Lake Martin, Alabama. Replicates Mean Taxon FFG A B C Total / 0.0225 m 2 AQUATIC INSECTS Diptera Ceratopogonidae P 1 1 0 2 30 Chaoboridae Chaoborus spp. P 63 50 24 137 2030 Chironomidae Chironomus spp. SH 2 5 0 7 104 Coelotanypus spp. P 3 4 0 7 104 Procladius sp. P 2 2 2 6 89 Tanypus spp. P 2 3 0 5 74 OTHER AQUATIC INVERTEBRATES Copepoda Cyclopoida 0 1 22 23 341 Nematoda 1 0 1 2 30 Oligochaeta Naididae Pristina spp. Tubificidae GC 7 0 1 8 119 Branchiura sowerbyi GC 1 6 5 12 178 Pelecypoda Sphaeriidae Sphaerium spp. FC 0 1 0 1 15 Total No. of Organisms 82 73 55 210 Total No. Taxa / Sampler 9 9 6 Density (No./m 2 ) / Sampler 3644 3244 2444 3111 96 Appendix XVI. Benthic invertebrates collected from dredge samples during the fall season in 1994 in the DCE in Lake Martin, Alabama. Replicate Mean Taxon FFG A B C Total / 0.0225 m 2 AQUATIC INSECTS Diptera Chaoboridae Chaoborus spp. P 10 8 33 51 756 Chironomidae Ablabesmyia annulata gp. P 0 1 0 1 15 Chironomus spp. SH 0 1 2 3 44 Coelotanypus spp. P 5 9 7 21 311 Glyptotendipes spp. FC 0 0 1 1 15 Procladius sp. P 0 0 1 1 15 Tanytarsus spp. FC 0 1 0 1 15 Unidentified pupae 0 1 0 1 15 Ephemeroptera Ephemeridae Hexagenia spp. GC 0 1 0 1 15 OTHER AQUATIC INVERTEBRATES Cladocera FC 0 0 1 1 15 Copepoda Cyclopoida 0 1 0 1 15 Oligochaeta Tubificidae Branchiura sowerbyi GC 0 0 2 2 30 Unidentified GC 0 1 2 3 44 Pelecypoda Corbiculidae Corbicula fluminea FC 0 1 0 1 15 Total No. of Organisms 15 25 49 89 Total No. Taxa / Sampler 2 8 7 Density (No./m 2 ) / Sampler 667 1111 2178 1319 97 Appendix XVII. Benthic invertebrates collected from dredge samples during the summer season in 2002 in the ECE in Lake Martin, Alabama. Replicates Mean Taxon FFG A B C Total / 0.0225 m 2 AQUATIC INSECTS Diptera Chaoboridae Chaoborus spp. P 105 137 116 358 5304 Chironomidae Chironomus spp. SH 1 0 0 1 15 Trichoptera Leptoceridae Unidentified larvae GC 1 0 0 1 15 OTHER AQUATIC INVERTEBRATES Cladocera FC 0 2 1 3 44 Copepoda GC 0 1 0 1 15 Nematoda 1 0 0 1 15 Oligochaeta Tubificidae Branchiura sowerbyi GC 0 5 3 8 119 Unidentified GC 0 1 0 1 15 Pelecypoda Sphaeriidae Unidentified FC 2 19 0 21 311 Total No. of Organisms 110 165 120 Total No. Taxa / Sampler 5 5 3 Density (No./m 2 ) / Sampler 4889 7333 5333 5852 98 Appendix XVIII. Benthic invertebrates collected from dredge samples during the summer season in 2002 in the DCE in Lake Martin, Alabama. Replicates Mean Taxon FFG A B C Total / 0.0225m 2 AQUATIC INSECTS Diptera Chaoboridae Chaoborus spp. P 162 82 130 374 5541 Chironomidae Coelotanypus tricolor 0 4 1 5 74 OTHER AQUATIC INVERTEBRATES Copepoda GC 7 7 13 27 400 Oligochaeta Tubificidae Branchiura sowerbyi GC 1 1 1 3 44 Unidentified GC 0 2 3 5 74 Total No. of Organisms 170 96 148 414 Total No. Taxa / Sampler 3 4 4 Density (No./m 2 ) / Sampler 7556 4267 6578 6133 99 Appendix XIX. Benthic invertebrates collected from dredge samples during the fall season in 2004 in the ECE in Lake Martin, Alabama. Replicate Mean Taxon FFG A B C Total / 0.0225 m 2 AQUATIC INSECTS Diptera Chaoboridae Chaoborus spp. P 59 20 8 87 1289 Chironomidae Chironomus spp. SH 6 25 14 45 667 Coelotanypus spp. P 1 0 1 2 30 Cryptochironomus spp. P 0 3 2 5 74 Procladius bellus var. I P 5 2 6 13 193 OTHER AQUATIC INVERTEBRATES Cladocera FC 1 1 1 3 44 Copepoda 0 3 0 3 44 Oligochaeta Tubificidae Branchiura sowerbyi GC 0 5 10 15 222 Unidentified GC 1 0 0 1 15 Pelecypoda Sphaeriidae FC 1 0 0 1 15 Total No. of Organisms 74 59 42 175 Total No. Taxa / Sampler 6 7 7 Density (No./m 2 ) / Sampler 3289 2622 1867 2593 100 Appendix XX. Benthic invertebrates collected from dredge samples during the summer season in 2004 in the ECE in Lake Martin, Alabama. Replicate Mean Taxon FFG A B C Total / 0.0225 m 2 AQUATIC INSECTS Diptera Chaoboridae Chaoborus spp. P 70 60 42 172 2548 Chironomidae Chironomus spp. SH 0 0 1 1 15 Coelotanypus spp. P 1 0 0 1 15 OTHER AQUATIC INVERTEBRATES Pelecypoda Sphaeriidae FC 1 1 1 3 44 Total No. of Organisms 72 61 44 177 Total No. Taxa / Sampler 3 2 3 Density (No./m 2 ) / Sampler 3200 2711 1956 2622 101 Appendix XXI. Benthic invertebrates collected from dredge samples during the summer season in 2004 in the DCE in Lake Martin, Alabama. Replicates Mean Taxon FFG A B C Total / 0.0225 m 2 AQUATIC INSECTS Diptera Chaoboridae Chaoborus spp. P 79 166 142 387 5733 Chironomidae Chironomus spp. SH 0 0 1 1 15 Glyptotendipes spp. FC 0 1 0 1 15 Nanocladius (N.) alternanthera 1 0 0 1 15 Diptera Unidentified pupae 1 1 0 2 30 OTHER AQUATIC INVERTEBRATES Copepoda 1 0 0 1 15 Oligochaeta Tubificidae Branchiura sowerbyi GC 0 0 2 2 30 Total No. of Organisms 82 168 145 395 Total No. Taxa / Sampler 3 2 3 Density (No./m 2 ) / Sampler 3644 7467 6444 5852 102 Appendix XXII. Benthic invertebrates collected from dredge samples during the fall season in 2004 in the DCE in Lake Martin, Alabama. Replicates Mean Taxon FFG A B C Total / 0.0225 m 2 AQUATIC INSECTS Diptera Chaoboridae Chaoborus spp. P 46 17 21 84 1244 Chironomidae Chironomus spp. SH 9 3 6 18 267 Coelotanypus spp. P 3 2 1 6 89 Procladius (Holotanypus) sp. P 0 0 1 1 15 OTHER AQUATIC INVERTEBRATES Cladocera Daphnidae FC 2 0 0 2 30 Copepoda (blank) 0 2 1 3 44 Oligochaeta Tubificidae Branchiura sowerbyi GC 6 7 7 20 296 Unidentified GC 0 0 6 6 89 Total No. of Organisms 66 31 43 140 Total No. Taxa / Sampler 5 5 6 Density (No./m 2 ) / Sampler 2933 1378 1911 2074