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# Executive Summary

This summary of the state of the knowledge rela ted to nutrient effects on spring ecosystems and biota represents a synthesis of the most important findings from the literature. The report was commissioned by the Springs Initiative of the Fl orida Department of E nvironmental Protection (DEP) to review the pertinent literature related to nutrient effects on spring ecosystems and biota.

To address the complexity of spring ecosystems, this literature synthesis was prepared by a diverse team of spring researchers at the Univer sity of Florida’s Water Institute, Center for Environmental Policy, School of Forest Resources and C onservation, Department of Environmental Engineering Sciences, Department of Soil and Water Science, Department of Fisheries and Aquatic Sciences, and The Center for Governmental Responsibility; as well as expertise from the environmental cons ulting firm, Wetland Solutions, Inc.

The report is organized into eight chapters:

1. Springs as Ecosystems

2. Springshed Nutrient Loading, Transport and Transformations

3. Biogeochemical Processes and Imp lications for Nutrient Cycling

4. Nutrient Effects on Sp ring Flora and Fauna

5. Invasive Aquatic Plants and Aquatic Plant Control in Florida Springs

6. Effects of Nutrients on Spring Ecosystems

7. Managing Nutrient Inputs to Fl orida Springs: The Legal Framework

8. Synthesis of Nutrients and Springs

DEP’s mandate was threefold: 1) to prepare a re view of what is known about the interactions between nutrients and springs, 2) to identify what information gaps exist that inject substantial uncertainty for springs restoration and manage ment, and 3) make recommendations for public policy and research to address these gaps. Th e following summary draws the most important highlights from each chapter according to this mandate.iii 

### Chapter 1: Springs as Ecosystems

This chapter focuses on descri bing the abiotic and bi otic components of sp ring and spring run ecosystems, their forcing functions and inputs, and the known ecology of these unique aquatic ecosystems.

The Literature: What is Known

From this review of the existi ng literature on spring ecosystems th ere are several conclusions that can be drawn:

• Springs and spring runs represent a unique class of aquatic ecosystems typified by high water clarity, relatively constant water te mperature and chemical conditions, and complex biotic interactions.

• In their natural state, springs persist in a quasi environmental steady-state and due to this stability have developed complex biolog ical systems that are highly efficient at converting solar energy into useful productive work.

• During the past century humans have been exer ting an increasing variety of external and internal stresses on sp ring ecosystems in Florida, the most pervasive include:

o Alteration of discharge regimes re sulting from increased groundwater withdrawals,

o Increased levels of nutrien ts, particularly nitrate-N in groundwaters discharging from springs,

o Extensive recreational di sturbance of springs and spring runs, and

o Increased disturbance in th e form of management actions, for instance exotic species control.

The Literature: What is not Known, with Emphasis on Future Research Directions

Several studies suggest that multiple stressors ca n affect the productivity of spring ecosystems, suggesting that the measurement of ecosystem metabolism may allow for the quantification of these stressors as well as the health of spring ecosystems. To furt her develop these concepts the following research directions are suggested:

• Development of a holistic management st rategy for springs to foster better understanding of the relative importance and synergism of direct stressors at the local level and indirect stresses that occur at the regional scale.iv • Ecosystem-level information su ch as monitoring of ecosystem metabolism may provide a more sensitive and quantifiable indicator of the response of springs to nutrient increases and the myriad ot her anthropogenic stresses.

• Use of whole-spring and in situ mesocosm manipulation experiments to clarify the effects of individual forcing functions on spring ecosystem response.



### Chapter 2: Springshed Nutrient Loadin g, Transport and Transformations

This chapter focuses on nutrient loading to sp ringsheds, with an emphasis on processes of transport and attenuation that are germane to their delivery to springs. It discusses intrinsic vulnerabilities of karst areas to nitrogen load ing, and examines the effects of changes land use/land cover and hydrology on various aspects of wa ter quality decline. The chapter ends with an examination of the state of our understanding of sp ringsheds as closed systems of water and solutes, with an emphasis on areas of uncertainty that are relevant questions for future research.

The Literature: What is Known

From this review of the existing literature on nutrient loading there are four essential areas for which managers can proceed with relative certainty:

• Complex flowpaths and a wide distributio n of residence times characterize the hydrogeology of the Floridan Aquifer. The timeline and source area of water delivered to springs is of critical impor tance to springshed management,

• Karst landscapes are vulnerable to polluta nts from anthropogenic activities at the surface. It is clear that the risks of aquifer cont amination are greatly attenuated in regions of the state where regional c onfining layers (e.g., Hawthorn Formation) limit the rate of interaction between surface a nd Floridan Aquifer waters.

• Water quality is changing in dramatic ways including: nitrogen enrichment, hardness, dissolved oxygen concentrations, and tr ace quantities of pharmaceuticals.

• The preponderance of N pollution appea rs to be from fertilizer sources. While there are several reasons to trea t this finding as an over-genera lization, most of the accumulated evidence from mass balance computations a nd isotopic tracer studies suggests that mineral fertilizers, and therefore not septic tanks and wastewater sprayfields, are the principal sources of N pollution. However, sin ce fertilizer use is hi ghly diffuse and based on hundreds of thousands of individual decisions, control of N from point sources such as municipal wastewater effluent disposal activi ties may be a more cost effective form of N load reduction.v The Literature: What is Not Known, with Emphasis on Future Research Directions

What is not known about loading, transport and tr ansformations injects s ubstantial uncertainty that has key implications on pollutant delivery in karst systems and thus represent important areas of future research.

• What are the temporal dynamics of water age? It is clear from the literature that there remain large uncertainties about the age mixtur e of water discharging from spring vents.

• What is the natural variability in spring flow, and how vulnerable is this to human appropriation? Ongoing efforts to set minimum flows and levels (MFLs) should consider the role of flow in wa ter quality response variables.

• What are the sources of N in springsheds? While there is accumulating evidence that the source of N is principally mineral fertili zer, there remains significant uncertainty about the interpretation of bul k stable isotope measurements in complex karst hydrologic systems.

• What are the causes and effects of water quality changes beyond elevated nitrate? The synthesis of the literature on springs is dominated by information about nitrogen, but essentially devoid of similarly detailed st udies of other water quality parameters.

• What is the assimilative capacity of the la ndscape and aquifer for nutrients and other contaminants? Understanding the assimilative capacity and how it might be affected by development and enhanced by management , should be a research priority.

• How responsive are springsheds to management interventions? As we begin to understand the links between springshed activit ies and declining water quality in spring ecosystems, evaluating the extent to whic h parameters are management-sensitive is relevant to both timelines for recovery and prioritization of management efforts.



### Chapter 3: Biogeochemical Processes a nd Implications for Nutrient Cycling

The potential biogeochemical reactions and path ways occurring in spri ngs are numerous, and have been studied in varying le vels of detail. In this chapter the following three issues are presented: 1) a description of the biogeochemical processes related to nutrient cycling in spring systems, 2) a summary of the available literatu re related to biogeochemi stry in Florida spring systems, and 3) a discussion of the needs a nd directions for future spring research.

The Literature: What is Known

• In springs and spring runs, biogeochemic al processes occur in the water column, sediments and flood plains and can regulate the role of springs as a ‘source’ or ‘sink’ for nutrients and pollutants. The effect of biogeochemi cal processes on spring watervi composition begins immediately upon emergen ce from the boil and continues throughout the entire length of a spring run.

• Key interfaces involved in el emental cycling in springs environments are biofilms, algal mats and hyporheic sediments zones. A variety of processes can occur in these zones of high microbiological activity including organic matter deco mposition, nitrogen fixation, nitrification, denitrificati on, iron reduction, sulfate reducti on, and associated processes involving phosphorus.

• Organic matter compositions and distribu tions greatly affect the variety and importance of biogeochemical processes especially those requiri ng anaerobic conditions (e.g., denitrification).

• There is a high likelihood that the effect of increased nitrate may be coupled to other nutrient levels (e.g., phosphorus or metals) or phys ico-chemical parameters (e.g., dissolved oxygen). Some of th e observed association of increased nitrate with increased algal growth could be the result of ot her reactions occurring during groundwater emergence (e.g., sulfide oxidation and calcium carbonate precipitation).

The Literature: What is Not Known

• Because there are so few data characterizing the physical, chemical and microbiological conditions in most springs, it is extremely difficult to describe the exact nature and extent of biogeochemical reactio ns that are occurring in these systems.

• There is little information to ascerta in the involvement of other nutrient processes/cycles in controlling nitrate levels, including nitrate conversion and loss (e.g., iron and sulfur involvement in denitrification or al ternative loss processes)

• Among the studies which could advance our understanding of springs as sinks, sources, and transformers of nutrients are:

o Expanding water quality data collection to include additional geochemical parameters (e.g., sulfide) could improve our ability to e xplain and predict individual spring responses to nutrients.

o Increasing the frequency of water quality data collection (e.g., diel patterns) will offer additional insight into our ability to identify pathways (e.g., aerobic vs. anaerobic, phototrophic vs. he terotrophic) involved in spring processing/cycling of nutrients.

o Basing the collection of water quality da ta on upstream and downstream locations of stream segments will improve our ab ility to understand pr ocesses and spring run characteristics contributing to nutrient uptake and transformation.

o Characterization of sediments (e.g., nutrien t storages, organic matter distributions, mineral composition) would enable better assessments of benthic and hyporheicvii processes affecting current and future st orages and fluxes of sediment nutrients (sediment memory).

o Studies are needed to document the potenti al for aquatic plant management efforts to alter biogeochemical conditions (i.e., creating anoxic zones) and nutrient cycles (e.g., increasing denitrificati on loss of nitrate or incr easing release of ammonium and phosphorus).

o More studies are needed to document the involvement of individual system components in observed declines in nitrat e with distance in spring runs (i.e., dilution vs. biotic uptake vs. denitrification).

o More exploratory work is necessary to determine the existence and importance of novel pathways of nitrate conversion (e.g., anammox and lithotrophic nitrate reduction).



### Chapter 4: Nutrient Effects on Spring Flora and Fauna

Many of Florida’s springs and spring runs ar e enriched in nitrate due to broad-scale contamination of groundwater s upplies, which fosters two prim ary and interrelated concerns regarding the effects on flora a nd fauna in spring systems:

1) high nitrate concentrations can promot e eutrophication that can lead to a variety of changes in flora and fauna

2) high nitrate concentrations in Florida’ s springs and spring runs can affect fauna directly through toxicity.

The Literature: What is Known

• A eutrophication progression scheme suggests that increased nutrient delivery to aquatic systems may induce change, by favoring production of fast-growing algae that ultimately out-compete and displace nativ e vascular plants (see Duarte 1995).

• In some of Florida’s spring systems, both the apparent pro liferation of nuisance algae and the apparent decline of native vascular plants are consistent with the proposed eutrophication progression scheme, which a cue that management actions are needed.

• In general, nutrients can affect the fauna l assemblages in aquatic systems in three primary ways: 1) toxicity, 2) changes to trophic webs and 3) changes in habitats.

o Elevated concentrations of ammonia and nitrate can lead to increased mortality and sublethal effects.

o At this time, lethal effects are not an overwhelming concer n in most spring systems.viii o Toxicity needs to be factored into management of spring systems, especially given uncertainty factors that coul d be applied to most criteria

o The potential for changes in trophic webs and habitats, along with flow-on effects on faunal assemblages does exist.

o The form and magnitude of any changes are determined by complex interactions among bottom-up and top-down processes, su ch as grazing pressure, habitat use and predation pressure.

o The outcomes of ecological interactions cl early relate to sustainable management of spring systems

The Literature: What is Not Known

• There are few quantitative data that defi nitively link changes comprising the eutrophication progression scheme to incr eased nitrate loads in spring systems

o The eutrophication progressi on scheme predicts that nonlinear responses may have “decoupled” these systems from contem porary concentrations of nutrients in the water column.

o Currently, patterns in the distribution and abundance of plants and algae do correlate with other physical and chemical parameters, such as light availability.

o Little is known about the individual or combined effects of the various forcing factors or the form and strength of potential feedback mechanisms, including those linked to nutrient concentrations.

• There are no data to support rigorous evaluation of bottom-up and top-down influences on faunal communities

o Changes in faunal assemblages have not been linked to changes in vegetated communities and changes in grazer abundance have not been linked to changes in floral assemblages.

o Sublethal effects of nitrate may not translate to effects on populations, assemblages or ecosystems.

Recommendations for Future Research

• Management of nutrients in spring ecosy stems would benefit from compilation of inventories and collection of baseline data. Initial suggestions for discussion include inventories and baseline da ta collection designed to:

o Measure nutrients, oxygen, carbon, light, flow , substrates, grazing, algae, plants, invertebrates, fish, and other key driv ers and valued components of spring systems chosen with the best available information and, eventually, with guidance from the results of diagnostic studies.ix o target systems where effectiveness of ma nagement can be assessed (e.g., systems with total maximum daily loads, mini mum flows and levels, pollution load reduction goals, surface water improvement plans or basin management action plans).

o Sample synoptically or at spatial and temporal scal es that support rigorous analyses of interactions (note that th ese inventories and baselines form the platform for adaptive management).

o Balance breadth with depth (e.g., optimi ze taxonomic detail and replication).

• Diagnostic studies focused on bottom-up and top-down interactions are also critical. Initial suggestions include the followi ng recommended diagnostic studies to:

o Determine assimilation of nutrients by flora as a key component of nutrient budgets, including assimilation by epiphytes , macroalgae, vascular plants and microphytobenthos or microalgae found in sediments.

o Determine how nutrients and other driver s affect overgrowth, shading and other relationships among periphyton, macroalgae an d vascular plants in an effort to ascertain the need for management acti ons, including aquatic plant management and restoration of native species.

o Clarify limitations to growth of algae and plants by:

− focusing on species of algae and pl ants that are hypothesized or known to play major roles in nutrient assimilation

− including potentially limiting f actors other than nitrogen and phosphorus and interactions among limiting factors that can significantly influence the out comes of management (e.g., micronutrients, iron, oxygen, carb on, light, flow and substrate type)

− exploring relationships and consequences across multiple levels of biological and ecological complexi ty, including uptake and other physiological responses, growth and reproduction of individual organisms, and ecological changes in populations or assemblages to promote the success of manageme nt actions at the system level (note that some of the key, large- scale relationships form the core of an adaptive management approach)

− elucidating the direc tion, magnitude and form of limitations (e.g., nonlinear and nonreversible) to iden tify suitable targets and goals for management actions.

o Identify and elucidate changes in com position and function of faunal assemblages related to changes in habitats in an effort to identify habitats to be protected or restored by management actions.x o Determine palatability of flora to graze rs, rates of grazing, and the form and magnitude of predation by primary consum ers in an effort to identify flora, grazers and trophic links to be protected or restored by management actions.

o Identify and elucidate the toxic e ffects of ammonia and nitrate beyond physiological and individual levels of organization (e.g., population and assemblage effects) in an effort to determine threshold levels, safety factors and the need for management actions.



### Chapter 5: Ecosystem Implications of Invasive Aquatic Plants and Aquatic Plant Control in Florida Springs

Nonnative and nuisance plants such as water hy acinth, water lettuce, and hydrilla are a primary management concern in many Florida springs. In fiscal year 2005 – 2006, the DEP’s Bureau of Invasive Plant Management spent approximate ly $173,000 to control these plants in springs, primarily through the use of chemical herbicides.

This chapter reviews a broad range of literature to outline what is known and, in some cases, unknown about: 1) the history of these plants and thei r control in Florida, 2) the growth potential of the nonnative plants in springs as a function of elevated nitrate- nitrogen concentr ations, 3) the social and ecological consequences of aquati c plant overgrowth, 4) the ecological risks associated with current aquatic pl ant control methods, and 5) the pot ential benefits of alternative aquatic plant management appro aches in some springs systems. The following are the chapter’s major findings and research suggestions:

The Literature: What is Known

• Major problems with nonnative plants in Flor ida began with the introduction of water hyacinth, into the St. Johns River in the late 19 th century . Water hyacinth was documented in several springs ecosystem along the St. Johns River by the mid 1890s. Chemical control programs have maintained water hyacinth populati ons at low levels throughout Florida since the mid-1970s.

• Historical sightings by Willia m Bartram indicate that water lettuce, a floating aquatic plant, has been present in a number of Florida springs since at least 1765. Scientists disagree as to whether water lettuce was presen t in Florida before European colonization, or was introduced by early Spanish settlers. Chemical control programs have maintained water lettuce at low levels thr oughout Florida since the mid-1970s.

• Hydrilla became established in several areas of Florida, including the Kings Bay/Crystal River springs complex, by 1960. Sustained control of hydrilla, a submersed aquatic plant, has proven more difficult than the floating plants in Florida. Most aquatic plant management costs in Florida springs ecosystems over 2005 – 2006 were associated with chemical control of hydrilla.xi • Observations from several springs suggest a “boom-bust” successional sequence in which nonnative plants first out-compete native plant communities, and then suffer catastrophic population crashes associated with aquatic plant control or natural disturbances. Succession of springs into algal-dominated ecosystem states may be promoted by the nutrient pulses and ecological openings associat ed with the rapid loss of aquatic plant populations.

• Water hyacinth and water lettuce emit allelopathic compounds capable of suppressing a number of algal taxa. The effects of such allelopath ic compounds on algal dynamics in springs ecosystem are not presently known.

• Ecosystem surveys indicate that water hyac inth, water lettuce, and hydrilla provide attractive habitat for crayfi sh, apple snails, amphipods, fish, manatees, and other springs fauna at moderate levels of coverage.

• Observational accounts suggest that aquatic plant control activities may sometimes have significant adverse effects on springs fauna. Depression of dissolved oxygen due to decaying biomass is a primary concern to animals following aquatic plant control. Copper and diquat herbicides also pose concerns in terms of direct toxicity to some animals at levels used for aquatic plant control.

• Water hyacinth and water lettuce are curre ntly being managed for algal-suppression, nutrient recovery, and biomass utilization in a number of tropical countries, including places in which they are considered nonnative. Careful experimentation with similar ecosystem recovery methods may be worthwhi le in highly degraded springs ecosystems where these plants are established.

• Biotypes of hydrilla that are resistant to fl uridone, a systemic herbicide commonly used for hydrilla control in Fl orida lakes, have been docum ented in recent years. Thus, there is increased concern about the potential evolu tion of hydrilla strains that are resistant to Aquathol®, the contact herbicide most comm only used to control hydrilla in springs.

• Establishment of selective biological contro l organisms is increasingly viewed as a priority for sustainable cont rol of hydrilla in Florida. A potential biological control for hydrilla, the hydrilla tip mining midge ( Cricotopus lebetis) , has been documented in Kings Bay/Crystal River, and may be suitabl e for experimental introduction into other springs systems.

The Literature: What is Not Known

• Although scientific litera ture indicates clear relationships between nitrogen enrichment and increased growth of water hyacinth, water lettuce, and hydrilla in nonflowing aquatic systems, the few studies av ailable for springs and other flowing waters have not definitively determined a concentra tion of nitrate-nitrogen in springs that would be limiting to any of the nonnative plants. Nutrient assays in flowing water mesocosms would be necessary to develop nitrat e-nitrogen limitation values for these plants in springs conditions.xii • Although the scientific literature suggest s that copper and diquat herbicides can have significant adverse effects on algal community dynamics, these potential effects have not been well-studied in the specific context of Florida springs. Lyngbya wollei

, a filamentous cyanobacterium of great concern in many Florida springs, is notable for its relative resistance to these compounds as compared to other common algal and cyanobacteria taxa.



### Chapter 6: Effects of Nutrients on Spring Ecosystems

There exists a wide-spread recognition of th e environmental and economic importance of artesian springs in Florida. An ecosystem approach is essential to provide a greater understanding of the relative inte ractions between the myriad phys ical, chemical, and biological fluxes present in springs and their normal res ponses to rising nutrient levels. A focused and logical research agenda will be critical to pullin g these precious but threatened natural resources back from their current declining path.

The Literature: What is Known

• The best evidence suggesting a decline in th e health of Florida’s spring ecosystems comes from studies that have been conducted over a half-century time period in Silver Springs. Both the direct measurements and the estimated system metabolism analyses indicate that the Silver Sp rings ecosystem may be consid erably less productive than it was fifty years ago (Munch et al. 2006). The key findings from that study are:

o Sagittaria kurziana remains the dominant submersed aquatic plant species in Silver Springs and represents one of the main physical features of the ecosystem.

o Biomass estimates for submersed aquatic plants in the summer season were not significantly different from estimates made by Odum in the early 1950s. However, estimates for winter biomass were 31% lower than Odum’s, who reported no seasonal difference in su bmersed aquatic plant biomass.

o Biomass estimates for the epiphyte commun ity in the summer were approximately three-fold higher than thos e reported by Odum, while winter values were not significantly different be tween the two studies.

o The largest disparity between the Munch et al. (2006) estimates of primary producer community biomass and thos e of Odum from the 1950s was the substantial increase in biomass for the benthic algal mat community.

o Total species richness for birds, fish, and reptiles in the Munch et al. (2006) study were similar to historical records at Silver River.xiii o Estimated annual average fish live-weight biomass in the Munch et al. (2006) study has declined in Silver Springs since Odum’s study in the early 1950s by about 96%; and by 61% since Knig ht’s 1978-79 study (Knight 1980).

o Annual average gross primary productivity (GPP) and community respiration declined 27% and 26% respectively in the period between the 1950s and today and net community production declined by about 59%.

• Recent studies of the Wekiva River and Ro ck Springs Run (WSI 2007a) (spring runs averaging 0.69 and 0.84 mg/L of nitrate nitrog en, respectively)found similar results as those observed at Silver Springs with signif icantly lower gross primary productivity and photosynthetic efficiency demonstrating an inverse correlation between rising nitrate concentrations and reduced aqu atic ecosystem metabolism.

The Literature: What is Not Known, with Emphasis on Future Research Directions

Despite a wide variety of spri ngs’ research, there remains a significant knowledge gap between the real and perceived threats that nitrate pollu tion plays on the ecology of spring ecosystems. In summary the following recommended applied research activities are suggested:

• Baseline data on ecosystem-level structur e and response to key forcing functions, including sunlight, flow, and nutrient levels are lacking. Studies are needed that are fairly long-term (multiple years of repeated measures) and conducted over a representative sample of springs with a ra nge of forcing functi ons including discharge rates, groundwater chemistry, nutrient con centrations, and recrea tional intensities;

• Community response to increasing nutrients under controlled conditions are lacking. We suggest the design of in situ complex (multi-species) mesocosm studies to allow replication of spring plant community respons es to a range of nut rient conditions under realistic and relevant spri ng environmental conditions;

• Understanding of impacts and effects of management interventions are lacking. Whole-spring manipulation st udies are suggested to test the effects of possible management techniques such as controlling le vels of springshed nutrient loading, human recreational activities, alternative control methods for i nvasive species, and estimation of optimal consumer carrying capacities ( e.g. , manatee density).



### Chapter 7: Managing Nutrient Input s to Florida Springs: The Legal Framework

The legal framework for managing the nutrient po llution of Florida springs is potentially as broad as the scope of human activities that cont ribute nutrients to the springshed. Regulations on the discharge of wastewater are key elements.xiv The Literature: What is Known

• The federal Clean Water Act provides a se t of mandates and incentives for state
programs. The focus of the Clean Water Act is protecting surface waters through the regulation of point source discharges under th e National Pollutant Discharge Elimination System (NPDES). Discharges to groundwater may be regulated under this program if there is a “significant nexus” to the qu ality of navigable surface waters.

• The Safe Drinking Water Act (SDWA) protec ts public water supplies by establishing minimum criteria for drinking water quality and requiring states to regulate the underground injection of pollutants. Groundwater quality stan dards for nutrients are thus focused on the protection of public he alth. Florida has been delegated NPDES permitting authority by the Environmental Protection Agency (EPA) and implements an Underground Injection Control (UIC) program that is consistent with the SDWA. In addition, Florida regulates certai n sources of nutrient pollution that do no t fall within the direct jurisdiction of the federal program, such as nonpoi nt sources, agricultural discharges, and additional di scharges to groundwater.

o The regulatory program requires discharges to achieve effluent limitations based on the application of specified levels of technology or to achieve water quality standards, whichever is more stringent. Do mestic wastewater treatment plants are generally only required to use secondary treatment, except in areas where the Legislature has required advanced wastew ater treatment (AWT). Concentrated Animal Feeding Operations (CAFOs) ar e currently regulated under state rules pending the adoption by EPA of a new federa l rule. Those stormwater dischargers subject to NPDES permitting are required to reduce the discharge of pollutants to the Maximum Extent Practicable.

o Florida’s surface water quality standards must be reviewed and approved by EPA every three years. They consist of designa ted uses, narrative and numeric criteria for each of those uses, and moderating provisions.

• Waters designated as Outstanding Flori da Waters (OFW) have a general limitation on their degradation. A narrative nutrient standard prohibi ts altering nutrient concentrations “so as to cause an imbalance in natural popul ations of aquatic flora or fauna.” Numeric standards can also be de veloped, most commonly as Total Maximum Daily Loads (TMDL).

• The TMDLs process begins with determin ing whether a particular waterbody is “impaired” and demonstrating it does not meet water quality standards for a specific criterion. If there is sufficient data demonstra ting the concentration of the specific pollutant causing the impairment, it can be added to the verified list. A TMDL is then calculated and the load reasonably and equita bly allocated to the various sources and basins contributing pollutants. A Basin Mana gement Action Plan (BMAP) is developed for implementing the adopted TMDL.xv • Stormwater is a significant source of pollutant s that can be regulated as a point source, because much of it is collected into pipes or channels, or as a nonpoint source, because much of it comes from diffuse sources and activities. The NPDES program regulates many of the larger stormwater systems an d stormwater associated with industrial activities. Systems subject to NPDES permitting must be periodically reviewed and ways to reduce pollution must be considered in permit renewals.

• Most stormwater systems in Florida are either unpermitted or regulated under an Environmental Resource Permit program . Although they must be operated and maintained, there is no program for periodic review for compliance or enhancements.

• Septic tanks and other Onsite Treatment and Disposal Systems (OSTDS) are another major potential source of nutrient pollution. The Department of Health has adopted regulations for the construction and siting of se ptic tanks and drainfields. In some areas where nutrients are a concern, such as the Fl orida Keys, performance-based systems have been required. Local governments can adopt more stringent requirements.

• Local governments are required to periodi cally revise and implement comprehensive plans that can provide a framework for local springs protection. Local comprehensive plans and land development regulations can limit the intensity and design of land development to limit the contribution of nut rients to springs. They can provide for improved stormwater and wastewater management.

• The Florida Legislature may require special pr otections or enact other legislation that could protect springs and springsheds. The Legislature has specifically required local governments in the Wekiva Study Area to ame nd their comprehensive plans to enhance springs protection. It has also limited th e ability of local gove rnments to regulate agricultural land uses an d the conversion of agricultural la nd to urban uses. A Consumer Fertilizer Task Force has recommended ne w restrictions on the ability of local governments to regulate the application of fertilizers.



### Chapter 8: Synthesis of Nutrients a nd Springs: Knowns, Unknowns, Research Priorities and Mana gement Implications

This chapter represents a synthesis of the most important findings; greater details can be found within each of the chapters.. This synthesi s chapter is organized into three parts: 1) a summary of the “knowns” regarding nutri ent effects on springs and spring biota, 2) the unknowns and suggestions for future research, and 3) suggestions for a research and adaptive ma nagement program for spring ecosystems in Florida.

Part 1: NUTRIENTS and SPRING S: State of the knowledge

There is growing evidence that nitrate concentr ations are increasing in the water discharging from springs. Given the importance of nitrate-N a nd its role as a driver of primary production inxvi many aquatic systems, N appears to be the aspect of springs that has r eceived the most study. Chapter 8 summarizes the current state of knowledge concerning nutrients in springs.

PART 2: SPRINGS RE SEARCH PRIORITIES

Our review of the availa ble literature has shown that we do not have a systematic understanding of historic changes, fundamental processes, or mechanisms of ecosystem function sufficient to definitively manage springs and their contributing areas. Summarized in previous sections of this Executive Summary are topic areas in need of research related to sp ring and the effects of nutrients and other stressors on spring biota.

Overall we propose a rather holis tic research agenda that incl udes diagnostic studies, paired comparative studies of springs, manipulation studi es, and fine grained monitoring to elucidate cause and effect relationships and to test various hypotheses regarding the im pacts of stressors on spring ecosystems and biota. In particular we suggest:

• that a number of smaller springs be selected that represent a variet y of stressors acting separately and in combinatio n as a means of testing va rious hypotheses regarding the impacts of stressors on spri ng ecosystems and biota. Tests might involve, for instance, choosing springs with low versus high DO, pr esence and absence of grazers, low versus high nitrate-N, absence versus presence of recreation and so forth; then changing environmental conditions such as nitrate-N concentrations or DO concentrations in springs to support grazers, etc

• a relatively intense monitoring campaign in a li ke number of springs that is designed to elucidate spatial va riability and temporal dynamics of stressors and impacts.

Sustaining this long term research program will require a sustained source of funding. We suggest a simple source of funding, tied directly to the resource that is not a tax, but rather is a contribution for springs research.

• One-cent per bottle of water sold in Florida contributed for spring research will raise up to $21 million per year.

• Companies who contribute will use the cont ribution as a marketing tool with a green sustainable water logo on each bottle.

PART 3: RECOMMENDATIONS for SPRINGS POLICY and MANAGEMENT

Florida’s springs are important, unique resources and are under considerable pressure from changes in environmental drivers, direct human use, and indirect alteration of their chemical and physical environment.

• Our ability to understand causality betw een stressors and indicators in spring ecosystems is not sufficient to allow a fine grain approach to management.xvii • The complexity of the issues as well as the complexity of the springs’ social ecological system, comprised of a multit ude of actors from microbes to state government, requires an Adaptive Management (AM) approach.

• An AM approach required are new institutions and institutional frameworks as well as new science and experimental framew orks. W propose the creation of up to 6 Adaptive Springshed Action Programs (ASAPs).

• The proposed ASAPs build on the Springs Working Groups model by developing overall management goals and objectives, a nd then initiating a round of management interventions to test hypothese s concerning the causes of cu rrent apparent declines in spring ecosystem condition.

• The social framework of the ASAP team s hould include a tripart ite of individuals from government, science, and the citizenry at large.

• Research within the AM context is designe d for the purpose of producing results that may be applied to and answer management questions.

o It must be able to operate under conditions of complexity, uncertainty and risk, and generally be applied over larg e spatial and temporal scales.

o AM research crosses many ecological a nd organizational boundaries increasing its complexity and the lead time re quired for accomplishing its objectives

• The research within each of the ASAPs should ultimately be formulated by their participants, as they each have their particular ecological issues, management potentials, and socio-ec onomic constraints.xviii TABLE OF CONTENTS



### Chapter 1 - Springs as Ecosystems .................................................................................... 1
Summary........................................................................................................................ ......... 1

INTRODUCTION TO FL ORIDA SPRINGS ............................................................................ 1
Background..................................................................................................................... ........ 1
Springs as Ecosystem Laboratories ........................................................................................ 3
Spring Stability and the Evolutionary Time Scale.................................................................. 4

SPRINGS ECOSYSTEM DIAGRAMS..................................................................................... 5
Global/Regional Scale .......................................................................................................... .. 5
Springshed Scale............................................................................................................... ...... 5
Spring Scale ................................................................................................................... ......... 9

ENVIRONMENTAL FORCING FUNCTIONS...................................................................... 13
Solar Inputs ................................................................................................................... ........ 13
Atmospheric Forcing Functions............................................................................................ 15
Groundwater Inputs ............................................................................................................. . 16
Anthropogenic Forcing Functions ........................................................................................ 20

ENERGY/MATERIAL STORAGES IN SPRINGS ................................................................ 22
Spring Basin/Spring Run ...................................................................................................... 22

Water and Dissolved Constituents........................................................................................ 24
Primary Producers.............................................................................................................. ... 25
Primary Consumers.............................................................................................................. . 27
Higher-Level Consumers...................................................................................................... 29
Decomposers.................................................................................................................... ..... 29
Spring Biomass Pyramid....................................................................................................... 30

ECOSYSTEM-LEVEL PROCESSES...................................................................................... 31
Primary Productivity........................................................................................................... .. 31
Community Respiration........................................................................................................ 34
Productivity/Respiration Ratio.............................................................................................. 34
Photosynthetic Efficiency ..................................................................................................... 3 4xix Community Export............................................................................................................... . 35
Pollutant (Nutrient) Assimilation.......................................................................................... 35
Secondary Productivity......................................................................................................... 35

HUMAN AND AESTHETIC USES ........................................................................................ 36

Recreational Uses.............................................................................................................. .... 37
Visitation Seasonality ......................................................................................................... .. 38

Visitation Numbers ............................................................................................................. .. 39
Economic Impacts............................................................................................................... .. 42
Effects of Recreation on Spring Ecosystems........................................................................ 42
Human Use Carrying Capacity ............................................................................................. 43

DISCUSSION..................................................................................................................... ...... 44

REFERENCES ..................................................................................................................... .... 46



### Chapter 2 - Springshed Nutrient Load ing, Transport and Transformations .. 53
Summary........................................................................................................................ ....... 53

The Literature: What is Known ............................................................................................ 53
The Literature: What is Not Known, with Emphasis on Future Research Directions.......... 54

INTRODUCTION ................................................................................................................... . 57

HYDROGEOLOGY................................................................................................................. 61
Hydrology in a Karst Environment....................................................................................... 61
Typology of Florida’s Springs.............................................................................................. 65
Springshed Delineation......................................................................................................... 67
Hydrogeology and Nutrient Loading.................................................................................... 73
Assessing Aquifer Vulnerability........................................................................................... 74

Timescales of Water and Nutrient Fluxes............................................................................. 75
TRENDS IN WATER QUA NTITY AND QUALITY ............................................................ 77
Water Quantity................................................................................................................. ..... 77
Nitrogen ....................................................................................................................... ......... 79
Phosphorus..................................................................................................................... ....... 85

Carbon......................................................................................................................... .......... 87
Other Water Quality Parameters........................................................................................... 89xx SOURCES OF NUTRIENTS ................................................................................................... 92
Nitrogen ....................................................................................................................... ......... 92

Natural Sources................................................................................................................ . 94
Inorganic Sources.............................................................................................................. 95
Organic Sources................................................................................................................ 99
Isotope Tracing of N Sources ............................................................................................. 101
Synthesis of N Springshed Budgets.................................................................................... 104
Phosphorus..................................................................................................................... ..... 107
Carbon (BOD)................................................................................................................... .. 108
Other Water Quality Parameters......................................................................................... 109

SINKS AND TRANSFORMATI ONS OF NUTRIENTS...................................................... 110
Nutrient Sinks in Karst Landscapes.................................................................................... 111
Nitrogen ....................................................................................................................... ... 111
Phosphorus..................................................................................................................... . 113

Carbon......................................................................................................................... .... 114
Other Water Quality Parameters..................................................................................... 114
Springshed nutrient budget – Synthesis of Sinks................................................................ 114

UNKNOWNS AND RESEARCH PRIORITIES ................................................................... 118
Hydrologic and Geochemical ............................................................................................. 118
Nutrient Loading............................................................................................................... .. 118
Subsurface Nutrient Processing .......................................................................................... 119

REFERENCES ..................................................................................................................... .. 121



### Chapter 3 - Biogeochemical Processes a nd Implications for Nutrient Cycling

............................................................................................................................... .......... 135
Summary........................................................................................................................ ..... 135

The Literature: What is Known .......................................................................................... 135

The Literature: What is Not Known ................................................................................... 136

INTRODUCTION .................................................................................................................. 1 37

SPRING ENVIRONMENT.................................................................................................... 137
Spring water chemistry ....................................................................................................... 13 7xxi Reactive Zones................................................................................................................. ... 139

Biofilm ........................................................................................................................ ........ 140
Benthic Algal Mats ......................................................................................................... 141

Sediments...................................................................................................................... ...... 142

BIOGEOCHEMICAL PROCESSES...................................................................................... 144

Carbon......................................................................................................................... ........ 144

Carbon Sources ............................................................................................................... 14 4
Decomposition .................................................................................................................. .. 145
Nitrogen ....................................................................................................................... ....... 148
Biological Nitrogen Fixation .......................................................................................... 149
Nitrogen uptake............................................................................................................... 1 50
Nitrogen mineralization .................................................................................................. 151

Denitrification................................................................................................................ . 152
Alternative N pathways................................................................................................... 153
Overall N Cycle .............................................................................................................. 15 6
Phosphorus..................................................................................................................... ..... 156
Forms of phosphorus in water column and sediments.................................................... 157
Mineralization of organic phosphorus ............................................................................ 159

Biotic and Abiotic Ca Interactions.................................................................................. 160

Biotic and Abiotic Fe Interactions .................................................................................. 161
Overall P Cycle............................................................................................................... 1 62

Sulfur......................................................................................................................... .......... 162

Sulfide Oxidation............................................................................................................ 16 2

Sulfate Reduction............................................................................................................ 16 3
Overall S Cycle............................................................................................................... 1 64

EFFECTS OF INCREASED NUTRIENTS ........................................................................... 164
Carbon/Organic Matter Impacts ......................................................................................... 164
Nitrogen Impacts............................................................................................................... .. 164
Phosphorus Impacts ............................................................................................................ 1 65
Sulfur Impacts................................................................................................................. .... 166

SUMMARY AND CONCLUSIONS ..................................................................................... 166xxii Management and Policy Implications................................................................................. 167

RESEARCH NEEDS.............................................................................................................. 168
Specific Research Recommendations................................................................................. 168
Biogeochemical Characterization ................................................................................... 168

Source and Fate of N....................................................................................................... 169
Carbon Dynamics............................................................................................................ 169

Sulfur Distribution and Function .................................................................................... 170
Coupled Biogeochemical Cycles .................................................................................... 170

REFERENCES ..................................................................................................................... .. 171



### Chapter 4 - Nutrient Effects on Spring Flora and Fauna ..................................... 179
Summary........................................................................................................................ ..... 179

The Literature: What is Known .......................................................................................... 179

The Literature: What is Not Known ................................................................................... 179

Recommendations for Future Research.............................................................................. 180

INTRODUCTION .................................................................................................................. 1 82
Nitrate as a Compound of Concern..................................................................................... 183

PRIMARY PRODUCERS IN SPRING ECOSYSTEMS AND FACTORS AFFECTING

THEIR ABUNDANCE AND DISTRIBUTION.................................................................... 183
Light.......................................................................................................................... .......... 184
Carbon and Oxygen Balance in Macrophytes .................................................................... 186
Stream Velocity and Substrate............................................................................................ 187

Nutrients...................................................................................................................... ........ 188
Community Responses to Increased Nutrients ................................................................... 192

EFFECTS OF NUTRIENTS ON FAUNA............................................................................. 194
Components of Spring Ecosystems .................................................................................... 195
Relationships Between Nutrients and Invertebrates in Florida Springs ............................. 196
Relationships Between Nutrients and Fish in Florida Springs ........................................... 200
Summary of Relationships Between Nutrie nts and Fauna in Florida Springs.................... 207
Nutrient Effects Operating Through Ecosystem Processes ................................................ 207
Nutrient Effects Operating Through Toxicity..................................................................... 210xxiii Toxicity of Ammonia...................................................................................................... 211
Toxicity of Nitrate........................................................................................................... 2 13
Summary of Nutrient Eff ects on Fauna in Springs............................................................. 217

RESEARCH TO IMPROVE MANAGEMENT OF NUTRIENTS IN SPRINGS ................ 218

REFERENCES ..................................................................................................................... .. 220



### Chapter 5 - Ecosystem Implications of Invasive Aquatic Plants and Aquatic

Plant Control in Florida Springs ........................................................................ 231
Summary........................................................................................................................ ..... 231

The Literature: What is Known .......................................................................................... 231
The Literature: What is Not Known, and Recommendations for Future Research............ 232

INTRODUCTION .................................................................................................................. 2 34
Major Invasive Aquatic Plants: HI story and Control in Florida............................................. 235
Water Hyacinth ................................................................................................................. .. 235

Water Lettuce.................................................................................................................. .... 239
Hydrilla ....................................................................................................................... ........ 240

NITRATE-NITROGEN AND NONNAT IVE PLANTS IN SPRINGS................................. 242
Water Hyacinth and Nitrogen Limitation ........................................................................... 242
Water Lettuce and Nitrogen Limitation.............................................................................. 244
Hydrilla and Nitrogen Limitation ....................................................................................... 244

NONNATIVE PLANTS AND ALGAL COMMUNITY DYNAMICS ................................ 245
Boom-Bust Hypothesis ....................................................................................................... 245
Allelopathy Hypothesis....................................................................................................... 24 7
AQUATIC PLANT CONTROL AN D ALGAL SUCCESSION ........................................... 248
Herbicide Selection Hypothesis.......................................................................................... 248
Copper......................................................................................................................... .... 249
Diquat......................................................................................................................... ..... 249

Endothall...................................................................................................................... ... 250
Glyphosate ..................................................................................................................... . 250
2,4-D .......................................................................................................................... ..... 251
Attractor – Catastrophe Hypothesis.................................................................................... 251xxiv Habitat Values of Water Hyacinth.................................................................................. 251
Habitat Values of Water Lettuce and Hydrilla ............................................................... 253
Non-Target Concerns...................................................................................................... 254

RECOMMENDATIONS FOR FURTHER RESEARCH...................................................... 256
Alternative Management of Floating Plants ................................................................... 258
Alternative Management of Hydrilla.............................................................................. 259

CONCLUSION..................................................................................................................... .. 260

REFERENCES ..................................................................................................................... .. 261



### Chapter 6 – Effects of Nutrients on Spring Ecosystems ........................................ 271
Summary........................................................................................................................ ..... 271

ECOSYSTEM STRUCTURE ................................................................................................ 273
Primary Producers.............................................................................................................. . 273
Algae.......................................................................................................................... ..... 274

Vascular Plants................................................................................................................ 279
Primary Consumers............................................................................................................. 2 80

Invertebrates.................................................................................................................. .. 280

Vertebrates .................................................................................................................... .. 280
Secondary Consumers......................................................................................................... 281

Fish........................................................................................................................... ....... 281
Higher-Level Consumers................................................................................................ 281

Humans ......................................................................................................................... .. 282

ECOSYSTEM FUNCTION.................................................................................................... 282
Community Metabolism ..................................................................................................... 282
Gross and Net Productivity................................................................................................. 283
Respiration .................................................................................................................... ...... 287
Productivity/Respiration Ratio............................................................................................ 287
Photosynthetic Efficiency ................................................................................................... 287
Community Metabolism Conclusions................................................................................. 288

HUMAN AND AESTHETIC USES ...................................................................................... 289

Visitor Satisfaction........................................................................................................... ... 289xxv Economic Impacts............................................................................................................... 289
Effects of Eutrophication on Human Recreation................................................................ 289

DISCUSSION..................................................................................................................... .... 290

RECOMMENDATIONS FOR FURTHER RESEARCH...................................................... 296
Proposed Research Priorities .............................................................................................. 297

Recommendations........................................................................................................... 298

REFERENCES ..................................................................................................................... .. 299



### Chapter 7 - Managing Nutrient Input s to Florida Springs: The Legal

Framework ................................................................................................................. 305
Summary........................................................................................................................ ..... 305

REGULATION OF DISCHARGES....................................................................................... 307

Clean Water Act.............................................................................................................. 30 7
State Discharge Permitting ............................................................................................. 309

Domestic Wastewater ..................................................................................................... 310
Industrial Discharges and Concentrated Animal Feeding Operations (CAFO).............. 311
WATER QUALITY STANDARDS....................................................................................... 312
Groundwater Quality Standards...................................................................................... 312
Surface Water Quality Standards.................................................................................... 313

TOTAL MAXIMUM DAILY LOADS.................................................................................. 315
Impaired Waters............................................................................................................. 31 6

The Verified List............................................................................................................. 3 17

Calculation and Allocation ............................................................................................. 317
Basin Management Action Plans (BMAP)..................................................................... 318

STORMWATER..................................................................................................................... 319
NPDES Stormwater Permitting ...................................................................................... 321

SEPTIC TANKS................................................................................................................... .. 321

LAND USE PLANNING AND REGULATION................................................................... 322xxvi 

### Chapter 8 - Synthesis of Nutrients and Springs: Knowns, Unknowns,

Research Priorities a nd Management Implications .................................... 327
Summary........................................................................................................................ ..... 327

INTRODUCTION .................................................................................................................. 3 29

PART 1 - NUTRIENTS and SPRINGS : State of the knowledge.......................................... 329
Fate, Transport, and Transformation: What we know .................................................... 329
Effects of Nutrients on Spring Organisms and Systems: What we know ....................... 331

Effects on Flora and Fauna ............................................................................................. 332

Effects on Invasive Plants and I nvasive Plant Effects on Springs.................................. 334
Ecosystem Scale Effects ................................................................................................. 334
The Legal Framework..................................................................................................... 335

PART 2: SPRINGS RESEARCH PRIORITIES .................................................................... 336
Nutrient Fate, Transport, and Transformation: What we don’t know ............................ 336

Nutrient Effects on Flora and Fauna: What we don’t know ............................................ 338
Invasive Plants and Springs: What we don’t know ........................................................ 338
Ecosystem Scale Effects: What we don’t know .............................................................. 339
Summary of Research Recommendations ...................................................................... 340

PART 3: RECOMMENDATIONS for SPR INGS POLICY and MANAGEMENT ............. 341
The Springs Initiative: Prot ecting Florida’s Springs ...................................................... 342
Adaptive Management: An Approach to Springs Protection ........................................ 343
A Tragedy of the Commons............................................................................................ 343
Managing in the Face of Uncertainty.............................................................................. 344
Adaptive Springshed Action Programs (ASAPs) ........................................................... 344
Adaptive Spring and Springshed Research..................................................................... 347

REFERENCES ..................................................................................................................... .. 3491 

# 1

Springs as Ecosystems

Robert L. Knight, 1, 2 and Sky K. Notestein 1

1 Wetland Solutions, Inc.
2 Department of Environmental Engineeri ng Sciences, University of Florida

Summary Springs and spring runs represent a unique class of aquatic ecosystems typified by high water clarity and relatively constant water temperatur e and chemical conditions. Structurally, springs are best characterized by thei r basin morphometry (shape and volume), their position in the landscape, and their flora and fauna. A classic aquatic food web and biomass pyramid in spring ecosystems includes numerous primary produ cers and smaller numbers of primary and secondary consumers. Research has demonstrated that in their natural state, many springs are highly productive and persist in an approximate steady-state compared to most other aquatic ecosystems. During the past two centuries humans have been exer ting an increasing variety of stresses on spring ecosystems in Florida. These ex ternal and internal stre sses include decreasing groundwater supplies and increasing levels of nutrients, extensiv e recreational disturbance, and management alterations. Due to their physical and biological complexity combined with their adaptations to provide optimal le vels of system productivity, a ho listic ecosystem-level approach to spring studies is necessary to provide a gr eater understanding of an thropogenic effects (such as nutrient enrichment) on spring ecosystems. Development of a management strategy for springs will be dependent upon a better understanding of the re lative importance and synergism of direct stressors at the local level and indirect stresses that oc cur at the regional scale. Systemslevel metrics have promise for providing a be tter understanding of the properties that many springs have in common. Use of ecosystem-level information is expected to provide a more sensitive indicator of the response of springs to nutrient increases and concurrent alterations due to the myriad other anthropogenic stresses.

INTRODUCTION TO FLORIDA SPRINGS

Background

Springs and their associated spring runs are a unique class of aquatic ecosystems. Since their principal water source is groundwater, many springs have water that is crystal clear, yet rich with dissolved nutrients and gases. Often this water quality provides the basis for a diverse assemblage of aquatic flora and fauna whose productivity is primarily determined by light availability and secondarily affected by the availa bility of macro and micr o nutrients and by the ambient groundwater temperature. Additional factors that may aff ect spring productivity include external forcing functions such as diffusion of atmospheric gases, rainfall inputs of water and nutrients, immigration of fauna, and anthropogenic perturbations.2 The large number and diversity of springs in Florida represents a globally significant concentration of these ecosystems. It is likely that many of the larger springs ecosystems have been in existence for approximately 15,000 to 30,000 years, since the end of the last major ice age (Martin 1966, Munch et al. 2006). This extended time peri od has allowed the evolution of highly adapted plant and animal communities.

Springs represent an important resource for human utilization as well, both by indigenous peoples as supported by archeologi cal evidence, and by present da y populations who utilize them for a variety of recrea tional purposes (Scott et al. 2002). There is little doubt that the intrinsic aesthetics of clear, cool water vigorously em anating from the underground will continue to interest and fascinate huma ns. Many of Florida’s olde st tourist attractions ( e.g. , Weeki Wachee mermaids, Homosassa fish bowl, and Silver Spri ngs glass bottom boats – see Figure 1) utilize spring ecosystems as their main draw. In ad dition, numerous spring boil areas have been modified to facilitate swimming, recreation, and ev en “health spas”. Today all of the largest springs in Florida, whether privately or public ly owned, are managed as recreational parks, which, in turn, attract a large number of visitors and generate many millions of dollars in revenue on an annual basis. Correspondingly, many springs have suffered declines (often unintentional) in their condition ( e.g., up-rooting of vegetation, bank erosi on, litter, etc.) due to visitation by ever increasing numbers of people.

However, over the last two decades other more seri ous factors, with the potential to permanently alter Florida’s spring ecosystems, have been incr easingly recognized. Two of the most apparent anthropogenic factors that may be causing significan t changes in spring ecosystems are: 1) the reduction in groundwater supply through cons umptive human withdrawals; and 2) the simultaneous pollution of groundwat er, principally with nitrate- nitrogen, resulting from human land use changes, cultural practices, and genera l population growth. The full spectrum of the combined effects associated with flow reduction s and nutrient enrichment on spring ecosystems is not yet clear. However, there is justifiable co ncern for potential negative consequences due to the obvious degradation of other aquatic habitats worldwide subj ected to declining flows and water levels, and increased nutrient loading. The potential consequences of nutrient enrichment in springs include an increase in opportunistic primary producer s and organic matter deposition, an increase in nuisance algae sp ecies biomass, a decrease in plant and animal productivity and diversity, a reduction in water quality and not in significantly, a reduction of the aesthetics these ecosystems have long provided. Clearly, an effort to synthesize the current state of knowledge concerning the effects of nutrien t pollution on spring ecosystems is warranted. It is in this context that this project began.3 Figure 1 – An artist’s representation of life in Sil ver Springs at the time of H.T. Odum’s 1957 landmark study (illustration by Elizabet h A. McMahan, from Odum et al. 1998).

Springs as Ecosystem Laboratories

Spring ecosystems have been the focus of some notable ecological studies, due to their diverse flora and fauna, and because their stable envi ronmental characteristics generate natural controlled settings suitable for general ecological research. Springs research is broad and varied, ranging from descriptive classifi cations based on water discharge and chemistry, to observations of endemic flora and fauna, to studies of community ecology through whole system measurements. Some attempt has been made to use the term “Crenobiology” (derived from

creno , the Greek root for spring) as a banner for "the study of life in springs" (Botosaneanu 1998).

The study of community ecology generally follows one of two methodologi es. The first, the “reductionist” method, in which single species a nd their associated eco logical controls are studied, in turn may be combined with other species-centric studies to assemble a working community model. This approach is generally li mited to the study of dominant species, and even if multiple components are quantified and combined, the resulting sum of the parts is unlikely to represent a functional community. The sec ond method, the “holistic” or “systems-level” approach, attempts to study the community in its entirety. The major challenge to the holistic approach is the concern that not enough useful information is obta ined concerning the fate of4 individual species and resources th at might be of greatest practical interest to managers and the public. Clearly, a comprehensive knowledge of ecosystems can only be obtained using both holistic and reductionist studies. A central thesis of this chapter is th at our understanding of spring ecology and how to apply effective mana gement and protection to spring ecosystems should be based on both type s of studies together.

Spring ecosystems (or, simply springs), are highly suitable for the study of community ecology and ecosystem function at a holistic level. This is because the external forcing functions that determine spring structure and function are more stable than for many ot her aquatic ecosystems. This environmental stability, in turn, results in le ss temporal complexity in springs than in other aquatic ecosystems and more cons istent food webs based on relati vely stable groups of primary producers. In addition, springs offer a range of different community structures as a response to their differences in physical and chemical conditi on, creating a natural experiment with which to relate ecosystem function.

Another important reason to study spring ecosystems is that springs provide an opportunity to characterize the groundwater upon which humans are so dependent. A better understanding of land use changes and their resu lting effects on groundwater qua ntity and quality changes has emerged through the study of springs (Munch et al. 2006; Cohen 2007). An appropriate analogy that supports the need for increased emphasis on springs monitoring and ecological research is that these environments are comparable to the “c anary in the cage” used to warn coal miners of impending disaster. Florida needs a greater focus on ecological research and monitoring to gain a better understanding of just how much ecological risk already exists due to the nutrient enrichment and eutrophication of our formerly pristine springs.

Spring Stability and the Evolutionary Time Scale

The concept of springs as stable-state ecosyst ems likely originated during the early 1950s from observations made by collaborating naturalists from the University of Florida, including D.K. Caldwell, J.H. Davis, D. Natelson, H.T. Odum , W.C. Sloan, L.A. Whitford, and J.L. Yount. During this time, these and other researchers were intensively surveying the larger springs of Florida to describe the flora and fauna found ther e and in recognition that these ecosystems could serve as ideal sites for experimental studies. In a report summarizing work completed during 1953, H.T. Odum states that community meta bolism, temperature, and major chemical constituents are “essentially constant” (Odum et al. 1953). The steady-stat e concept of springs was further expanded by Whitford (1956) in his obs ervations that discharge, temperature, gas, and mineral content of spring waters were al l relatively constant w ithin individual spring ecosystems (although quite different between springs ) and that only light varied greatly as a consequence of temperate latit udes. In his landmark paper Primary Production in Flowing Waters (1956) Odum uses the term “chemostatic” to describe Silver Springs and to propose that spring ecosystems can be studied as quasi stea dy-state systems. The stability of spring ecosystems must also be viewed in context of larg er time scales, such as evolutionary or geologic time scales. The stability of Florida’s springs can extend only back in time to, at most, the last ice age (~ 15,000 to 30,000 years befo re present) when sea levels we re dramatically lower. Yet, there are no other ecosystems outside of the tr opics and the ocean that have had such a long period of relatively constant external forcing functions a nd resulting community adaptation.5 As dynamic entities composed of physical, chemi cal, and biological com ponents and fueled by a variety of external energy flows, springs and spri ng runs have all the qualif ications required to be considered as “ecological systems” or ecosystems. Hence, this chapter provides an overview of Florida’s artesian springs and spring runs as ecosystems, with an emphasis on describing their system-level properties. This systems approach is in contrast to the more common descriptions of aquatic ecosystem structure as defined by in dividual water quality pa rameters or plant and animal species lists. When available, additi onal information from springs and spring run ecosystems from other localities in the United States and elsewh ere is utilized to emphasize, enhance, or contrast information from Florida springs thr oughout this chapter.

SPRINGS ECOSYSTEM DIAGRAMS

Spring and spring run ecosystems are products of their environments and may be evaluated by considering their surroundings and their relationship with this environment. Appropriate scales for consideration of springs are global, regional, and local. Examples of how each of these scales affects spring ecosystems are provided below.

Global/Regional Scale

Florida’s springs are a part of the global and continental environment and economy. Their location between north latitudes 30.8 and 27 (d ecimal degrees) puts them in the southern temperate zone of North America, an area rich in rainfall and groundwater, influenced by moderate temperatures, and subject in more coas tal areas to salt water influences. As global and national economies influence land us e changes in this region of Flor ida, springs are affected. For example, agricultural products in the Suwannee River area of Fl orida include dairy products and row crops, both industries that are subject to regional market factors. These land uses, in turn, alter spring hydrology and water quality throu gh their irrigation requirements and increased loads of nitrogen and other pollu tants to underlying groundwater.

A second regional influence on spring ecology results from their unique aesthetic and recreational opportunities for humans. Several of Florida’s springs ar e economic engines, attracting a high rate of intern ational and interstate visita tion (5% and 31.5% of visitors, respectively, Bonn 2004). These visitors spend money that fuels local economies that have the potential to affect th e springs’ ecologies thro ugh mechanisms such as groundwater pumping, increased surface water runoff pollutant transpor t, and the direct and indirect effects of recreational use and aquatic weed control.

Springshed Scale

The life-blood of Florida’s springs is groundwater, principally derived from the Floridan aquifer which is largely supplied by regi onal rainfall. The quantity a nd quality of this groundwater determines the basis for much of the aquatic ecology of Florida’s springs. Just as surface water features rely on a watershed, spri ngs are dependant upon their “sprin gshed”, the area of land that constitutes the majority of their source of water to the aquifer, and can also be called the capture zone, catchment basin, or contributory area.6 Figure 2 illustrates the springshe d delineations of many of Florid a’s largest spring ecosystems. These springshed delineations are held by th e Florida Geological Survey (FGS) and were produced using multiple methods using many data sets including surface water and ground-water flows, potentiometric levels, subsurface condu it maps and dye trace studies and many other hydrogeologic information (FGS 2007). The inset portion of Figur e 2 shows an image of the Ichetucknee Springs springshed located in nor th central Florida. A large number of anthropogenic activities w ithin this springshed have the poten tial to affect th e ecology of these springs.

Figure 2 – Florida springshed delineations, with an in set of the Ichetucknee springshed in North Central Florida (FGS 2007).

The symbolic “energese” language of H.T. Odum is shown and briefly defined in Figure 3a, while Figure 3b represents a conceptual diagram that illustrates the linkage between terrestrial ecosystems, developed land uses, and groundwat er within a Florida artesian spring at the7 regional scale. The diagram includes the major fo rcing functions of sunl ight, atmospheric inputs (wind, rain, and storms), with associated wate r and nutrients, and economic drivers (tourism, markets, and goods) interacting with developed land use changes. This conceptual energy diagram focuses on flows of nitrogen and ener gy, with counter flows of money for those pathways mediated by human activities. Variable s in the diagram include: natural lands and low intensity land uses that do not receive signifi cant inputs of nitrogen from anthropogenic uses; developed lands that include a ll other land uses; the water and nitrogen associated with the natural and developed lands; the artesian groundw ater within the basin; and the physical and biological structure of a spring. Key system -level exports in the diagram include evapotranspiration, outflows of goods (recreationa l services), and downstream exports of water and organic material.

Figure 3a – Energy symbols in the “Energese” mode l language, from Odum et al. (1998).8Figure 3b –Conceptual springs regional diagram (M.T. Brown 2008). Key forcing function atthe regional scaleinclude: sunlight,wind, precipitation, and associated nutrients,hurricanes, and human goods and services.Springsheds are in turn influenced andmodified by their associated undevelopedand developed terrestrial environments withfertilizer and waste nutrient inputs andcumulative synergistic impacts resulting from consumptive water uses, recreation, and various resource management activities.9 Documentation of the effects of regional land us e changes in springshed basins on water quality changes in major springs (especially nitrate concentrations) has substantially improved in the past five years (Munch et al. 2006; Cohen 2007). Munch et al. (2006) conducted a detailed evaluation of land use changes in a five-mile radius around Silver Springs and nitrate concentrations in the Silver Springs system. Th is analysis found that approximately 75% of the water discharging from Silver Sp rings originates within a two-y ear capture zone contained in a 6.4 km (4-mile) radius around the spring covering about 135 square km (52 square miles). Analysis of aerial photos taken during the past fifty years suggests that land cover within the two-year capture zone has changed from a pr edominantly natural landscape to a mostly urban/agricultural area. This cha nge corresponds to increased nitr ogen loading, principally as a result of increased groundwater nitrate concentr ations. Figure 4 provides another example of the observed correlation between increases in human population and increases in nitrate concentration of the ground water dischargi ng from Weeki Wachee Spring in Hernando County.

Figure 4 – Population growth (diamonds) in Her nando County, Florida and nitrate nitrogen concentration (triangles) at Weeki Wach ee Spring from 1923 to 2006 (from DEP 2006).

Spring Scale

Springs in Florida occur wher e the underlying aquifer is at or near the surface allowing groundwater with net positive head pressure to be discharged. Springs are also found where surface water features have inci sed into the underlying karst geological formations. Many of Florida’s springs are found along the banks or the bottom of streams, and are therefore integral components of larger stream systems.

A typical spring ecosystem exhibits a large amount of physical, chemical, and biological complexity. The proximal external factors affecting this complexity are sunlight, artesian10 groundwater inputs, biological exch anges, and human activities. In general, a spring ecosystem includes all the internal abiotic and biotic components typical of most aquatic ecosystems including: aquatic macrophytes (h igher plants) that support a di verse assemblage of attached algae (periphyton, benthic algae, and pseudo-plankt on), detritus (dead plant and animal material generally associated with benthi c organic sediments), associated animals feeding on detritus, and a faunal food web of herbivores, omnivores, a nd carnivores. Because of their continuous discharge, springs commonly serve as headwaters for streams, rivers, and sometimes lakes and ultimately contribute a great deal of freshw ater to downstream inland and coastal aquatic ecosystems.

Figure 5 provides a systems diagram illustrating so me of the immense complexity inherent in a relatively unaltered Florida artesian spring ecosy stem. This model provides a summary of the major external forcing functions, internal ener gy storages and processes, and interactions between the spring ecology and the human econom y. Functional groups with in trophic levels illustrate the potential complexity of a spring ecosystem and the multiple interactive linkages.

Figure 6 provides an aggregated diagram of a spring ecosystem modified by human activities. Aggregation of functional groups is useful for the sake of combining elements of like structure and function into larger wholes. At this level of diagram aggregat ion illustrated fo rcing functions include: sunlight, atmospheric i nputs of water and nutrients, gr oundwater inputs of water and nutrients, sediment derived input s of nutrients, and anthropogenic inputs such as management actions, goods, services, and people. Aggreg ated outflows in the diagram include evapotranspiration, surface water discharges wi th downstream particulate matter and nutrient export. The spring ecosystem diagram illustrated in Figure 6 is entirely included within the springshed landscape diagram illustrated above in Figure 3.11Figure 5 –Detailed conceptual diagram of a Floridaartesian spring ecosystem (M.T. Brown 2008).12Figure 6 –Aggregated spring ecosystem diagram that incorporates human recreation and management activities (M.T. Brown 2008).13 ENVIRONMENTAL FO RCING FUNCTIONS

Solar Inputs

Solar inputs comprise the largest portion of the total external energy delivered to natural communities, and, in turn, are converted by phot osynthesis into organic matter and other forms of potential energy, with the entropy of heat lo sses equally balancing inputs. The conceptual model of energy flow through a natural comm unity, described and quantified by Odum (1956, 1957a, 1957b) in his work at Silver Springs, Florid a, documented that only a fraction (24%) of the total solar energy (insolation) delivered to the system was absorbed by primary producers, which subsequently converted only about 5% of this energy into gross primary production (Figure 7). While numerically low, this conversi on of solar energy to usef ul plant sugars (also called photosynthetic efficiency) is high compared to most natural and human-cultivated plantbased systems.

As noted earlier, the majority of springs in Florida are located between 3 0.8 and 27 degrees north latitude, and are subject to seasona l variation in incident light due to the orientation of the sun. Maximal solar inputs occur during the spring an d summer and minimal solar inputs occur during winter, with corresponding patterns in daylight length. Solar inputs occur acro ss much of the range of the electromagnetic spectrum (< 200 to 5,000 nm) but only certain portions of that spectrum are readily utilized by photosynthetic organisms for primary production. This portion of the spectrum, spanning the range of 400 to 700 nm wavelengths, commonly referred to as photosynthetically activ e radiation (PAR), is typically m easured with a quantum sensor and expressed in units of μmoles of photons per s quare meter per second (Kirk 1994). Figure 8 illustrates a typical annual cycle for insolation as well as the PAR portion of that insolation.14 Figure 7 - Estimated energy flow diagram for Sil ver Springs, Florida, with all units in kilocalories per square meter per year (from Odum 1957b).15 0 50 100 150 200 250 300 350 400 450 500

1 /30/ 0 4 3/2 0/

0 4
5 / 9/04 6/28/ 0 4

8 / 1 7/

0 4 10/6/0 4

1 1/

25/04

1 /14/ 0 5 3 / 5/05PAR (uE/s/m2

0 50 100 150 200 250 300 350

1 /30/ 04 3/2 0/

04
5 /9

/04 6/28/ 04

8 / 1 7/

04 10/6/04
11 / 25/04

1 /14/ 05 3 /5

/05Total Insolation (J/s/m2)

Figure 8 – Annual cycle of daily insolation and phot osynthetically active radiation data recorded at Silver Springs, Marion Count y, Florida (from M unch et al. 2006).

Atmospheric Forcing Functions

Broadly, atmospheric inputs incl ude both wetfall (precipitation and associated dissolved and particulate substances contained within precipitation) as well as dryfall (dry particulate matter carried by the wind). Average annual precip itation (USGS records from 1951 to 1980) for mainland Florida range from more than 64 inches in the panhandle to less than 48 inches near Tampa Bay. Most of Florida’s springsheds re ceive between 52 and 60 in ches of rainfall on an annual average basis (Miller 1997). The majority (~ 80%) of th is precipitation is lost to evapotranspiration, a portion goes to surface runoff or shallow aquifers where it is delivered to streams and lakes (14%), and the balance (6 %) percolates downward to recharge deeper groundwater aquifers (Fernald and Purdum 1998).

Springs receive atmospheric nutrie nt inputs directly on their wate r surface area and indirectly through their groundwater inputs and surface wate r runoff. Nitrogen, phosphorus, and numerous other elements and compounds are deposited on the surface of the land and water via the atmosphere. Due to their relatively small open water areas, springs and spring runs derive the16 majority of their water and nutrient inputs from groundwater; direct at mospheric inputs are generally negligible.

Atmospheric nitrogen can be delivered to the environment as inorganic forms dissolved in rainwater or as particulate or ganic and/or mineralized forms. Atmospheric nitrogen exists primarily in the form of inert nitrogen gas (N 2 ), but inorganic nitrogen reaches the atmosphere through volatilized ammonia from industrial emissions, liquid ammonia fertilizers, and organic wastes. Emissions of oxides of nitrogen (NO X ) to the atmosphere occu r through denitrification, the incomplete combustion of fossil fuels and industrial processes.

Because atmospheric nitrogen deposition occurs over the entire springshed, the cumulative load of nutrients to a spring via atmospheric inputs can be substantial. Measurements of total inorganic nitrogen (nitrate and ammonium) atmo spheric deposition are obtai ned from monitoring stations managed by the National Atmospheric Deposition Program (see website for information and data: http://nadp.sws.uiuc.edu/ , accessed December 2007). Detailed descriptions of the magnitudes of common sources and si nks for nutrients in springsheds are described in Chapter 2, and as demonstrated, the atmospheric deposition lo ading to springsheds can represent substantial inputs. Mitigating factors that may reduce the impact of atmospheric deposition on spring nutrient concentrations include low concentration per unit area, low daily input, and the likelihood of uptake by plants.

In springs, the gas exchange of oxygen and carbon dioxide between air and water is critically important (Matson and Harriss 1995). Due to thei r subsurface water source, springs frequently have low concentrations (and pe rcent saturation) of dissolved oxygen in their upstream reaches. Atmospheric oxygen is the primary driving force to bring dissolved oxygen levels to saturation, with some input of dissolved oxygen being supp lied by primary production. Carbon dioxide is an essential nutrient for plant growth in spri ngs and other aquatic ecosystems. Hence, the atmospheric supply of this gas (about 3.5% by volume) is a component in replenishing the supply of inorganic carbon in many spring runs , although spring waters can initially be supersaturated in dissolved carbon dioxide due to their groundw ater source (Duarte et al . 2007).

Other important atmospheric forcing functions in clude the effect of wind on mixing of waters and the transport of seeds and reproductive prop agules. Many algal and plant species are highly motile in their resting and reproductive stages (Gurevitch et al. 2002). Those species with propagules that are carried by wind easily co lonize spring heads, which are otherwise inaccessible to floating propagules. Atmospheri c forcing functions may also be broadly interpreted to include invertebrate and vertebra te air borne inputs. The adult forms of aquatic invertebrates commonly occurring in springs (especially midges, stoneflies, and aquatic moths) depend upon flying skills to propaga te in and re-colonize head sp ring areas. A large variety of aquatic and wetland dependent bird s are very active in spring ecosystems and regularly migrate into and out of these systems. (Hoyer et al. 2006).

Groundwater Inputs

Groundwater inputs are the major h ydrological, elemental, and mineral forcing functions for the artesian spring ecosystems of Florida. The gr oundwater supplies of Florida are derived from17 precipitation, primarily within the state’s boundaries, alth ough some groundwater is derived from Alabama and Georgia precipitati on (Fernald and Purdum 1998).

Florida has five principle aquifer systems. From largest to smallest they are: the Floridan, Biscayne, Surficial, Intermediate, and the Sand a nd Gravel aquifers (Mill er 1997), each of which has corresponding importance as a s ource of freshwater for human us age. The largest aquifer in Florida and the one that is of pr imary importance to artesian springs is the Floridan (Figure 9). Key points of importance related to an unders tanding of the relationship between groundwater and springs in Florida are that:

• rainfall is the principal forcing function for groundwater recharge in Florida and is relatively independent (for prac tical purposes) of human control;

• artesian aquifers are the primary s ource of water for Florida’s springs;

• artesian aquifers are also the primary source of water used in the state by humans; and

• as rainfall percolates into aquifers, its chem ical characteristics are largely determined by the natural and anthropogenic us es of the overlying lands.

Figure 9 – General map of the extent of the confines of the Florida aquifer and the locations of the principal artesian first magnitude springs in Florida (from Fer nald and Purdum 1998).18 These points result in two principal connections between humans and the quality and quantity of water in Florida’s aquifers: the first is that human consumptive use draws from aquifer storage and has lowered potentiometric head (the forci ng function for spring disc harge); the second is that cultural land use activities may result in in creased quantities of nutri ents and other chemical constituents in water delivered to the aquife rs. Human-induced changes in groundwater quantity and quality have an indirect, but potentially significant effect on spring aquatic habitats, which obviously are aquifer dependent.

Evidence of the relation between groundwater w ithdrawals by humans and spring discharge is illustrated by long-term data for Rock and We kiwa Springs groups near Orlando (WSI 2007a). Figure 10 depicts measured and modeled discharg e from Wekiwa Springs for the period from 1959 to 2005. The measured spring discharge data show a general decline in spring discharge, apparently resulting from a co mbination of increased consumptive use (withdrawals) and declining rainfall totals over the corresponding time period (Inte ra 2006). The dashed line in Figure 10 corresponds to the minimum annual mean discharge for Wekiwa Springs as adopted by the St. Johns River Water Management District in 1991.

Evidence for the impact that land use activity ha s on aquifer water quality has been documented as well (Katz et al. 1999, Munch et al. 2006; Cohen 2007). Of primary consideration for this study are nutrients, principally nitrogen in the form of the compound nitrate. Nitrate is a naturally occurring component of aquatic ecosystems, and nitr ogen is one of the essential elements for plant life. However, nitrate is readily utilized by a quatic plants and algae, and in excess can lead to the development of nuisanc e aquatic plant problems (Rabalais 2002; see Chapter 4). Table 1 provides an example of this type of land use analysis for Silver Springs in which modeled and actual data for springshed ni trogen loads and loads a ppearing at the spring boil show fairly good agreement. Other studies ha ve “fingerprinted” specific nitrogen sources (animal waste, inorganic fertiliz er, septic tank, etc.) based on ra tios of stable nitrogen isotopes appearing in the spring vent s (Phelps 2004; Toth 2003).19 y = -0.3124x + 688.31
R 2 = 0.4654

50 55 60 65 70 75 80 85 90 95 100

1950 1960 1970 1980 1990 2000 2010

YearAnnual Mean Flow (cfs) Intera 2006 (Predicted)

Measured

Minimum Annual Mean Flow = 62 cfs

Figure 10 – Time series plot of annual average flows (measured and modeled) at Wekiwa Spring compared to the existing St. Johns River Water Management District minimum annual mean spring flow of 62 cfs estab lished in 1991 (from WSI 2007a).

Concentrations of nitrate vary widely among the different springs and sometimes even amongst vents within large spring groups , but in a large percentage of springs, current measurements indicate nitrate concentrations are significantly elevated over those observe d in springs contained in undeveloped springsheds (Champion 2001, Munch et al. 2006). This correlation between developed land use in a springshe d and nitrate concentration of groundwater contained therein suggests a cause and effect relationship (Jones et al. 1997, Katz et al. 1999, Cohen 2007). Katz et al. (1999) utilized isotopic analyses to show that substantial por tions of nitrate nitrogen found in the Upper Floridan Aquifer and in spring discha rges are derived from anthropogenic activities such as fertilizer application for agriculture and residential uses, livestock waste, and human
waste.20 Table 1 . Observed and simulated spring flows and n itrogen loads at Silver Springs, Florida, for the period from 1957 through 2055 based on land use /land cover analysis from historic aerial photographs and estimated loads and resulting sp ring boil total nitrogen concentration based on 2055 projected land uses (Munch et al . 2006).

Year Observed Spring

Flow MODFLOW Simulated Spring Flow Observed Nitrogen

Load (lbs/yr) Land Use / Land Cover Model Est. N Load (lbs/yr) Observed Spring Nitrogen Concentration (mg/L) Land Use / Land Cover Model Estimated N Concentration (mg/L)
1957 640.0 716.5 94,416.0 399,054.10 0.10 0.38
1979 778.0 710.8 814,898.6 802,633.06 0.71 0.76
1995 720.0 708.3 955,962.0 1,036,198.93 0.90 0.99
2005 680.0 687.6 1,057,606.7 1,120,813.63 1.07 1.10
2055 N/A 687.6 N/A 1,760,000.00 N/A 2.02

Similar work has been reported for Silver Sp rings (Phelps 1994, 2004), Rainbow Springs (Water and Air Research 1991), Wekiwa Springs (MACTEC 2007), and Wakulla Springs (Chelette et al. 2002). Figure 11 illustrates estimates of th e contributory nitrogen sources for two spring systems, Wakulla Springs south of Tallahasse e and Wekiwa Springs near Orlando. In the Wakulla springshed, human-derived wastewater s ources (Tallahassee sprayfield, on-site septic tank systems, and wastewater residuals disposal ) are estimated to contribute about 61% of the total nitrogen load at Wakulla Springs (Chelette et al. 2002). While atmospheric sources are estimated to provide 26% of the load to the springs, other research has found that those contributions are elevated when humans clear fore st land and replace it with pasture and grass covered land uses (MACTEC 2007). For the Wekiwa springshed, residential fertilizer applications (20%), agricultural fertilizer appli cations (26%), and septic tanks (22%) represent the largest input categories, while atmospheric deposition was estimated to contribute 2% of nitrogen loading (MACTEC 2007). These two examples illustrate the variable nature of the source of nitrogen loading by springshed.

Anthropogenic Forcing Functions

As the discussion above reveals, indirect anthropogenic inputs to springs can be measured from the influence they have on nutrient delivery vi a atmospheric deposition and groundwater supply. Indirect anthropogenic impacts on springs may also result from changing downstream conditions through construction of dams and weirs, st ream channelization and sedimentation, and harvesting or otherwise manipula ting fish and other wildlife popul ations that normally include the spring and spring run in their home range.21 Figure 11 – Top pie chart illustrates th e relative contribution of to tal nitrogen load to the springshed of Wakulla Springs for the period from 1990-1999 (Chelette et al. 2002); bottom pie chart illustrates the relative contribution of to tal nitrogen load to the springshed of Wekiwa Springs for 2004 (MACTEC 2007).

Direct anthropogenic inputs or forcing functi ons to Florida’s arte sian springs include: recreational activities that trample or dislodg e sediments and submersed aquatic vegetation; aquatic plant management activitie s (principally in th e form of physical, chemical, or biological22 control of non-indigenous or “ undesirable” plant and animal spec ies); modification of spring habitats by construction that reduc es connectivity (bulkheads, shor eline modifications, damns, or weirs); and general modificati ons to food webs by introduction of non-indigenous species and reductions in top level consumers thr ough hunting, fishing, and habitat alteration.

Nutrient loads to the groundwater represent an indir ect anthropogenic forc ing function at the regional level while nutrient loads delivered via stormwater runoff to sp rings are an indirect factor at the local scale.

ENERGY/MATERIAL STORAGES IN SPRINGS

Spring Basin/Spring Run

Spring basins (and their various analogues: boil or pool) and spring runs are lotic aquatic habitats whose size corresponds to the volume of groundwater being discharged. Typical spring basins are round or oval, as is commonly observed with solution features in karst geology areas and are connected to their associated spring run. Spri ngs have been classified based upon geomorphic features, and can be further subdivided as spring s, sinks, river rises, or karst windows (Wilson and Skiles 1989). As karst features, spring pool s (or vents) may initially form through catastrophic collapse, but their overall genesis may require millennia. Exceptions include springs discharging from karst fissures which underlie su rface drainage rivers such as the Suwannee or Santa Fe Rivers. The smallest spring basins can be less than 3 m (~ 10 ft) in diameter while the largest spring groups ( e.g. , Wakulla, Rainbow, Silver) have ba sins approaching 90 m (~ 300 ft) in diameter. The bathymetric profiles of spring ba sins are variable, but a typical spring basin has gradually sloping sides which des cend to the vent/s like a funne l. Because of this sloped gradient, spring basins are susceptible to the deposition of sediments from upland erosion. Depths of spring basins rarely exceed 15 m (~ 50 ft) and are commonly less than 5 m (~ 16 ft), although the vent/s of the springs often connect to a network of underwater caves, tunnels, and passages which can reach depths greater than 100 m (~330 ft). The unde rwater structure of many of Florida’s largest spring systems have been explored a nd mapped, none more extensively than Wakulla Springs (Figure 12).

Spring runs begin as a direct extension of a spring basin, drain accord ing to surface elevation gradients, and typically join with other surface water features. Spring runs can be defined as streams which derive 50% or more of thei r base-flow from groundwater sources (Scott et al. 2004). In the karst-rich portion of Florida, ma ny streams receive substantial base flow from springs. The size of spring runs corresponds to th e volume of water they discharge, ranging from a few meters to 100 m (~330 ft) in width, while leng ths range from tens of meters to tens of kilometers for the largest spring runs ( e.g. , Silver, Rainbow, Wakulla, Wekiva).23 Figure 12 – Three-dimensional diagram of the Waku lla Cave and Spring System in Wakulla County, Florida (from L oper et al. 2005).

Spring flows are created by a difference in poten tiometric head between the groundwater and the receiving water. A portion of the po tential energy contained in this water is converted to kinetic energy as water discharges from a spring vent, determining the discharge rate and the resulting flow velocity. Given that the size of a spri ng run corresponds to the quantity of water it discharges, water residence times are typically s hort, in the range of hours, and preclude the development of measurable plankton abundance (Hynes 1970).

Spring flow can expel floating macr ophytes such as water lettuce ( Pistia stratiotes ) and water hyacinth ( Eichhornia crassipes ) that might otherwise block the light essential for photosynthesis in submersed primary producers. Uninterrupted spring flow provides a nearly continuous source of essential elements ( e.g. , oxygen, carbon, hydrogen, nitrogen, phosphorus, potassium) and lessens boundary layer effects that might otherwis e limit critical gas exchanges (Westlake 1967).

Stream velocity is also related to sediment composition (Butcher 1933, Hynes 1970; see Chapter 4). Favorable sediments for macrophyte growt h, such as sandy clays (Power 1996) may be scoured away at velocities > 0.30 m/s (Hynes 1970) and sand substrates begin to give way to gravel and large rocks at stream velocities of 0.60 m/s or greater (Butcher 1933). When spring24 flow is reduced by natural or anthropogenic factor s, scouring velocities are not maintained and spring basins have been observed to fill with coarse and fine sediments (Scott et al. 2004). Reestablishment of spring basin morphology may re quire considerable time if spring flows are restored.

High water clarity often facilitates littoral zone s that include the entire spring and spring run, although other factors including human recr eation, herbivore grazing, stream velocity, sedimentation, and riparian shading commonly in fluence the distribution and abundance of plant life (see Chapter 4).

Water and Dissolved Constituents

Arguably the most important structural com ponent of a spring (other than its basin morphometry) is its water. In a flowing spring system, there is a continual supply of inorganic and organic materials in both dissolv ed and particulate forms. This is beneficial to many forms of life, both autotrophic and hete rotrophic, and provides the basis for the observation that some springs have been found to be among the mo st productive ecosystems on the earth (Odum 1971). Water discharged from artesian springs associ ated with the Floridan Aquifer typically has slightly basic pH, is commonl y under-saturated in dissolved oxygen and over-saturated with dissolved carbon dioxide, and contai ns a variety of elements in trace amounts. Spring water is typically low in inorganic particulate matter as we ll as colored dissolved organic matter, with the exceptions of river rises , whose discharges are principally composed of surface waters reemerging after their entry in a sinkhole or swallet (Butt et al. 2007). Water is a natural solvent and through its movement, provides a mechanism of transporting substances. Consequently, nitrate produced by natural processes or depos ited by anthropogenic activ ities on surface lands and waters can be transported to groundwater and subsequently di scharged to spring ecosystems.

The first and most commonly utilized classificati on for Florida springs is based on magnitude of discharge as initially descri bed by Meinzer (1927), subse quently utilized by Rosenau et al. (1977), and continues to be used today (Table 2).

Table 2 . Spring classification by magnitude of discharge (after Meinzer 1927).
Magnitude Average Discharge First 100 cfs (64.6 mgd) or more Second 10 to 100 cfs (6.46 to 64.6 mgd) Third 1 to 10 cfs (0.646 to 6.46 mgd) Fourth 100 gpm to 1 cfs (448 gpm) Fifth 10 to 100 gpm Sixth 1 to 10 gpm Seventh 1 pm to 1 gpm Eighth Less than 1 gpm

cfs = cubic feet per second; mgd = million gallons per day; gpm = gallons per minute; pm = pints per minute; 1 ft 3 = 0.0283 m 3 ; 1 gal = 3.785 liter; 1 pint = 0.473 liter25 Because the groundwater discharged from spri ngs is qualitatively influenced by surrounding geology, soils, and land surface activities, the re sulting differences in water chemistry have allowed springs to be classified and grouped based on their predominant ions (Whitford 1956, Slack and Rosenau 1979, Woodruff 1993). Using clus ter analysis of the six predominant ion (calcium, bicarbonate, magnesium, su lfate, sodium, and chloride) concentrations for 170 springs, Woodruff (1993) arrived at four distinct spring types and thei r percent occurrence): calciumbicarbonate springs (76%), sodium -chloride (salt) springs (12 %), mixed calcium/salt springs (8%), and low ion (softwater ) springs (4%, Table 3).

Table 3 . Ranges of ion concentrations for the six mo st predominant ions ( calcium, bicarbonate, magnesium, sulfate, sodium, and chloride) from 170 Florida springs which were utilized to create four distinct spri ngs groupings via cluster anal ysis (from Woodruff 1993).

Spring Group Ca +2 (mg/L) HCO3 - (mg/L) Mg +2 (mg/L) SO4 - (mg/L) Na + (mg/L) Cl - (mg/L)

Low-ion 0.7 – 4.8 1 – 20 0.2 – 2.6 0.3 – 7.6 0.8 – 8.2 1 – 15

Calcium bicarbonate 1 – 74 1 – 255 0.2 – 20 0.2 – 78 1 – 17 1 – 21

Mixed 32 – 167 13 – 396 10 – 77 12 – 477 6 – 127 10 – 235

Salt 40 – 600 76 – 205 20 – 546 43 – 1683 160 – 4400 240 - 9400

Data sources: USGS, SRWMD, and SJRWMD

Primary Producers

Primary production in spring ecosystems is domin ated by photoautotrophic life forms which use energy derived from sunlight to fix or ganic matter from inorganic compounds. Chemoautotrophic bacteria are also capable of generating organic mate rial from inorganic sources, and may exist in anaerobic springs, but their ecology or relative importance is unknown. Additional sources of organic ch emical energy to spring ecosyste ms are largely derived from litter fall from surrounding riparian vegetation. The amount of these allochthonous inputs varies widely, depending on site specific conditions and seasonal cycles . The variability of these organic matter inputs must be co nsidered in regard to the tr ophic structure and food webs supplemented by these inputs.

The relative clarity of spring water facilitates the primary productivity and growth of autotrophic organisms at a variety of scales, from microscopic attached algae to macroscopic vascular plants. Most visible in many of Florid a’s springs are the macrophytes, a grouping that includes vascular plants, bryophytes, and macroscopi c and colonial algae. The va scular aquatic plants have internal transport structures (xylem and tracheids ), are principally angios perms (flowering plants with enclosed seeds) evolved fr om land plants, and are comprise d of relatively few species (< 1%) in comparison to the divers ity of terrestrial vascular pl ants (Sculthorpe 1985). Bryophytes lack xylem and true roots and are repres ented by mosses and liverworts (Stream Bryophyte Group 1999), with native Florida spring genera including Brachelyma and Fontinalis

.26 Aquatic macrophytes can be furthe r subdivided by their growth ha bits as submersed, floating, or emergent. Common native aquatic vascular plant species incl ude strap-leaf sagittaria ( Sagittaria kurziana ), tapegrass or wild celery ( Vallisneria americana ), southern naiad ( Najas guadalupensis ), coontail ( Ceratophyllum demersum ), a variety of pondweeds ( Potamogeton
spp. ), and fanwort ( Cabomba caroliniana ). Non-indigenous aquatic vascular plant species are present in many Florida spring eco systems and include hydrilla ( Hydrilla verticillata ), Eurasian milfoil ( Myriophyllum spicatum ), Brazilian water weed ( Egeria densa ) water hyacinth ( Eichhornia crassipes ), and water lettuce ( Pistia stratiotes ).

The macroalgae are distinct in that they lack a vasc ular system or true root s, and are composed of individual cells that ar e independent and undifferentiated. M acroalgae that occu r in springs and spring runs include the genera Chara and Nitella which have vascular plan t like appearances, as well as filamentous genera principally from the green and cyanobacteria groups of algae such as Cladophora

, Enteromorpha

, Lyngbya , and Vaucheria . A survey of 29 Florida springs by Stevenson et al. (2007) found that almost all springs ha d macroalgae growing in them, with an average of 50% of the spring bottoms covere d by macroalgae; many algal mats achieved thickness greater than 0.5 m, with seasonal variation in macroalgae biomass largely related to turbidity and sedimentation from flood events. (Stevenson et al. 2007). Although the two most common reported taxa of macroalgae were Lyngbya wollei and Vaucheria spp., 23 different macroalgal taxa were observ ed in the spring survey.

Collectively, macrophytes provide a physical st ructure upon which smaller primary producers can attach, thereby preventing their downstream export and facilitating access to a better light environment. These epiphytic communities cons ist of a myriad of microalgae, diatoms, cyanobacteria, and microbes that have variously been labeled as biofilms, periphyton, or as aufwuchs (German for growth) and Mattson et al. (1995) reported that spring run periphyton communities are typically dominated by epiphytic di atoms. This epiphytic community has high rates of primary production which can eclipse the productivity of macr ophytes. Odum’s (1957b) quantification of the trophic structure of Silver Springs observed that approximately two-thirds of the primary production was derived from the epiphytic community, even though the strap-leaf sagittaria (Sagittaria kurziana) macrophyte community represented about three quar ters of the standing biomass.

Productivity of submersed vascular plants in spring ecosystems can be high as well. The annual cycle of biomass and productivity of wild celery ( Vallisneria americana ) was studied in Kings Bay, (Citrus Co.), using a modified hole-punching technique and app lied to shoot density data to obtain estimates of production (Hauxwell et al. 2007). Areal leaf produc tion ranged between 3.6 and 23.0 g/m 2 /d of dry biomass, with highest productivity observed in late spring/summer, with an annual total leaf produc tion estimate of 2.7 kg/m 2 ; placing the productivity of Vallisneria in Kings Bay above most literature values for other Vallisneria populations as well as global averages for seagrasses (Hauxwell et al. 2007). Examples such as these demonstrate the high levels of primary production in cl ear-water spring ecosystems.

The lotic nature of spring ecosystems would appear to preclude the formation of a self-sustaining phytoplankton community. However, stream a nd spring run waters can contain autotrophic27 material made up of protists, algae, and cyanoba cteria which have origin ated from the sloughing of benthic or epiphytic grow ths (Odum 1957; Allan 1995; Munch et al. 2006). This observation was supported by Swanson and Bachmann’s research (1976) in Iowa streams, which reported that the average daily amount of chlorophyll exported downstream in the summer was highly correlated (r > 0.9) with the to tal standing crop of chlorophyll on the stream bottom, as well as with the bottom surface area upstream of the samp ling point. A similar pattern was observed in three Florida spring-fed streams, with an increase in chlorophy ll concentration with distance downstream, likely the result of cumulative increases in benthic supply of autotrophic material (Frazer et al. 2006). Within Florida spring runs and streams, pseudo-plankton (also known as tychoplankton) appears to derive largely from sloughing of periphytic a nd benthic autotrophic communities, and, as a result, the species compositi on is likely to be similar. Although perhaps not a true phytoplankton community, this auto trophic material represents a steady downstream export of chlorophyll and primary production available to primary consumers.

Primary Consumers

Primary consumers in spring ecosystems are repres ented by herbivores and omnivores. Vascular aquatic plants are directly cons umed by manatees and turtles like the herbivorous Florida cooter ( Pseudemys floridana) and the omnivorous yellow-bellied slider ( Trachemys scripta) . Snails are also important primary consumers, with larger species such as the Florida applesnail ( Pomacea paludosa) consuming vascular plants and smaller species such as members of the Elimia genus consuming periphyton. As noted in the prev ious section, the highe st rates of primary productivity are derived from the epiphytic comm unity, which, in turn, supports a diversity of aquatic macroinvertebrate primary consumers. I nvertebrate densities in spring ecosystems can be substantial (Table 4), with important orders being shrimp -like crustaceans (Amphipoda), midges (Chironomids), aquatic worms (Oligochaeta), larval flies (Diptera), snails (Gastropoda), mayflies (Ephemeroptera), caddisflies (Trichoptera ), and aquatic beetle larvae (Coleoptera). Mattson et al. (1995) reported that macroinvertebrate communities are typically dominated by chironomids, ephemeropterans, and trichop terans. Observations made by Munch et al. (2006) found chironomids, trichopterans, and lepidopteran s dominated collections made from Silver Springs and suggested that the daily chironomid emergence is correlated with dominance of strap-leaf sagittaria (Sagittaria kurziana) and much lower in areas dominated by flamentous attached and benthic algae.

Other primary consumers include freshwater muss els which are found in Florida springs, but not at the diversity or abundance observed state wide, perh aps a function of lower plankton availability in spring runs. A survey of 16 major Florida springs documented 18 species of freshwater mussels, of which the most comm only occurring was the non-indigenous Asiatic clam, Corbicula fluminea (Walsh and Williams 2003). In th at study, a general increase in the abundance of mussels downstream of the spring ve nt was observed. As noted earlier, there is often a pattern of increasing pseudo-plankton (tychoplankton) downstream in spring runs, and together these observations suggest that susp ended material may limit the abundance and distribution of filter feeders in these ecosystems.28 Table 4 . The mean number (and std. dev.) of orga nisms collected from vegetated substrate in Mill Pond Spring, a feeder spring of the Ic hetucknee River (from Woodruff 1993).

Subclass / Order Number of Organisms (average #/m 2 ) Standard Deviation

Oligochaeta 1,051 1,135

Hirudinea 584 903
Diptera 633 573
Ephemeroptera 24 73

Odonata 195 230
Lepidoptera 24 73
Trichoptera 925 1,132
Coleoptera 97 292
Amphipoda 26,767 17,540
Isopoda 2,360 3,332

Spring ecosystems, which have a navigable conn ection to coastal waters, are utilized by manatees as a place to feed and as a thermal re fuge during cold seasons. Manatees are obligate herbivores whose annual average feeding rates have been estimated at 33.2 kg (73 lb, fresh weight) of vascular plants per day (Bengtson 1983), although consumption rates are likely influenced by animal size and reproductive stat us. Manatees have b een observed feeding on a variety of submersed and floating vascular plant species, but do not appear to select for or measurably feed upon filamentous algae.

Although aquatic plants have co-e volved with herbivores and shoul d be able to recover from cropping by herbivores, when consumption by he rbivores continually exceeds vegetation production rates, overgrazing and loss of vegetation can result. Non-indigenous snails may cause overgrazing, with possible dest ructive avenues being diverse f eeding abilities such as those possessed by the channeled applesnail ( Pomacea canaliculata ), or due to the very high densities observed in smaller species, i.e. 10,000 Melanoides tuberculatus per square meter in the St. Johns River (Thompson 2004). Seasonal loss of su bmersed vascular plan ts appears to have occurred in some spring ecosystems due to grazing by manatees (Jacoby et al. 2007); particularly during winter seasons when manatee densities are highest and reduced light levels correspond to lower plant productivity. Major reductions in th e abundance of submersed vascular plants and a shift from vascular plants to non-grazed fila mentous algae has been reported as a result of manatee feeding at Kings Bay, a spring group whic h forms the headwaters of Crystal River in Citrus County, Florida (Hauxwell et al. 2004).

Obligate herbivorous fish are ra re in Florida springs as most species which consume plant material are also general omnivores and are likely ingesting a mix of detritu s and invertebrates. Mullet ( Mugil spp.) is a good example of an omnivorous fish common to Florida spring systems linked by navigable connection to marine waters. Other largely herbivorous species include the non-indigenous sailfin catfish ( Pterygoplichthys multiradiatus and P. disjunctivus ), the sucker29 mouth catfish (Hypostomus sp.), and blue tilapia ( Oreochromis aurea) , which have become common in several spring systems, including Rock, Wekiwa, Volusi a Blue, and others in the St. Johns River drainage. These non-indigenous fish species are undergoing a range expansion following their introduction from a variety of acci dental and purposeful ev ents in state waters beginning in the 1960’s. While non-indigenous fi sh can also be found in stream and lake systems, the thermal refuge provided within sp ring ecosystems may facilitate their expansion northward in Florida. Sailfin or suckermouth ( Pterygoplichthys and Hypostomus spp. ) catfish were not observed in a 1992 survey of fish fauna from seven springs contained in St. Johns River drainage basin (Woodruff 1993), but have been obser ved in several of these same springs since at least 2005 (Davis and Herring 2005, author’s observations).

Higher-Level Consumers

Faunal studies of Florida springs are perhaps best represented by fi sh collections, many of which are maintained by the Florida Museum of Natural History. In a review of museum data for occurrence of fishes in Florida springs, Wa lsh and Williams (2003) documented more than twenty thousand fish specimens collected from Florida springs representing no less than 99 species from 65 springs. These species represent approximately three quarters of the 126 native fishes that occur in Florida freshwaters and illustrate the collective importance of spring ecosystems as fisheries habitat (Gilbert 1987). Within individual springs , fish diversity is influenced by spring size and it s connectivity with other draina ges. Silver Springs, which appears to have the best historic fishery records available, has relatively stable fish species diversity, as evidenced by the following: Hubbs a nd Allen (1943) reported thirty-five species in the 1940s, Walsh and Williams collected tw enty-nine species in 2002, while Munch et al. (2006) observed thirty-three sp ecies in 2004-05. Munch et al. (2006) pointed out th at while the number of fish species utilizing Silver Springs was simila r to historical observations, the loss of channel catfish and mullet, presumably as a conseq uence of the downstream construction of the Kirkpatrick/Rodman Dam, has resulted in a major decline in overall fish biomass.

Other higher consumers found at variable population densities, but relatively unknown importance in springs include: a variety of bird species ( e.g. , ibis, herons, egrets, cormorants, wood ducks, osprey, hawks, owls, and bald eagles), reptiles such as water snakes and alligators, and mammals such as otters, raccoons, and of course, humans.

Decomposers

Detritus and dissolved organic matter (DOM) from macrophytes and a llochthonous material forms an important reduced ch emical energy pool available fo r consumption by a range of invertebrate detritivores and microbial decomposers. The recycling of energy and matter in spring ecosystems is complemented by the detrital pathway. Bacteria and fungi are two broad functional groups which utilize organic de tritus as a primary energy source.30 Spring Biomass Pyramid

H.T. Odum was the first published researcher to apply general system ecology principles from aquatic to spring ecosystems (Odum 1957b, Kemp and Boynton 2004). During his multi-year project researching over 20 Flor ida artesian springs, Odum noted that the biomass of spring trophic levels ( i.e. , producers, primary consumers, seconda ry consumers, etc.) forms a classic “pyramid”, with lower biomass and numbers of living organisms at th e higher trophic levels (Figure 13). This figure illustrates the predominance of strap-leaf sagittaria ( S. lorata = S. kurziana ) and associated periphytic algae as the bi ological flywheel of the Silver Springs ecosystem. Average standing stock (dry weight bi omass) of these plants was estimated by Odum as about 809 g/m 2 . Herbivores, including turtles, midges, snails, and the herbivorous fish mullet ( Mugil cephalus ) are the next step up on the biomass pyr amid with an estimated average biomass of only 37 g/m 2 . Higher trophic levels include primar y and secondary consumers such as carnivorous fish, alligators, and numerous fish-eating birds such as the double-crested cormorant ( Phalacrocorax auritus ), anhinga ( Anhinga anhinga ), herons and egrets (Ciconiiformes). The biomass of the primary consumers was estimated as 11 g/m 2 while the top consumers had an estimated biomass of only 1.5 g/m 2 . An important parallel metabolic track occurs in the detrital food chain that includes decomposer organisms such as bacteria, aquatic insects, and crayfish. Continuing work by Odum and other research ers in general systems studies found this relationship to be the typical at all scales for adapted systems (Odum 1983).

Figure 13 – An example of the pyramid of biomass for the Silver Springs community (from Odum 1957b).31 ECOSYSTEM-LEVEL PROCESSES

The utilization of measured dissolved oxygen ch anges to estimate whole ecosystem metabolism in flowing aquatic ecosystems has been implem ented in only a limited number of Florida springs by Odum (1956, 1957a, 1957b), although current data for several spring systems have recently become available (Munch et al. 2006, WSI 2006, 2007b), as well as metabolism estimates from a variety of interstate and international streams ( e.g. Naiman 1983, Marzolf et al. 1994, Young and Huryn 1999, Mulholland et al. 2001).

Community metabolism can be conveniently and rapidly assessed for entire spring ecosystems by measuring the mass of dissolved oxygen at upstream and downstream stations. The availability of data sonde s that collect and store in situ measurements of dissolved oxygen and temperature, as well as quantum sensors to do th e same for photosynthetica lly available radiation has greatly improved the ability of researchers to generate unattended single or multiple station data for the estimation of community metabolism. A suitable spring run segment is identified and equipped to provide a record of hourly di ssolved oxygen concentration, percent saturation, and temperature conditions. Di ffusion is estimated from knowle dge of percent saturation and average flow velocity. Oxygen mass transpor t and rate-of-change are calculated from continuous flow records, hydraulic residence time, and dissolv ed oxygen concentration changes. Five components of community metabolism can be estimated using this methodology:

• Gross primary productivity (GPP), the total pr imary productivity of the entire submersed photosynthetic community, includi ng all algae and macrophytes;

• Net primary productivity (NPP), the net differe nce between gross primary production and community respiration equal to the organic (fixed) carbon availabl e at the base of the aquatic food web;

• Community respiration (CR), the dark respir ation or oxygen consum ption of the entire aquatic community, including microbial popu lations, algae, macrophytes and fauna;

• P/R ratio, ratio between gross primary produc tion and community respiration, indicating autotrophic vs. heterotrophic stat us of the aquatic ecosystem;

• Photosynthetic (or GPP) effici ency, the conversion efficiency of useable PAR to gross primary production.

Primary Productivity

Primary productivity can be defined as the rate of formation of or ganic material from inorganic carbon by photosynthetic activity and through the photosynthetic generation of oxygen, can be measured as a rate value representing the conv ersion of solar energy to reduced chemical energy (Bott 2007). In the process of creating this organic material, some energy is lost through respiration (R a ), while the resulting formation of organic material is often defined as net primary productivity (NPP), and the combination represents gross primary productivity (GPP). This can be represented as:

GPP = NPP + R a (Eq. 1)32 Gross primary productivity (GPP) is the total prim ary productivity of the entire submersed plant community, including all algae a nd macrophytes, and typically expr essed in terms of the amount of oxygen produced per unit area and unit time, ( i.e. , g O 2 /m 2 /d). Net primary productivity (NPP) is the net difference betw een gross primary production and community respiration and is equal to the organic carbon avai lable at the base of the aqua tic food web. Da ily dissolved oxygen rate of change measurements integrate a number of ecological fu nctions and are readily convertible to units of en ergy (Reeder and Binion 2001).

While there are primary producer bacteria which generate organi c mater without solar inputs ( i.e. chemoautotrophs), they do not generate oxyge n and their importance in Florida spring ecosystems is unknown but believed to be minor . Therefore the meas urement of primary production in spring ecosystems is based on the equation describing photosynthesis:

6 CO 2 + 12 H 2 O → 6 O 2 + C 6 H 12 O 6 + 6 H 2 O (Eq. 2)

The rate of photosynthesis can be measured vi a changes in dissolved oxygen (or alternatively carbon dioxide) typically in units of mass per area per time, such as g O 2 /m 2 /d. The utilization of dissolved oxygen changes to measure community metabolism in streams was pioneered by Odum (1956) in his study of Florid a springs. It was based on the theoretical considerations that oxygen is introduced into the water as a result of photosynthetic prim ary production during the daylight hours; oxygen is consumed by living or ganisms (bacteria, animals, and sometimes chemical oxidation); oxygen is diffused into air or water depending on the saturation gradient; there can be a potential influx of oxygen throug h accrual of groundwater and surface drainage inputs (see Figure 14). Therefore, changes in dissolved oxygen c oncentrations (C) are a function of photosynthesis (P), community respiration (R), atmospheric diffusion (D), and corrected for the accrual (if needed) from discrete water inputs from surface or groundwater inputs:

C oxygen = P – R ± D + A (Eq. 3)330123456789105/15/05 0:00 5/15/05 6:00 5/15/05 12:00 5/15/05 18:00 5/16/05 0:00

Dissolved Oxygen (mg/L)UpstreamDownstream-1-0.500.511.55/15/05 0:00 5/15/05 6:00 5/15/05 12:00 5/15/05 18:00 5/16/05 0:00

DO Rate-of-Change (g/m 2 /hr)CorrectedUncorrected-1.0-0.8-0.6-0.4-0.20.00.20.40.60.81.01.25/15/05 0:00 5/15/05 6:00 5/15/05 12:00 5/15/05 18:00 5/16/05 0:00

Corrected DO Rate-of-Change (g/m 2 /hr)050010001500200025005/15/05 0:00 5/15/05 6:00 5/15/05 12:00 5/15/05 18:00 5/16/05 0:00

PAR (umol/m 2 /s)AirPlant LevelAir = 39.30 mol/m2/dPlant Level = 28.52 mol/m2/dGPP = 8.45 g/m2/dCR = 8.40 g/m2/dNPP = GPP- CR = +0.05 g/m2/dPAR Eff. = GPP / PAR = 0.30 g O2/molFigure 14 –Example determination of ecosystemmetabolism based on upstream-downstreamdissolved oxygen data from SilverSprings (from WSI 2007b).34 Community Respiration

Community respiration (CR) is the oxygen c onsumption of the entire aquatic community (ecosystem), including microbial , floral and faunal populations. Respiration is reported as community respiration because there is no practi cal way to differentiate between the metabolism of heterotrophs, such as microbes, invertebrate s, and fish; and that of autotrophs. Community respiration occurs throughout day and night peri ods and it is assumed that nighttime CR equals daytime CR. Community respiration can be strongly affected by the biogeochemical oxygen demand of the microbial community, which can rapidly increase followi ng inputs of organic matter such as riparian litterfall or aquatic plant detritus follow ing herbicide applications (WSI 2006).

Because respiration can not practically be differe ntiated between autotrophs or heterotrophs, it is collectively termed community re spiration and is commonly expr essed on a daily basis as CR 24 , and which adds some uncertainty in the estimati on of GPP and NPP (see e quation 1). Hence, net primary production (NPP) may be more accurately termed net daily metabolism (NDM, or net ecosystem productivity or net community productivity) or the net change in oxygen concentration through biol ogical activity, thus:

NDM = GPP - CR 24 (Eq. 4)

Productivity/Respiration Ratio

The ratio of gross primary production to commun ity respiration (P/R) was introduced by Odum (1956) as a simple index of the relative dominan ce (or balance) of auto trophic and heterotrophic metabolism in stream ecosystems. As such, the P/ R ratio was intended to classify ecosystems as either autotrophic (P/R > 1) or heterotrophic (P/R < 1) depe nding on whether ecosystems were net producers or net consumers of organic matter. Heterotrophic state (community respiration) and autotrophic state (gross primary production) are related (Dodds 2006, as an ecosystem with high autotrophic activity will likely have high heterotrophic activity and certainly have high community respiration; however, an ecosystem with more heterotr ophic activity will not necessarily have more autotrophic activity ( e.g. , the condition obtained with high biological oxygen demand loading). An asset of P/R ratios is that they provide for consideration of enrichment by organic carbon in addition to n itrogen and phosphorus, while recognizing that stream food webs can be based on autotroph ic or heterotrophic organisms (Dodds 2006).

Photosynthetic Efficiency

Photosynthetic (or gross primary productivity, GPP) efficiency represents the efficiency of conversion of light energy to organic matter through the photosynthetic activity of primary producers. Photosynthetic effi ciency considers the light en ergy that primary producers can utilize, commonly referred to as photosynthetica lly active radiation (PAR), which has wave lengths between 400 and 700 nm (Kirk 1994) and expres sed in units of light (quanta, moles, or Einsteins) per unit area and time. As gross primary production is typically expressed in terms of35 the amount of oxygen produced per unit area and unit time, ( i.e. , g O 2 /m 2 /d), the resulting photosynthetic efficiency term can be expressed as g O 2 /mol of PAR.

Community Export

Community export refers to downs tream losses of organic carbon fr om the spring system. Plant matter (algae and macrophyte biomass) is typical ly lost from attached plant communities and flows downstream, partially forming the basis for downstream food webs. This downstream export can be readily estimated by sampling two components (WSI 2006, 2007b):

• Floating macrophytic and algal/plant materials

• Fine particulate export of micros copic plant and algal materials.

Pollutant (Nutrient) Assimilation

Flowing aquatic ecosystems are highly produc tive (Odum 1956), and hence have tremendous potential to productively assimilate nutrients a nd other pollutants within definable concentration ranges. Effectively assimilated nutrients are eith er used by the aquatic ecosystem to enhance its own productivity and food-chain support functions, are permanently buried in sediments, or are lost to the atmosphere (nitrogen only). Net nut rient assimilation rate can be estimated based on upstream and downstream nutrient concentrations , discharge rates, and effective volume and surface area of the aquatic system (WSI 2006, 2007b) . Detailed process studies are essential to understand and quantify gross nutrient flux in aquatic systems (see Chapter 3).

Ecosystem-level metrics of spring (and stream or aquatic) ecosystems may provide essential methods for assessing impairment of designated bene ficial uses. Three diffe rent but interrelated measures of the whole ecosystem function for spring (and stream) ecosystems are: community metabolism; community export; and nutrient assi milation rates. When quantified, they can be compared to reference sites and have been found to be sensitive to pertur bations and pollutants.

Secondary Productivity

Consumers convert the chemical energy contained in their food into their own biomass in a process called secondary productivity and the effi ciency of this transfer of energy up trophic levels has been termed ecological efficiency (Lindemann 1942). In other words, ecological efficiency is a measure of the amount of energy in the biomass that is produced by one trophic level and subsequently incorporated into th e biomass produced by a higher trophic level; on average only 10 percent of energy is transferred with the balanc e lost through respiration and heat loss (Ricklefs 1990).

Based primarily on his research at Silver Spri ngs and other Florida sp rings, H.T. Odum (1983) hypothesized that community metabolism as measur ed by gross primary productivity or by total community respiration is maximized in adapted ecosystems. Odum’s hypothesis was named the “Maximum Power Principle” and its importan ce becomes increasingly clear as researchers attempt to explain the complex cause and effect relationships of multiple stressors in spring36 ecosystems. Knight (1980) tested this hypot hesis at Silver Springs and found that the in situ periphyton productivity of mesoco sms in Silver Springs was in fact maximized when consumer densities were approximately the same as those that naturally occur in the spring ecosystem. Ecosystem productivity was found to be lower when consumer densities were lower and higher than this apparent optimum.

In his metabolism measurements at eleven Flor ida springs, Odum (1957a) found that most were operating at similar ecological efficiencies, or in ot her words, their abilities to effectively utilize solar inputs to fix carbon ( i.e. , primary producer biomass) were approximately equal. Until proven otherwise, it may reas onably be assumed that spring ecosystems cannot consistently exceed these observed ecological e fficiencies. This assumption is borne out by observation that the gross primary productivity values measured in relatively unaltered sp rings are at the same approximate level as gross primary productiv ity documented in other highly adapted and subsidized ecosystems ( e.g. , coral reefs, tropical rain forests, fertilized agricultural operations, etc.).

The importance of the Maximum Power Principl e for assessing nutrient effects in spring ecosystems cannot be overstated. All significant bi ological activity in a t ypical spring ecosystem is dependent upon autochthonous productivity. Ultimately the number and secondary productivity of macroinvertebrates, amphibians, reptiles, fish, bi rds, and mammals is dependent on the efficiency of the conversion of sunlight to reduced carbon by photosynthetic plants living in the spring flow. If in their natural state, sp rings are indeed already adapted to run at optimal efficiency, then any perturbation that upsets the natural balance of the system will reduce the gross primary productivity (power) of the spring fl ora and thereby lower th e life-support capacity of the spring. Work by Knight (1980) confirmed th at this general principle (also known as the “subsidy-stress hypothesis”, [E. P. Odum et al. 1979]) operates in Silver Springs in situ mesocosms when densities of primary consumer s (snails) and s econdary consumers (small fish) were altered from their normal, adapted levels. Ch apter 6 pursues this line of reasoning further in evaluation of the proximal and ultimate eff ects of nutrient increas es on spring ecology.

HUMAN AND AESTHETIC USES

The use of springs by humans has always been important, although their primary uses have likely changed over time. In prehistoric times, springs would have represented important locations for high quality drinking water and as areas that would have attracte d prey resources. Based on these attributes, some early human settlements were a ssociated with springs. The use of springs for drinking water supply still continues, comm only as a source for bottled water. Human usage and the resulting direct impacts are often critical in the health of spring ecosystems. Important components of human usag e include: the total number of visitors, the number of visitors actually entering the springs and spring runs for wate r dependent activities (wading, tubing, kayaking, canoeing, swimming, sc uba diving, or snorkeling), and number of visitors for other non-water depe ndent activities (picni cking, hiking, and bi rd watching). The human component includes the inputs of money sp ent to visit and utiliz e springs for recreation and aesthetic purposes. In one perspective, maintaining and pr otecting spring ecosystems from degradation ensures that moneta ry inputs are not diminished. Unfortunately, human recreation can alter spring ecosystems by direct damage to primary producers through erosion, trampling,37 cutting, and uprooting of macrophyt es resulting from wading, pow er boat propellers, and boat anchors. Indirect human effects include chan ges in the quality and quantity of groundwater supplying spring ecosystems from land use cha nges and population increases around springs. Ironically, it may be the aesthetic attraction that springs hold for human beings which leads to their subsequent recrea tional over exploitation.

Recreational Uses

During the last century, Florida’s springs have been utilized for recreational activities and, consequently, many springs have been physica lly modified for this purpose. Physical modifications found at a number of springs incl ude the following: seawalls, rip rap, dams and weirs, deposition of beach sand, boardwalks a nd docks, concrete-lined swimming pools, and large covered facilities for boats and people. Br oadly speaking, springs attract users who will be engaged in two types of outdoor recreat ion, those that are water dependent ( i.e. water contact) such as swimming, snorkeling, wading, bathing, tubing, or canoeing, and those that are nonwater dependent but focus upon the spring’s aest hetics; these include sunbathing, picnicking, hiking, bird and people watching, or camping. As Figure 15 illustrates, detailed summer weekend counts at Rock and Wekiwa Springs in Or ange County, Florida, found that there was an approximately even distribution between water c ontact and non-contact activ ities, as well as a demonstration that recreational activities vary by spring due to site-speci fic characteristics such as water depth and flow velocit y. Unique recreational or tourism features exist at many springs as well; Silver Springs is noted for its gl ass bottomed boat tours, Homosassa Springs its underwater “fish bowl” viewing ch amber, Ichetucknee Springs is a noted tubing destination, and Volusia Blue Springs is noted fo r winter time viewing of manat ees. The one generalization that can be made concerning springs today is that all of the larger springs (first magnitude) are managed as parks, either by private or public entity.38 Wading
12.91%

Bathing
15.33%

Relax
50.99% Nature/Photo
0.00%

Picnic 1.12% Sunbathing
3.45% Fossiling
1.26% Swimming
0.04%
Snorkeling

1.04% Tubing
13.86% ROCK SPRINGS (KELLY PARK) (8/11/07)

Wading
2.50%

Bathing
40.22% Relax
49.46%

Snorkeling
1.84% Sunbathing

1.52% Swimming
0.49% Nature/Photo
0.02%

Floating
3.96% WEKIWA SPRINGS (8/12/07)

Figure 15 – Water dependant recreational activity by category for Rock Springs (Kelly Park) and Wekiwa Springs State Park on August 11 th and 12 th , 2007 (WSI 2007a).

Visitation Seasonality

Visitation has seasonal characteristics, as most water dependent recreation takes place during the warmer seasons when spring waters can provide a refreshing experience. One study of eight springs (Silver Glenn, Silver, Alexander, Apopka, Bugg, Ponce de Leon, Gemini, and Green) in the St. Johns River Water Management District of Florida documented that these springs39 experienced peak use in spring (Apr-Jun) and summer periods (Jul-Sep) (Bonn 2004). Comparisons of human use at two springs with ve ry different seasonal patterns are illustrated in Figure 16. The influence of non-contact water recreation can be detected in the results of a study of spring visitation in four springs located in Florida state parks: Ichetucknee, Wakulla, Homosassa, and Volusia Blue. Peak attendance fo r these springs occurred from spring through summer for Ichetucknee and Wakulla, where in-water activities are the dominant forms of recreation and from winter thr ough spring for Homosassa and Volu sia Blue (which had a July peak also), reflecting the infl uence of non-contact recreation using winter tourism, fish observation, and manatee watching (Bonn and Bell 2003). Seasonal recreational shifts are observed at Volusia Blue Spring State Park, where in-water activ ities are prohibited in the fall and winter when large numbers of manatees are present; but are offset by the great number of visitors which come to the park during these tim e periods for manatee observation. Patterns in the weekly usage were reported in a comprehe nsive study of Rainbow Sp rings river recreation (Holland and Cichra 1994), in which weekend use leve ls were twice or more the use levels of the total for the weekdays.

Visitation Numbers

In-water recreation use numbers in springs are dependent on time of year, day of week, and weather, with highest use duri ng summer holiday weekends (Memor ial Day, Fourth of July, and Labor Day). Estimates of annual numbers of Rainbow River users yielded about 33,000 people recreating on the river in 1994 (Holland and Cichra 1994). This relatively low number is an estimate of just those persons engaged in wa ter dependent recreation on the Rainbow River (mostly tubers floating downstream) and did not include persons recreating at the headsprings and associated park where annual visitation is substantially higher. Another study of water dependent recreational users was conducted on Silver Glenn Spring (Pandion Systems, Inc. 2003) in which the main types of recreation observed were swimming, fishing, and boating. Swimmers utilized both th e park’s recreation area and boats to gain acce ss to the spring, but all fishing was done from boats rangi ng in types and size s from small runabouts less than 5 m to houseboats approaching 20 m in length, and many of the larger vessels stay ed in the spring run for multiple days at a time. During summer holid ay events, as many as 200 boats could be found in the Silver Glenn Spring run, an ar ea of approximately 6.1 ha (15.2 ac).40 0 10,000 20,000 30,000 40,000 50,000 60,000 70,000Jan-01Jan-15Jan-29Feb-12Feb-26Mar-12Mar-26Apr-09Apr-23May-07May-21Jun-04Jun-18Jul-02Jul-16Jul-30Aug-13Aug-27Sep-10Sep-24Oct-08Oct-22Nov-05Nov-19

Calendar DayMonthly Total Paid Attendance Paid Attendance (Model)
Traffic Count Traffic Count (Model) Traffic Count Y = 21627 (1+ 0.802 cos [(2 pi() / 365)(t-156.2)])
R 2 = 0.8372 Paid Attendance Y = 12799 (1- 1.04 cos [(2 pi() / 365)(t+159.2)])
R 2 = 0.6925

Paid Traffic
Attendance Count Statistics (#) (#) Average 12,773 21,598 Median 9,265 19,150 Maximum 58,139 63,473 Minimum 107 1,102 Std Dev 11,360 13,445
Count 96 96 Std Err 1,159 1,372

0 500 1000 1500 2000 2500 3000

Jan-1 Feb-20 Apr-11 May-31 Jul-20 Sep-8 Oct-28 Dec-17
Julian DayAverage Number of Visitors (#/day) Overnight Visitors Daily Visitors

Figure 16 – Comparison of human use seasonality at Rock Springs and Volusia Blue Springs (from WSI 2007a and WSI 2003 respectively). Volusia Blue Spring Rock Springs41 Visitation numbers are influenced by the unique f eatures of the spring, as well as other factors such as adequate parking and regulations ( i.e. , no alcoholic beverages a llowed at state parks). The number of people utilizing a representative sa mple of Florida’s springs is available in two detailed socioeconomic reports (Bonn and Bell 2003, Bonn 2004). Bonn and Bell 2003 examined attendance records from 1992-2002 for f our springs-focused state parks each with large numbers of human utilization. Over this 11 year period of record, annual attendance to the Ichetucknee Springs State Park increased from 134,000 people to nearly 189,000 people (41% increase), Wakulla Springs annual visitor attendance increased from 163,000 to 184,000 (13% increase), Homosassa Springs State Park annua l visitor attendance in creased from over 200,000 to nearly 266,000 (33% increase), and Volusia Bl ue Spring State Park an nual visitor attendance decreased from 360,000 to 337,000 (6.4% decline).

In Bonn 2004, visitation numbers for eight different Florida springs in Marion, Lake, and Volusia Counties of Florida were estimated using 2002/2003 data with annual attendance estimates for the smaller spring ecosystem s being 5,776 for Apopka, 2,888 for Bugg, 14,439 for Green, and for the larger springs, 77,233 for Alexander, 57,755 for Gemini, 259,000 for Ponce de Leon, 1,096,000 for Silver, and 39,235 visitors a year for Silver Glenn Springs.

Between these two studies which used data from the years 2002 and 2003, total attendance for the six first magnitude springs (Silver Spri ngs excluded) was approximately 1,092,468 persons, representing an average annua l attendance of approximately 182,000 persons per first magnitude spring. Silver Springs was not included in this average value because as many people visited this one spring as the other six first magnitude sp rings combined, which would have skewed the calculated average visitation statistics.

Excluding river rises and karst windows, Florid a has 24 first magnitude springs/ spring groups, the majority of which provi de public recreation (Scott et al. 2002); multiplying this number by the average number of visitors calculated above results in over 4.4 million potential visitors and when combined with the approximate 1.1 million visi tors to Silver Springs alone, results in 5.5 million visitors per year in 2002/2003 numbers. The Bonn and Bell (2003) data suggest an annual average increase of 1.26% in visitation, which if applied to the previous estimate, results in over 5.7 million visitors to first magnitude springs per year during 2006/2007. Smaller springs (2 nd and 3 rd magnitude) also attract numerous visito rs, sometimes as many or more than larger springs. For example, Ponce de Leon Spring (Holmes County), at tracted an estimated 259,000 visitors in 2002, although visitation numbers for smaller springs are probably in the thousands to tens of thousa nds per year (Bonn 2004).

The estimate used above of applying the aver age attendance figures from six first magnitude springs to total number of statewide first magn itude springs is simplif ied and certainly an approximation. As the detailed studies by B onn and Bell (2003) and Bonn (2004) illustrate, there is much variation from spring to spring, both in visitation numb ers (and economic impact) due to the unique characteristics of each spring and the recreational oppor tunities it presents. Nevertheless, this estimate of the potential number of vis itors has value as it quantifies this likely significant ecological effect, namely the number of visitors (greater than five million per year), that may be utilizing Florida’s springs for recreation.42 Economic Impacts

Bonn and Bell (2003) report direct economic impact s for four major springs-related Florida state parks estimated by visitor spending in 2002, which ranged from nearly $23 million at Ichetucknee Springs to $10 million at Blue Spri ng. Individually, the Ichetucknee Springs and Wakulla Springs had similar estimates of overall spending for the two parks (approximately $22 million). Although Ichetucknee Springs attracts mo re visitors from outside the area, Wakulla Springs’ visitors spend more than those visi ting Ichetucknee Springs; on average $89 a person per day is spent at Wakulla Spri ngs and $34 at Ichetucknee Springs (2003 dollars). At the lower end of economic impact were Homosassa Spring s and Volusia Blue Spring with total spending estimates of $13.6 million and $10 million respectivel y in 2002. These estimates were the result of these parks being heavily atte nded by visitors from outside the county with relatively low spending per visitor. In terms of wages and salaries , Ichetucknee Springs generated the highest wages at $5.09 million and Wakulla Springs generated the most employment at 347 jobs (Bonn and Bell 2003).

The Bonn (2004) study of eight different springs revealed that total an nual direct spending by visitors accounted for over $65 million, generate d over $13 million in wages, and created over 1,100 jobs. Average daily spending was estimated by category and varied widely between parks; however, four park average spending by category was 22.9% for restaurants, 19.9% for admission fees, 14.8% for groceries, 13.8% fo r shopping, 13.5% for lodging, 6.4% for evening entertainment, 4.3% for transpor tation, and 4.3% for all other spending. Silver Springs, which has a large theme park associated with the natu ral springs’ environment, accounted for 94% of the total economic impact, while the remainder of the springs contributed only 6%. This illustrates that the revenues, jobs, and economic im pacts generated by Silver Springs has more to do with its development as a tourist site (wit h commensurate gate fees and retail purchase opportunities) rather than simply being one of Florida’s largest springs, and that the economic impact generated by the Silver Springs attraction cannot be directly app lied to other springs. This may be especially true considering springs- focused parks operated by governmental agencies that do not have a pr ofit-focused business plan. Neve rtheless, based on the examples where economic impacts have been estimated, it can be suggested that Florida’s major springs make substantial economic contributions.

Effects of Recreation on Spring Ecosystems

A study of Silver Glenn Spring ob served the majority of the sp ring run (6.1 ha) showed impacts from recreation, with direct or indirect mech anical damage from boat propellers and anchors (Pandion Systems, Inc. 2003). While swimming generally caused little impact on submersed aquatic vegetation, wading caused trampling and areas near the spring boil where visitors accessed the water were devoid of submersed aquatic vegetation, as was the western part of the run where boaters preferred to anchor (Pandion Systems, Inc. 2003). The Holland and Cichra (1994) study also characterized the effects of recreation on the submersed aquatic vegetation in Rainbow River. During periods of low recreatio nal use, the quantity of uprooted and cut drifting plants did not increase as recr eation activity increased throughout the day, but during periods of high use, the quantity of drifting plants did increase with increased recreational activity. It was also noted that very few areas throughout the river appeared to be devoid of vegetation due to43 recreational activity, instead appear ing to be influenced more by natural factors, such as water depth and substrate. Shallow areas were prone to propeller damage; because recreational activity of several different user groups ( i.e. tubers, boats, canoes) occurr ed at the same time, it was difficult to determine which user group was responsible for the observed plant damage. However, observational evidence suggested that the majority of the damage to submersed aquatic vegetation was due to power boats. Th e findings of these two studies (Holland and Cichra 1994, Pandion Systems, Inc. 2003) suggest that the type of recreation strongly influences the effect on vegetation and that power boats have the most poten tial to cause harm. Personal observations indicate that people wading in shallow areas result in the near total absence of submersed vascular plants, while filamentous a nd attached macroalgae can remain and appear substantially more resistant to damage from trampling.

Perhaps the most influential study to gauge the imp act of recreational use (largely tubers) on the submersed aquatic vegetation and benthic hab itats was conducted on th e Ichetucknee River (Dutoit 1979). This study demonstrated a linkage between recreational ac tivities and damage to the ecology of the spring run, prim arily through physical disturbanc e in the form of trampling, uprooting, breakage, and generalized destruction of submersed and emergent aquatic vegetation. Recognizing that the aquatic vege tation stabilizes sediments and forms the basis for the food web of this spring system, park management adopted limits to the number of recreationalists on the Ichetucknee River. Subsequent manageme nt changes to the number and frequency of recreational tubers using the river were impl emented to minimize damage to the submersed aquatic vegetation. This entailed setting the daily number of tubers to 750 in th e upper river and 2,250 at the midpoint tube launch, as well a closing the shallo w upper section from September through May following the peak su mmer use period; such limitati ons allowed for aquatic plant re-growth (Florida Department of Environmental Protection 2000).

Human Use Carrying Capacity

The concept of carrying capacity has been applied to out door recreation as a technique to manage the impact of human use on natural re sources. Various facets of carrying capacity as applied to springs in public use areas includ e: the physical amount of space needed to accommodate recreation; visitors needs such as pa rking or restrooms; th e ecological impacts of recreation on air, water, soil, plant, or animal resources; and the social carrying capacity defined as the level of human use above which the quality of recreation w ould be diminished (Billington 1995). Thus recreational carrying capacity has tw o components, ecological and social, with ecological relating to environmental damage a nd social relating to th e visitor (Shelby and Herblein 1986).

Published case studies of the carrying capacity of Florida’s spring ecosystems have been conducted on the Rainbow and Iche tucknee Rivers and relied on the visitor surveys to characterize maximum usage levels. Generally, perceptions of increased crowding resulted in decreased satisfaction among user s. According to Holland a nd Cichra (1994) visitors’ perceptions of crowding on the Rainbow River i ndicated that 65% of weekend users rated the river “slightly” to "moderately” crowded, and 8% reported "extremely” crowded, while 83% of weekday river users reported the river as “not at a ll” to “slightly” crow ded and 2% reported “extremely” crowded. Perceptions of crowding by vis itors were correlated with self-reported44 levels of satisfaction of thei r springs experience by Holland and Cichra (1994), with 52% of weekend users and 49% of weekday users rating th eir trip was "perfect” or “excellent”, and only 5% of weekend users and 4% of weekday users ra ted their trip as fair or poor. A subsequent study of visitor tolerance for crowding was made in an effort to better gauge social carrying capacity through the perceptions and experiences of 464 recrea tional visitors utilizing the Rainbow River between May and September 1994 (Billington 1995). This study treated perceptions of social carrying capacity as a function of crowding among three user groups: tubers, motorized boaters, and non-motorized boater s. Results suggest that perceived crowding was highest among motorized boaters and lowest among tubers, while satisfaction levels were highest for non-motorized boaters and lowest am ong tubers. The relation between crowding and satisfaction suggests that the soci al carrying capacity for tubers may be approached during heavy use days, but overall, insufficient evidence exists to conclude that soci al carrying capacity has been exceeded for the majority of river us ers at the Rainbow River (Billington 1995).

A final study examining the social carrying capaci ty was provided by Paulauskas’ (2001) research examining the experience of recreationalis ts using the Ichetucknee River in relation to perceptions of crowding . Results were derived from 320 re spondent surveys c onducted on site between May and June 2000. The author reported that if res pondents encountered fewer than expected recreationists, they rated their experi ence as “highly satisfyin g” and that respondents with no expectations or who encountered as ma ny as expected, also rated their experience as “highly satisfying”. However, perceptions of crowding did appear to influence overall satisfaction levels of river re creationists and the study revealed an inverse relationship with crowding on overall satisfaction ratings, in that as perceptions of crowdi ng increased, satisfaction levels decreased (Paulauskas 2001). The results from this study indicate that despite heavy weekend use, when between 50% and 65% of th e respondents rated th e river as crowded, perceptions of social carrying cap acity did not appear to be exce eded. This suggests that for Ichetucknee Springs State Park, a good balance between the need to meet recreational use demands and the need to limit recreational numbers for ecosystem protection has been achieved.

DISCUSSION

Spring and spring runs are complex aquatic ecos ystems with a variety of forcing functions, energy storage structures, and ecosystem processes. These ecosystems are entirely dependent upon the amount and nature of the energies and materials they re ceive from local, regional, and even global systems in which they are embe dded. Far too often, the study of these spring ecosystems has focused on select forcing functions or individual structural components with inadequate consideratio n of the whole. While these species or process-level observations contribute towards our knowledge of spring ecosystem s, their interpretation and applicability is likely to be limited to the spring from which they were gathered and provide little insight into the generalized responses of springs to anthropogenic impacts.

As evidenced throughout this report, dozens of research and monitoring projects have been conducted in Florida’s springs over the past decade. While our knowledge about specific indicator organisms or limited troph ic groups in springs has increased as a result of these studies, they have pointed more to the factors that are different between springs rather than to their underlying similarities. Quantification and comp arison of multiple forcing functions, entire45 trophic level structure, and ecosy stem-level processes such as gross primary productivity and community respiration provides a more holisti c view of springs and holds promise for developing commonalities among the hundreds of diffe rent springs that occur in Florida.

Chapter 6 of this report provides an in-depth revi ew of examples of holistic studies in spring ecosystems, as well as conclusions from that research and recommendations for future research. Increased nutrient inputs in many of Florida’s springs are an observed fact, as described above and in the following chapters. Nutrient inflows to springs are one ecological forcing function that has the potential to significantly alter the form and function of the impacted spring ecosystems. However, for improved understanding and management of springs as natural resources, elevated spring nutrient concentrations mu st be placed into the context of the overall spring ecosystem, including the numerous additional external forcing functions described in this chapter. A holistic approach to providing some quantif ication or estimation of all of the principal forcing functions and key details of spring struct ure and function should be considered in future spring studies.46 REFERENCES

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# 2

Springshed Nutrient Loading, Transport and Transformations

Matthew J. Cohen School of Forest Resources and Conservation University of Florida

Summary A discussion of the effects of nutrients on spri ngs biota and ecosystems of necessity requires an understanding of the manner and mechanism of nutrient delivery. More over, as evidence accumulates for particular consequences of nut rient and contaminant loading, understanding how those loads can be attenuated is crucial. This chapter focuses on nutrient loading to springsheds, with an emphasis on processes of transport and attenuation that are germane to their delivery to springs. It starts by describing the flowpaths a nd hydrogeology of Florida’ s karst systems, with an emphasis on intrinsic vul nerabilities to nitrogen loading in particular. A discussion about land use and land cover change, and the sources of nutrient pollution, follows. Nitrogen loading is by far the most well studied spring contaminant, but examination of other changes that have occurred in springs, including changes in flow, dissolved oxygen, hardness and conductivity that have been documented within an d across springs is informative. Attention is also set on the future of nutrient loading, including the pot ential saturation of phosphorus along subsurface flowpaths, with important implications for the timeline of nutrient load reduction. Finally, the state of our understanding of springsheds as closed systems of water and solu tes is examined, with an emphasis on areas of uncertainty that are relevant questions for future research.

The Literature: What is Known From this review of the existing literature on nutrient loading there are four essential areas for which managers can proceed with relative certainty:

• Complex flowpaths and a wide distribution of residence times characterize the hydrogeology of the Floridan Aquifer . The timeline and source area of water delivered to springs is of critical importance to springs hed management, but these physical attributes are poorly constrained in karst systems. The presence of la rge conduits, uneven limestone porosity, complex potentiometric gradients and shallow relief make the delineation of springsheds an area of active research.

• Karst landscapes are vulnerable to pollu tants from anthropogenic activities at the surface . The implications of this are profound fo r land use planning. In particular, it is clear that the risks of aquifer contamination are greatly attenuated in regions of the state and springshed where regional confining laye rs (e.g., Hawthorn Formation) limit the rate of interaction between surface and Floridan A quifer waters. Nitrogen (in nitrate form) is particularly problematic because it is effectiv ely inert, and therefore readily transported,54 in the low organic matter, relatively high dissolved oxygen environment of the upper Floridan.

• Water quality is changing in dramatic ways . Because of the intrinsic vulnerability of Florida’s karst geology to contaminant transport, this fact is well do cumented. What is perhaps not as well documented is the uneven manner in which water quality is declining. Much emphasis on nitrogen (in nitrate form) enrichment has ensued from as much as 350-fold increases above background concentr ations in some springs; overall, spring nitrate concentrations appear to have increa sed by a factor of 10. In contrast, phosphorus concentrations have not markedly increased, presumably due to high exchange capacity with the limestone matrix. Other water quality parameters have also changed; hardness appears to have increased systematically si nce the 1970’s, perhaps a signal of changing subsurface flowpaths that deliver increasi ngly old groundwater. Similarly, dissolved oxygen concentrations have declined sin ce the 1970’s, in many cases crossing ecologically significant threshol ds (~ 1.0 mg/L); this cause for this decline is unknown, but may also be related to the age of the wa ter emerging at spring ve nts. Finally, regular observation of trace quantities of pharmaceu ticals and other chemical by-products of modern society (e.g., DEET, caffeine) illustra te the degree of connectivity between the groundwater and nearby human activities.

• The preponderance of N pollution appea rs to be from fertilizer sources . There are several reasons to treat this finding as an over-generalization, but most of the accumulated evidence from mass balance co mputations and isot opic tracer studies suggests that mineral fertilizers, and th erefore not septic tanks and wastewater sprayfields, are the pr incipal sources of N pollution. One key caveat discussed, is that the isotope method is relatively poorly constraine d for subsurface flowpaths, which may lead to interpretive errors. More importantly, th e source determination for each spring needs to be considered uniqu e; management of those sources must therefore be calibrated to local conditions.

The Literature: What is Not Known, with Emphasis on Future Research Directions

• What are the temporal dynamics of water age? It is clear from the literature that there remain large uncertainties about the age mixtur e of water discharging from spring vents. Given significant discord between observed dye tracer residence times (days to weeks) and nominal ages (decades), a research progr am to better underst and water age over time seems critical. Decomposition, auto-correlation and cross- correlation analyses of resulting time series should enable improve d understanding of flow regimes and also links with other water quality variables that are strongly relevant to springs ecology (nutrients, salts, dissolved oxygen). Specifica lly, if binary mixing models are useful simplifications of spring vent hydrology, time series will allow decomposition of flows into two components, with in sight into the relative importa nce of each, and changes in mixing with regional climatic trends. These methods need to be applied cognizant that different springs are likely to have differe nt hydrogeologic properties, and inference from one spring may not apply in others. This un certainty has key imp lications on pollutant delivery in karst systems, which can be highl y variable. Pulses of nutrients and other water quality parameters (dissolved oxygen, salts, toxi ns) have been invoked for55 catastrophic shifts in other systems, partic ularly when ecosystem drivers have broadly changed. The presence of such pulses (even long duration anomalies might be considered pulses) in spring systems is unknown, and the implications on springs ecology equally uncertain.

• What is the natural variabili ty in spring flow, and how vu lnerable is this to human appropriation? Ongoing efforts to set minimum flows and levels (MFLs) should consider the role of flow in what are gene rally considered to be water quality response variables. For example, changes in nitrate that are implicated in springs ecological change have occurred along with changes in fish access, vent discharge, and salinity, among other things. Moreover, in advective sy stems, flux is often a better predictor of ecological enrichment than is concentration, and with declining flows, fluxes are affected. Understanding the role of flow both directly and indirectly in spring ecosystems, and linking that to how water has been appropria te is likely to be high value research.

• What are the sources of N in springsheds? While there is accumulating evidence (see above) that the source of N is principally mi neral fertilizer, there remains significant uncertainty about the in terpretation of bulk stable isotop e measurements in complex karst hydrologic systems. First, th ere is a broad omission of N fixation from N budgets (which would tend to lighten the isotopic ratio of overall N delivery). Second, nitrification processes are poorly understood: since that process results in large kinetic fractionation (lightning) and denitrification is negligible in the aquifer, the potential impact on the overall N isotope ratio is both unknown and of crucial importance for understanding flow paths. This is also relevant for questi ons beyond nitrogen delivery, since there is a plausible connection between loading of or ganic wastes and declines in dissolved oxygen. Constraining the isotopic kineti cs in springsheds by observations along flowpaths will likely improve understandi ng of water and pollutant sources.

• What are the causes and effects of wate r quality changes beyond elevated nitrate? The synthesis of the literature on springs is dominated by information about nitrogen, but essentially devoid of similarly detailed studies of other water quality parameters. Where surveys have been done across a broad-spectrum of water quality attributes (Stevenson et al. 2004, for example), attributes such as sa linity, flow, dissolved oxygen and phosphorus were at least as predictive of ecosystem change as nitrogen, despite the dramatic enrichment of N that has been observed. Ther e is a clear and present need to understand how variability in these attributes occurs, and the extent to which that variability is in response to anthropogenic activ ities. For example, si gnificant negative correlation between DO and algal cover exists (Stevenson et al. 2004); there are myriad mechanisms (water age, increased BOD loading) that might lead to reduced DO c oncentrations in the Upper Floridan, but these remain critically understudied. Septic tanks, animal waste disposal systems and sprayfields generate enormous quantitie s of biological and nitrogenous oxygen demand if discharged unoxygenated into the groundwater. Since dissolved oxygen concentrati ons are strongly spatially va riable, have significantly declined across a broad spectrum of springs (Scott et al. 2004), and were predictive of algal cover, understanding th e roles of hydrogeologic and an thropogenic processes in low DO at spring vents is a firs t-order research unknown.56 • What is the assimilative capacity of the landscape and aquifer fo r nutrients and other contaminants? Efforts similar to the springshed nut rient budgeting for watersheds in the northeastern US yield an estimat e of 20% for the fraction of terrestrial N load that is transported by rivers to the sea. In a kars t environment where N movement is essentially unimpeded, conservative estimates of landscap e attenuation are as high or higher (70- 90%), without accounting for N fixation. Even where efforts were made to estimate the load reaching the aquifer (MACTEC 2007), thos e estimates are 6 fold higher than what emerges at the spring vent. Elaborating th e reasons for this discordance, and how it might be affected by development and enha nced by management, should be a research priority. Issues that will n eed to be dealt with are response lags due to aquifer residence times, and the relative importance of surface vs. vadose zone vs. aquifer processes. Moreover, as land development increases, ther e may be loading thresholds that exceed the assimilative capacity of the landscape and lead to el evated load breakthrough. The potential for observing this effect is lo w for nitrogen pollution, but may be high for phosphorus. Ensuring that P saturation fronts are not advancing towards spring vents is important for forecasting.

• How responsive are springsheds to management interventions? As we begin to understand the links between springshed activit ies and declining water quality in spring ecosystems, evaluating the extent to whic h parameters are management-sensitive is relevant to both timelines for recovery and prioritization of management efforts. This is relatively well understood for N (see Obreza 2002 for an assessment of BMPs on N delivery), but not at all for other water quality metrics in springsheds. If other parameters (P, DO, salts, flow) figure prominently in holistic springshed management, understanding how to best manage them is a clear priority.57 INTRODUCTION

The objective of this chapter is to synthesize literature on nutrient load s, transformations and sinks in springsheds, emphasizing processes occurr ing prior to spring water emergence at the boil. After a brief background on nutrient cycles and karst hydrogeology, the chapter discusses sources (loading models, water and nutrien t source tracing, temporal dynamics, spatial dynamics) and then transformations and sinks (abiotic attenuatio n, biotic attenuation, management-inducible attenuation). The chapte r will close with a synthesis of the knowledge stores relevant to springshed management, and the gaps that should be targets for future research.

Nutrient enrichment is a water quality problem with global scope; the eff ects of nutrients on the primary productivity of fr eshwater (lotic and lentic), estuarin e and marine environments has been widely discussed and demonstrated (H utchinson 1957, Schindler 1977, Carpenter et al. 1998a, Rabalais 2002). Moreover, broad scale nutrient en richment - principally resulting from the use of modern inorganic fertilizers but also from fossil fuel combustion and la nd use change - has dramatically altered global biogeochemical cycl es (e.g., doubling the rate of global N cycling - Vitousek et al. 1997) and watershed loading above what is predicted as background levels (Omernik 1977). For nitrogen (N), anthro pogenic influence is particularly pronounced; commercial fertilizer use worldwide now accounts for 140 Tg (10 12 g) of N yr -1 , a figure that exceeds the upper estimate of the quantity fixed naturally in terrestria l ecosystems (Vitousek et al. 1997). N delivery to freshwater and marine ecosystems is prin cipally via riverine transport (Boyer et al. 2002) accounting for ca. 25% of watershed lo ad in Northeastern rivers (Figure 1).

The addition of nutrients is responsible in ma ny cases for changes in the rate of primary production, and therefore indirectly on the co mposition of primary producers as competition constraints are altered (Smith et al. 1999). For example, non-linear shifts in lake primary producers moving from rooted aqua tic plants to phytoplankton have been broadly attributed to P enrichment (Vollenweider 1976, Canfield and Bachman 1981, Wetzel 1983, Carpenter et al. 1998b), with important implications for how the P loading to these lakes is managed. Cascading effects of nutrient enrich ment typically include changes in oxygen and light dynamics of aquatic systems, with deleterious eff ects on animals adapted to pre-en richment conditions (Hynes 1970), and on whole system metabolic processes (Dodds 2006). Genera lly, nitrogen (N), phosphorus (P) or both (Elser et al. 2007) limit ecosystem productivity, a nd as such, the focus of nutrient attenuation efforts.

While the discussion of eutrophi cation in lake system s is relatively mature (e.g., Schindler 1977), the hydraulic circumstances in flowing water system s, such as constant advection to maintain diffusion gradients and turbulent mixing, make ge neralization about eutrophication in all aquatic systems from lake models alone problematic. A new synthesis of eutrophi cation in lotic systems has begun to emerge (Tank and Dodds 2003, Dodds 2006, Hilton et al. 2006) with important but often subtle variations from lentic eutrophica tion, particularly with regard to alternative mechanisms for changes in primary productivity (Caraco et al. 2006). As such, while phosphorus may be the limiting nut rient in lakes (Carpenter et al. 1998a), and nitrogen most frequently the limiting nutrient in marine system s (Smith 2006), the influe nce of nutrients on flowing water ecosystems58 Figure 1 – N inputs (kg m -2 yr -1) vs. riverine delivery (Boyer et al. 2002; see reference for threeletter codes). Loads have increased ~ 20 fold si nce the mid-1800’s. SE USA (red dot) falls near line defined by rivers from NE USA; a global model was identical (Vitousek et al. 1997).

may be more ambiguous (alternating co-limitation or no limitation – Smith et al. 1999, Elser et al. 2007). Consequently, management of both N and P from anthropogenic sources is a critical priority for managers, regulators and scientists worldwide. Aggressive management schemes that limit the quantity or improve timing and uptake of fertilizer applicati ons, enhance or restore the landscape assimilative capac ity, reduce the load from fossi l fuel combustion, and decouple human waste from waterways are needed globally. Evidence has shown that these efforts can work, but the probability of success in mana ging the problem is maximized by understanding stores, fluxes and transformations intrin sic to the system being managed.

Addressing the observed eutrophica tion (principally overgrowth of filamentous algae) in spring ecosystems in Florida requires attention to th e myriad mechanisms and potentially complex dynamics of nutrient loading and pr ocessing. As such, a synthesis of nutrient effects in Florida’s springs starts with a discussion of the s ource and chronology of nutrient loading. The mechanisms and temporal dynamics of nutrient lo ading to spring vents ar e relevant to their effect. Perhaps as importantly, a synthesis of nutrient loads to springs at the same time as consideration of the effects, allows risks and op portunities for future management to be linked. For example, a role of phosphorus in the de cline in spring biological condition has been suggested. Because phosphorus concentration in spri ng discharges has, in general, not increased through time, it has been essentia lly neglected in the literature on springshed management (a point to which this review will return in de pth). The implicit assumption for why P has not increased along with N appears to be that the carbonate aquifer matrix has sufficiently high sorption capacity to attenua te anthropogenic loads. This partic ular assumption is being tested as part of the loading literature (Nair et al. 2004), but has direct relevan ce to synthesis of ecological effects if, as those authors predict, P saturation and breakthrough can occur.59 Emphasis on N loading vis-à-vis other water qua lity parameters has occurred within the scientific and regulatory communities due to stro ng evidence for anthropogenic N enrichment in the karst environment of North Fl orida; evidence for contemporane ous P enrichment is absent. Karst is vulnerable to nitrate-N transport; indeed Katz et al. (2005) and Phelps (2004), among others, use N 2 partial pressure in spring-vent water to conclude that denitrification is negligible, and nitrate behaves like a conservative solute. As such, increasing human development in karst regions inexorably leads to increased N load s without major efforts to manage sources.

Figure 2 - Nitrate load to the Gulf of Mexico from the Suwannee River: a) changes in loads between 2001 and 2005 (driven by changes in water flo ws), b) time series n itrate concentrations at 5 1 st magnitude springs along the Suwannee, and c) estimates of nitrate load by sub-basin. Basins where a regional aquitard is absent (shown in grey) are 25% of the area, but contribute 65% of the load. (Hornsby et al. 2002, 2003, 2004, 2005; Mirti and Mantini 2006).

For example, Figure 2 summarizes recent trends in N loading to the Suwannee River over time; these trends, replicated nearly ubiquitously where springsheds in clude intensive human land uses (Scott et al. 2004), are the principal motivating force behi nd this chapter. Clear upward trends in load are evident, as is spatial identifica tion of sources (i.e., sub-basins shaded grey, Georgia

Florida60 representing 25% of the area a nd 65% of the load). Observed increases in N loads to the Suwannee River are principally attributed to springs and diffuse groundwater (Hornsby and Mattson 1997, Pittman et al. 1997). Nitrate-N concentrati ons of springs, which discharge directly from the upper Floridan Aquifer, have increased substa ntially from near background concentrations ( ≤ 0.1 mg/L) (Kat z 1992, Maddox et al. 1992) to > 5 mg/L at present (Katz et al. 1999).

In general, considerat ion of nutrient loading requires an accounting effo rt. The goal is to tabulate accurately the inputs and outputs of a ke y nutrient for comparisons across catchments or through time. Van Breeman et al. (2002) present an approach for 16 watersheds in the Northeastern United States; for N, they tabulate the important sources and sinks and arrive at nominal relationships between loading and export that are useful for understanding the magnitude and vulnerability of sinks, the im portance of land use and the key areas of uncertainty. Their methodology, and examples of similar efforts for the more complex and confounded environments of karst springsheds, ar e an important component of this report.

Landscape loading and subsurface transport/tran sformations of nutrients are the most well studied aspect of springs, at least with respect to nitrogen. Indeed, while links between nutrients and ecological effects may remain somewhat ambiguous (e.g., Stevenson et al. 2004), the collective understand ing of nutrient loading, te mporal nutrient delivery trends, travel times of water and nutrients, principal s ources of loads, and subsurface transformations is relatively mature. By way of example, the current synthesi s effort identified key literature sources in each of several categories within the general framework of nutrients and Florida’s springs. Figure 3 summarizes the distribution in th e number of references compile d, illustrating the significant focus on sources, and more limited focus on downstream effects.

Figure 3 – Number of publication on each of a variet y of topics germane to nutrient effects in springs identified for this study. This result highlights important k nowledge stores and gaps.61 HYDROGEOLOGY

Hydrology in a Karst Environment

Hydrology in Florida is complicated by two aspects of its geological setting (Berndt et al. 1996). First, the State is flat, leadi ng to anastomosing flow paths, l ong residence times, closed surface basins (isolated wetlands, swalle ts), and dominance of inter-flow (i .e., instead of surface runoff). Second, the geologic basement of Florida is limest one, leading to a karst la ndscape (defined here as the presence of a self-organizing aquifer, in this case made of carbonate). Champion and Starks (2001) suggest that an understanding of karst is a prerequisite to understanding the hydrology and water quality of springs. Ewers ( 2006) suggests that failure to appreciate the complexity of karst aquifers (predominantly by assuming isotropy in hydrogeologic modeling) has led to poor decision making, and wasteful spending on poorly predictive models. The dissolution processes that characte rize the development of karst le ad to subsurface flow conduits, massive groundwater flow anis otropy, and numerous sinkhole and swallet features that link surface flows (streams and rivers) to the groundwater matrix and conduits (Kincaid et al. 2004, Ewers 2006). The resulting complexity of subs urface flowpaths means attribution of hydrologic and pollutant loads to particular areas is le ss clear than for surface drainage networks.

The principal deep aquifer system in the State is the Floridan Aquifer (Figure 4), which covers the entire state, and portions of South Caro lina, Georgia, Alabama and Mississippi. This carbonate aquifer is among the most transmissive ka rst aquifers in the world, making it an ideal water source, and consequently, the source of water for over 93% of Florida’s population (Fernald and Purdum 1998). Protecting the quality and quantity of this resource is of paramount policy importance, and nutrient poll ution to the aquifer is one of a constellation of stressors that are already deleteriously affecti ng it. While affects on the Flor idan Aquifer exist throughout the state, springs do not. The vast majority of springs occur in North and Central Florida, principally because of the proximity of the Floridan to the surface in areas and greater landscape relief.

Several features particular to Florida karst (eogenetic or not d eeply buried; Florea and Vascher 2007) are relevant to nutrient loading to springs. First, eogenetic karst has high matrix porosity (Martin and Dean 2001), which leads to highly accessible storage; the result is that spring discharge is very weakly responsive to short-te rm weather, a condition quite different from the karst regions of Kentucky, Misso uri, Texas and elsewhere around the globe where the karst is more deeply buried. Indeed, an analysis of flow flashiness in 12 Florida springs (Florea and Vascher 2007) suggests that peak flows are less than 5 times minimum flows over 90% of the time, a homeostasis typical of ba salt aquifer springs, not karst. The reservoir from which spring flow derives is extremely large, and as such nom inal residence times are long; the importance of this for the time domain of nutrient delivery is extremely important, and will be discussed in detail later. A second peculiarit y of the Floridan Aquifer karst is that flow through the matrix and conduits is a dynamic interplay driven by hydrologic conditions. Ma rtin and Dean (2001) have observed in the Santa Fe sink-rise system that during low flow, matrix and secondary conduit contributions are high, but that at hi gh flow, major conduits dominate. This has implications for nutrient delivery to springs; during periods of hi gh flow, primary short-term62 Figure 4 – Extent and confinement of the Floridan Aquifer. The Fl oridan is among the world’s most productive karst aquifers and source of most of Florida’ s drinking water (Miller 1990).

Figure 5 – Plan view of Briar Cave (Marion Count y) showing complex anastamosing channels that generally follow matrix fractures (NE-SW), but include caves and lateral connections (Florea and Vascher 2007).63 recharge areas are likely to be the dominant source where surface water loads via a swallet or sinkhole. At low stage, matrix flow is more im portant, and dispersed nut rient loading across the unconfined region is likely to be more predic tive of what emerges from the spring vent.

The matrix that contains the Floridan is not uni form, leading to significant anisotropy in flows. Vertically, limestone deposits of different ages have different pr operties, and harbor water with different chemistry. For instance, in the S ilver Springs region, Suwannee and Ocala are the shallowest and Oldsmar and Cedar Keys are the deepest; corresponding differences in salinity and ion profiles exist. The Uppe r Floridan Aquifer (UFA) refers to water in the Ocala formation which ranges in thickness between 0 and 180 feet; it may be exposed or buried beneath Miocene clay and Plio/Pleistocene sand deposits. Laterall y, anisotropy of aquifer properties is extreme, with high fracture porosity (Mar tin and Dean 2001), and complex anastamosing networks of large and small conduits that m ove water through the subsurface in predictable (Werner 2000)

Figure 6 - Hydrogeologic units in the Silver Spri ngs basin (after Phelps 2004, Scott et al. 2004).64 but still massively complex ways (Kincaid et al. 2004, Florea and Vasche r 2007 – see Figure 5). Cave diving experience (e.g., Fisk and Exley 1976) visually confirms the complexity and anisotropy of the matrix.

The presence of the Miocene confining unit (the interbedded marine sands and phosphatic clays of the Hawthorn Group) significantly affects th e connectivity between surface and groundwater (Scott et al. 2004). Where that confin ing unit is thick and intact, hydrologic and nutrient processes almost exclusively occur in the surface sediments and the surficial aquifer; nutrient migration into the aquifer is expected to be slow. Where the Hawthorn has, over time, been eroded away via surface water action, the Ocala limestone has been exposed, along with the UFA within it. Figure 4 shows th e extent of that exposure (areas in brown are unconfined). The zone of active erosion of the Hawthorn is known as the Cody Escarpment (Cody Scarp), and this divide is hydrogeologica lly critical (Figure 6).

The importance of the Cody Scarp for hydrologic processes is enormous. At and below the scarp, strong groundwater-surfacewater interactions lead to unique f eatures such as the Santa Fe River Sink at O’Leno State Park where the river is captured in its entirety by a sinkhole, emerging 5 km downstream. Below the Scarp wher e the limestone is exposed, prevalence of surface water features (we tlands, stream channels, lakes) drops dramatically and are replaced by spring vents and karst windows where the potentiome tric surface of the Flor idan intersects the land surface. Springs and spring runs are theref ore surface expressions of the level and quality of the aquifer.

Similarly, the importance of the Cody Escarpm ent in delineating z ones with distinct biogeochemical processes is also difficult to ove rstate. As observed by Hornsby (2007), above the scarp where the Miocene Hawthorn Formation precludes surface wate r percolation to the Florida, the surface waters and th e surficial aquifer are generally elevated in dissolved organic carbon (DOC), and depleted in N, presumably because large wetland areas and areas with terrestrial soils that ha ve high water tables, leads to rapid ra tes of denitrification, and variably enriched with P; indeed N:P ratios in these headwater areas where the Hawthorn is intact and land use intensity is low may be less than 1.0. In groundwater near the scarp, P levels climb dramatically due to interactions with the phospha tic clays of the Hawthorn, but then decline due to sorptive processes occurring be tween the phosphate and the CaCO 3 aquifer matrix. However, it should be noted that some springs have natura lly high levels of P (to be discussed at length later) suggesting that P sorption ma y not be uniformly high. Nitrat e concentrations are generally low until water from the upper Floridan Aquifer (UFA) that has been enriched with nitrates (i.e., near areas of intensive human land use) becomes an important component of river flow. Figure 2 illustrates this trend, with loading occurring in areas shaded grey in the unconfined region.

These confined, transitional and unconfined hydrologi c zones (Figure 7) and associated expected nutrient conditions (moderate P/low N, high P /low N, low P/high N, respectively – Hornsby 2007) underscore why spring nutrient enrichment ha s focused on nitrogen. Reasons for intrinsic vulnerability of unconfined landscapes to N enrichment will be described later, but first it is relevant to address how the contributing areas for spring flow have been defined.65 Figure 7 – Spatial survey of hydrogeologic zones in the eastern part of the Suwannee River Water Management District (after Upchurch et al. 2007). The Cody Escarpment defines the boundary between where the Floridan is confined and unconfined. Springs (blue dots) and surface streams (blue lines) are shown.

Typology of Florida’s Springs

Karst springs (i.e., not water table seeps – Scott et al. 2004) occur where the potentiometric surface of the Floridan Aquifer in tersects the ground surface. They are categorized according to median discharge (Figure 8a from Scott et al. 2004 and Rosenau et al. 1977, after Meinzer 1927). The logarithmic partitioning of discharge in to magnitudes partially obscures the dramatic differences in flow between for example a 1 st magnitude (> 65 million gallons per day) and 3 rd magnitude (< 6.5 million gallons per day) sp ring. Comparison of hydraulic and ecological processes between springs of such differing size is potentially problematic, and scaling relationships for generalizing process-level measur ements made at a spring of a particular size are clearly required. Examples of factors expected to be diffe rent for varying spring magnitude are light regime, depth, flow velo city, recreational impacts, and densities of large herbivores.66 Springs occur most prevalently below the Cody Escarpment where the UFA is suddenly unconfined by the Hawthorn Formation aquitard (Wa kulla, Ichetucknee and Silver form in this manner), and near the channels of major rivers (e.g., Santa Fe, Suwannee, Ocklawaha) traversing

a) b)

Figure 8 – a) Nomenclature for spring magnitude ; cfs = cubic feet/ second, mgd = million gallons/day, gpm = gallons/minute) . There are 720 springs (39 1 st , 191, 2 nd , 151 3 rd magnitude). b) Nomenclature for spring types based on where and how water is discharged.

Figure 9 – Locations of Florida’s > 700 springs, includ ing those visited by Scott et al. (2004).67 the unconfined karst region. Figure 9 shows th e location of the springs; the sweep of the Suwannee River is clearly evident from the tremendous density of springs that discharge in that region. Numerous springs also occur at the coast, where the land surface dips and intersects the Floridan; Homossassa springs, Kings Bay, Cha ssahowitzka, Weeki Wachee springs among many others are formed in this manner. Tidal e ffects (Yobbi 1992) can co mplicate the flow and nutrient dynamics of these systems.

There is also a preponderance of sub-aqueous springs, including Blue Springs (currently inundated by Rodman Reservoir), Apopka Springs (discharging to Lake Apopka) and numerous submarine vents (Lambert and Burnett 2003, Slomp and van Capellen 2004). Nutrient effects at the locations of submarine groundwater discharge sites can be important for estuarine ecology (Sigua and Tweedale 2003, Paytan et al. 2006) including the incidence of red tide (Hu et al. 2006). Other springs occur where riverine flows reemerge afte r sinkhole capture and passage through subsurface conduits (e.g., Santa Fe Ri ver Rise); while varying amounts of water emerging in these systems is river water, and Wilson and Skiles (1989) argue that they should not be categorized as springs, Martin and Dean (2001) demons trate that there is indeed substantial flow-dependent mixing with ground water. A typology of springs based on where they occur (Copeland 2003) is presented in Figure 8b.

Springshed Delineation

A springshed is defined as the area over which water falling at the surface (but ultimately traveling through the subsurf ace) will collect at a particular spring vent (Upchurch et al. 2004b). More often than not, the contributin g area for a particular vent or small isolated spring cannot be delineated from catchment areas delineated for a spring complex. For example, the springs of the Ichetucknee complex are fed by water from different parts of the springshed, but only the springshed delineation for their ag gregate catchment is possible given existing potentiometric surface data.

A springshed is most often delineated in the same manner as a surface watershed, with the exception that the terrain boundaries are based on the el evation of the groundw ater potentiometic surface (Shoemaker et al. 2004). This presents seve ral challenges. First a nd foremost is that the “terrain” that defines boundaries (i.e., groundwater levels) is extremely flat, and delineating edges without exceedingly dense well stage data is impossible. Sec ond, the potentiometric surface that is the groundwater terrain is dynamic, depending on rainfall and recharge rates. As such, the boundaries of springs heds tend to be fuzzier to accommodate the interannual variability. Third, the measuremen t of the “terrain” by groundwater wells is less frequent and of lower spatial resolution than would be true for a surface drainage system (i.e., a digital elevation model), so the precision of the delineation is lower. A schematic of a springshed (Figure 10 – after Upchurch and Champion 2004) shows the va rious focal points of springshed delineation and protection. Loper et al. (2005) use similar terminology (p rimary, secondary and tertiary protection zones) for the level of nutrient mana gement recommended to reduce loads to Wakulla Springs. The combined inputs of surface water a nd subsurface drainage are extremely important to parse for purposes of understanding nutrient lo ads arriving at the spring vent (Figure 10).68 Figure 10 – Schematic of springshed with surface water drainage in the confined region, a sinkhole (blue dot) capturing flow after passage into the semi- and un-confined regions, a historic stream channel that is the presumed flo w course of the karst conduit (“trace”), and the spring. Priority areas for water quality prot ection are shown (Upchurch and Champion 2004).

By way of example, Figure 11 shows the delineat ed springshed for the Ichetucknee complex in the Suwannee River Water Management Distri ct (SRWMD) (SRWMD 2002). The dashed line delineates areas where, over some time fram e, water in the subsurface moves towards Ichetucknee Springs; this boundary was for 2002, and would be expected to vary year to year, depending upon rainfall patterns. The presence of confining units (H awthorn Group) in the northern watershed leads to surf ace water drainage networks that clearly diverge from the groundwater basin. For example, the headwaters of Rose Creek (the main tributary draining south from Lake City), Olustee Cr eek (which drains via surface ch annels to the east, and out of the springshed) and Falling Creek (a tributary of the Suwannee, dr aining via surface channels to the north) are in the springshe d, but dominant flowpaths for the latter two are not to the Ichetucknee. This complicates the process of developing a hydrolog ic budget, which also complicates the ability to constr uct a nutrient budget. Moreover, much of the characterization of subsurface flow boundaries is pr edicated on model assumptions ; the size and geography of springsheds appears to be strongly sensitive to th ese assumptions (Shoemaker et al. 2004), with as little as 40% spatial c oncordance between different m odels of the springshed area.

Another mechanism for delineating springsheds is the use of major ion chemistry via a process named hydrochemical facies analysis (Champion and Upchruch 2003, Upchurch et al. 2004a). The central premise of this approach is that po ols of water in the subsurface have major ion chemistry profiles, and since most of these ions ar e conservative (i.e., don’t react, at least in the subsurface), similarities in major ion chemistry between two disparate locations can be used to infer flow linkages between them. Use of disc riminant and principal components analyses on well water samples is then used to identify regions with generally similar and generally distinct hydrochemistry. Hydrochemical facies have be en used to delineate flow paths in the69 Figure 11 – Delineation of the Ichetucknee springshed in May 2002, including Lake City. Note the disappearing streams (red dots are sinks), the historic stream channel (the Ichetucknee Trace) and the movement of surface water from the northern springshed into Olustee Creek to the east and Suwannee River tributar ies to the north (SRWMD 2002).

Ichetucknee springshed (Upchurch and Champi on 2004 – Figure 12), springs typology in the Southwest Florida Water Management District (Champion 2001), and broad groundwater basins in the Suwannee River Water Mana gement District (Hornsby 2007). The two principal benefits of this is method are 1) confirmatory (i.e., vis-à-vis the potentiomet ric delineation) and 2) discriminatory at a scale that the potentiometric method is not. Figure 12 shows the ion profiles for water from various springs in the Ichetuck nee complex, and up-gradie nt wells; concordance between springs and certain wells suggests a hydro logic connection. The in ference is that there are two distinct catchments (1: Ichetucknee H ead, Cedar Head, Blue Ho le; 2: Mission, Mill Pond, Devils Eye). Similar research from th e St. Marks system (Barrios and DeFosset 2006) suggest that four groundwat er flow components compri se the St. Marks Rise. Rose Creek

Falling Creek

Cannon Creek Olustee Creek

Olustee Creek70 Figure 12 – Use of major ion chemistry to delinea te the flowpaths and source areas for particular springs in the Ichetu cknee complex (A, B, C, D, E, F). The shape of the “star” represents the relative concen tration of various ions, and concordance between a spring vent star and well samples upstream (blue stars) suggests a hydrologic conne ction. Clear groupings appear to be A/B/C and D/E/F. These springs are notably different in their nitrate concentration, among other wa ter quality parameters.

Delineation of springsheds has utility for develo ping budgets of water and nutrients for springs. However, several important caveats about deliea ted boundaries should be made clear. First, as already mentioned, certainty abou t flowpaths of rainfall in th e geographic bounds of a delineated springshed is complicated by surface water draina ge that may transport water elsewhere.

Second, where there is a predominance of condu it flow for moving water through the aquifer, some of the water (and therefore nutrients) de livered to the springshed may be transported in conduits that emerge outside the springshed. For example, Toth and Fortich (2002) speculate for the Wekiva that some of the springshed c onduits bypass the spring co mpletely, discharging directly or diffusely to the St. Johns River. Anisotropic flowpaths, possibly orthogonal to contemporary potentiometric gradients, might lead to this unusual circumstance.

Another example of the complex nature of the conduit system co mes from dye trace studies on the springs of the Ichetucknee River. Despite the convergence of water to the river from 9 springs, the conduit systems are not the same for each. A regional conduit system from the Ichetucknee Trace (see Figure 11) appears to load to Blue Ho le, while a more local conduit71 system loads to Mission, Mill Pond and Devils Ey e springs. The latter system apparently moves water to Devils Eye spring (on the west side of the river) from a conduit that also discharges to Mission and Mill Pond (on the east side of the river). As suc h, delineation of the Ichetucknee springshed is actually a composite of smaller springsheds, each with different hydrochemical

Figure 13 - Delineation of springshed (red dashed lin e) and capture zones (500 year travel time – pink line, 10 year travel time - green line) for the Silver Springs springshed (Phelps 2004).

properties, conduit maturity, depth and no minal residence times. Understanding the hydrogeologic setting for a given candidate spring is likely to be extremely important in understanding the potential impacts of nutrients.72 Finally, while springsheds do delineate flow direc tion, they do not imply flow velocity. Shallow potentiometric gradients, flow anisotropy and extremely variable recharge rate s lead to the condition where the time of concen tration (i.e., time required for a molecule of water to move from precipitation to a particular gaged location, in this case, th e spring vent) can vary from hours to millennia. This leads to strongly inco ngruous evidence about the residence times and transit velocities of water and nutrients in the subsurface. Fo r example, Karst Environmental Services (1997) reports, rapid transit times (nominally 8 days) for water moving from Rose Creek (Figure 11) sink to Ichetucknee springs. In contrast, age dating techniques for the water (Katz et al. 2001, Katz et al. 2004, Katz et al. 2005) indicate recharge ages of 20-25 years. The implications are quite significant: if the water is uniformly 20 years old (i.e., piston flow model), then the nutrients and other contaminants emerging today are the signal of loading that occurred in the mid-1980’s. More likely, ho wever, is that the wa ter at the spring vent is a binary mixture of young (days to years) and old (decades to mill ennia); nutrients, calcite undersaturation, and oxygen are all associated with the younger water (Toth and Katz 2006). Moreover, there is evidence of mixing with Lower Floridan Aquifer (LFA) water at sites n ear the St. Johns River (Knowles et al. 2002), which may explain spatial vari ability in nitrates, dissolved oxygen, salinity and calcite saturation (Toth and Katz 20 06). Understanding the time domain of loading is important for interpreting current loads, anticipating load reduction response times, and determining the influence of drought and flood conditions on spring ecology. While efforts to implement dye tracing (Karst Environmental Services 1997, DeFosset 2003, Moore and Martin 2005) are expensive and provide inco mplete pictures of groundwater movement, they are critical.

Given this uncertainty, a key refinement in sp ringshed delineation is the definition of timespecific capture zones. This technique relies on modeling and makes potentially problematic assumptions about subsurface flow, but the resultin g map product is of considerable utility for regional planning. Phelps (2004) produced a map (Figure 13) that shows 10-year and 500-year capture zones for Silver Springs; the proximity of these boundaries under scores the uncertainty associated with the approach. Upchurch and Champion (2004) define priority protection areas for springs water quality in this way; Figure 14 shows this delineation for Ichetucknee Springs.73 Figure 14 – Ichetucknee Springshed’s low, medium and high priority areas for water quality management. Ichetucknee Trace (Spring Protecti on Area) has highest priority due to short capture times and highly pervious soils. Surface Wate r Protection Area delineates areas that drain to surface features captured by sink holes (e.g., Rose/Cannon Creek sinks) feeding spring conduit systems. Lowest prior ity is the Springshed Protection Area, where aquifer confinement and surface drainage minimize risk to spri ng water quality (Upchurch and Champion 2004).

Hydrogeology and Nutrient Loading

The importance of hydrogeology on ni trate loading to surf ace water is exemplified in the Santa Fe River basin, which has the contrasting hydrogeology of conf ined and unconfined regions. Specifically, the aquifer in an 820 mi 2 area of the upper basin is confined and surface water dominates transport, whil e the aquifer in a 546 mi 2 area in the lower basi n is unconfined and the primary transport is via the Floridan. Loads by r each (Table 1) indicate that although the lower basin occupies only 41% of the watershed, it gene rated roughly 98% of the load over the last 5 years (see Figure 2 for spatial partitioning). Wh ile land use activities (crops and improved pasture) and animal activities (chicken and dairy farms, improved pastures), which are the likely sources of nitrate (Lamsal et al. 2006) are expected to load at similar rates in both areas, the intrinsic capacity of the landscape to denitrify is dramatically different, resulting in a massively elevated contamination risk in unconfined aquifer regions.

Table 1. Nitrate-N loads (kg/km 2 /yr) generated by the two reaches of the Santa Fe River watershed during the year s 2001-2005 (Data sources: Hornsby et al. 2002, 2003, 2004, 2005; Mirti and Mantini 2006). Santa Fe River 2001 2002 2003 2004 2005 Upper Basin (confined) 4 3 29.6 15 56 Lower Basin (unconfined) 475 581 696 1128 130274 Assessing Aquifer Vulnerability

In a setting like the karst landscape of North Fl orida, water quality vulnerability is far from uniform; the example from the Santa Fe Rive r basin underscores the role of geology in regulating contamination risk from nitrogen. Across all c ontaminants, the degree of confinement, the rate of recharge, the amount of rainfall, the position of conduits, the proximity to springs, the intensity of land use, and the efficacy of manage ment controls figure prominently in understanding the mass and rate of delivery to the aquifer and springs. This has necessitated the use of predictive models to help prioritize hot-spots. Hot-s pots in Florida conform with areas where the aquifer is unconfined and land use is inte nsive. It is important to note that the model was not developed with sufficient density of karstic calibration site s to be of pred ictive utility. However, this map underscores the sorts of methods used to unders tand the spatial distribution of risk and therefore guide management.

To account for both known multiple causes of aquifer water quality decline, and inherent lack of understanding of karst flowpaths, the Florida Ge ological Survey (FGS) developed the Floridan Aquifer Vulnerability Assessment (FAVA) based on a weight-of-evidence synthesis of risk factors (Arthur et al. 2007). Among the key drivers of that model are hydrogeology and confinement; the result ing map product (Figure 15) conforms closely to environmental geology. Similar local efforts (e.g., Wekiva A quifer Vulnerability Assessment – Cichon et al. 2005) refine these maps for local planning purposes.

Figure 15 – Relative vulnerability of the Floridan Aquifer System to pollution from overlying land uses (from Arthur et al. 2007). Area s in white are outside the study bounds.75 Timescales of Water an d Nutrient Fluxes

Times of flux between recharge and emergence at a spring vent represent a source of significant and important uncertainty, with direct relevanc e to understanding current and predicting future nutrient concentrations observed in springs. B ecause the drainage network consists of subsurface pathways through stores with unknown eff ective volume, it is generally not possible to use conventional hydrograph tec hniques to determine watershed convergence properties. Moreover, the reported times of concentration based on modeling and dye trace studies (e.g., Phelps 2004, Karst Environmental Services 1997) vary over 5 orders of magnitude (days to centuries) within a given springs hed. Phelps (2004) further repor ts two estimates of water age from the same spring made 3 months apart that ar e 1-5 years and >50 years associated with large observed shifts in spri ng discharge. Katz et al. (2004) suggested that rech arge of recent origin coming from shorter groundwater flow paths and/ or from features th at are hydraulically connected to the upper Floridan A quifer (e.g., sinkhole) could contribu te a larger fraction of total spring discharge during high flow conditions.

There are two complimentary lines of evidence that are widely used to determine the timedomain of connectivity between particular points in the lands cape and the outflow of spring vents. The first is an aggregated method ba sed on the concentration of chronometric tracers; typically these are gases whose concentration has varied in well described ways in the atmosphere (Figure 16). Assuming these gases ar e unreactive, and that no gas exchange occurs once water has recharged, concentrations of thes e gases in water is indicative of the time at which it was recharged (Katz et al. 2001). When this technique is implemented, numerous tracers are used because they re present different time scales (e.g., 14 C has been used for longterm age dating – Toth and Katz 2006), or checks in event of c ontamination. Probably the most promising technique is the combined measurement of 3 He and 3 H (the daughter product), both based on nuclear bomb-testing that peaked in the early 1960’s. The concentra tions of the various tracers then need to be interpreted in terms of how water flows through the system; that is, the computed nominal age of the water varies ba sed on whether piston flow (limited mixing, water moves through the system with other water rech arged at the same time), binary mixing (two water sources at various fractions) or exponent ial mixing (mixing of multiple water ages) is assumed (Katz et al. 2001). Evidence varies for which of th ese flow models is most plausible. Katz et al. (2001) show evidence supporting a piston fl ow model, while Toth and Katz (2006) working in some of the same systems, argue fo r a binary mixing model of moderately old water (pre-1940) and young water (ca. 1990). Moreover, the various tracers disagree to a significant extent (see Figure 17), at least with respect to the placement of water age within a fairly well defined range. There is no intrin sic reason to expect springs to be have uniformly with regard to hydrogeology, but site-specific un certainties arising from unknow n flowpaths, residence times and mixing dynamics significantly affect ma nagement strategies and expectations.

The general picture that emerges from the agin g measurements is wate r that, on average, has spent between 10 and 30 years in the subsurface. This estimate emerges regardless of which hydrologic flow model is invoked (Katz et al. 2001). However, if a binary mixing model holds (for which Toth and Katz 2006 pr ovide evidence, and for which nitrogen delivery dynamics are further supportive), then the results from among the earliest efforts to characterize spring water age (Katz et al. 2001) suggest that between 30 and 70% (and in some cases more) of the water at 12 springs was from water with residence times less than 4 years.76 Figure 16 – Concentrations of various tracers (d ecay corrected) of groundwater age as a function of time. Water recharged prior to 1950 is considered tracer free (after Katz et al. 2000).

Figure 17 - Estimated water ages for Florida springs (Katz et al. 2001) using multiple dating techniques. While the measures are generally po orly correlated, they are consistently suggestive of recharge dates between 5 and 35 years before present.

Another method for assessing flow connectivity and rate between surface locations via subsurface pathways is the use of injected fluo rescent dye (Karst Envi ronmental Services 1997, Butt and Murphy 2003, DeFosset 2003, Kincaid et al. 2004, Moore and Martin 2005). In general, while some dye injections are never rec overed over the period of observation, the travel77 times of dye plugs suggest much more rapid flux of water in conduits; for example, a transit time of 8 days was observed between surface flows into Rose Creek sink and outflows at Ichetucknee Springs (Karst Environmental Services 1997). Th is discordance in time domains for subsurface pathways is strongly suggestive th at a piston flow model is not a ppropriate, calling into question the utility of the nominal water age for manageme nt purposes. Indeed, if, as demonstrated in Toth and Katz (2006), there is strong negative covariance be tween nutrient, dissolved oxygen (Katz et al. 1999), and ionic saturati on deficits of water and its age, then it is plau sible that shortresidence time flowpaths dominate nutrient deliver y. This generates dramatically different expectations regarding the time domain of responses to changes in land use or management practices.

TRENDS IN WATER QUANTITY AND QUALITY

Water Quantity

Flows in springs are generated at a rate of roughl y 1 cfs of flow per square mile of delineated springshed area (Upchurch and Champion 2004). St rong correlations in spring discharge with climatic forcing are observed (Knockenmus a nd Yobbi 2001), but the tim e domain of correlation is apparently over 3-years (that is, flow is correlated with rainfall over the last three years). Chelette et al. (2002) report remarkable va riability (5-28 cm/s) in fl ow velocities over 6 months of observations in Wakulla Sp rings; the responsiveness of spring hydraulics to antecedent rainfall conditions is an area of needed reseerch . Interestingly, only a sm all fraction (6-23%) of the incident rainfall on the entire springshe d (both confined and unconfined areas) can be accounted for in spring flow; since evapotranspi ration (ET) accounts for only 67% of rainfall on a landscape basis, a significant fraction of flow (as much as 25%) leaves th e land via unknown pathways, presumably diffuse groundwater flow or submarine discharges. It is important to reiterate here that the uncertainties related to springshed delineation/size and surface runoff routing make closure of springshe d water budgets highly problematic.

Flow matters in springs for a variety of demonstr ated and hypothetical reas ons. Clearly, as flows change, fluxes of contaminants dissolved in the water also change. Hilton et al. (2006) argue for flux rather than concentration as the predictor of biological response. Moreover, the classical evolution of eutrophication in re sponse to nutrient enrichment so well described for lakes (Wetzel 1983) is not as clear in running wa ters (Hynes 1969, Dodds 2006) where biological access to nutrients is constrained in part by reside nce times. Residence time clearly plays a role in the degree to which lakes respond to P en richment (Vollenweider 1976, Schindler 1977), and Hilton et al. (2006) invoke residence time as a critic al regulator of phytoplankton responses in rivers. As flow declines, residence times in the sub-surfaces would be expected to increase (perhaps non-linearly) which may alter the degree to which nutrient effects can be observed in springs. It is worth noting that residence times in nearly all of the ma jor springs are extremely short, despite some evidence for flow reductions (e.g., Munch et al. 2006 in Silver Springs).

Declines in flow can affect both the load and the residence times of nutrients in spring runs, and can also be expected to indir ectly affect biota by changing th e hydraulic environment in which they live. Indeed, Whitford (1956) speculates on the role of flow in the spatial distribution of epiphytes in Silver Springs. The most obvious and alarming change in spring ecosystem primary78 production is increases in filamentous algal grow th, and algal attachment has been shown to respond to shear stresses i nduced by advection (Wade et al. 2002). Effects of flow velocity on animals (e.g., gastropods) are also likely; as flow volumes decline, flow velocities generally decline commensurately. Animals (and indeed plants) for which the hydraulic environment is critical for life history are likely to be a ffected by such changes, with unknown ecological
consequences.

Evidence for flow declines are subject to some un certainty because of climatic variability and a poor understanding of time lags between rainfall inputs and spring discharges. However, long term statistically significant downward trends have been observed at several springs (Silver and Ponce de Leon) and weaker trends observed at Wekiva and Rock springs (German 2003). Notably, however, several springs with significant algal problems near the vent (Silver Glen, Alexander) showed no decline in flow with time. Chelette et al. (2002) provide some evidence that current flows at Wakulla Sp rings (~350 cfs) are markedly lower than historic values, which were over 600 cfs. Champion and Starks (2001) re port dramatic declines in Crystal (Figure 18), Sulphur and Lithia springs between 1937 and 1997, preceding the major drought effects of the last decade; some evidence of d ecline at Rainbow Springs, among the largest springs in the state, is also evident. Garman and Garey (2005) repor t on declines in a submarine spring discharge to the point where, since 1961, no discharge has been observed at all at Jewfish Sink (a submarine spring). Kissingen Springs in the P eace River basin, historically a 2 nd magnitude spring, is another example where discharge c eased in the 1950’s. The cause of this lack of flow due to changes in the aquifer surface has been ascribed to climatic effects, and extraction for municipal and agricultural uses (e.g., weak statistical effects of consump tive use on flow in Knockenmus and Yobbi 2001, strong effects in Weber and Perry 2004). Odum (1957) speculates that the fate of Kissingen is possible for other springs if judicious use of wate r resources is not made a high priority. While it is clear that such a fate ha s not been widely realized despite significant pressure on the UFA for agricultural and municipal supply, it is also clear that flow regulates all processes in springs, and continue d vigilance to protect flow ra tes by avoiding excessive aquifer drawdown is essential.

Figure 18 – Flow at Crystal Springs, Hillsborough County since 1933 (Weber and Perry 2006).79 Nitrogen

The nitrogen cycle is shown in Figure 19. Excelle nt summaries of the behavior of N in the environment exist (Bowden 1986, Rabalais 2002). In all, it is estimated that human-derived sources contribute between 60% and 80% of th e global N deposited on land annually, and that nearly 50% of the N transported in the world’s rivers (~ 36 x 10 12 g N/yr) is of anthropogenic origin (Galloway et al. 2004). Key facets for understanding N fate and transport in Florida are: 1) high assimilatory (plant uptak e) and dissimilatory (denitrificat ion) potential in regions with high water table (e.g., confined regi ons), and 2) negligible sorpti on and denitrification potential once nitrate is in the karst aquifer (Katz et al. 2004). To understand why N enrichment in springs is occurring requires understanding the spatial concorda nce between intensive land uses (e.g., dairies, row crops, poultry houses) and regions where the Floridan Aquifer is unconfined. Water enriched with mineral fertilizers and organi c N spends very little time in the root zone (where plant uptake can occur) and encounters c onditions unfavorable for dissimilatory removal in the shallow UFA, which is generally oxi c (limiting the potential to use nitrate (NO 3 ) as a terminal electron acceptor) and low in availa ble carbon (the electr on donor for microbial respiration).

Figure 19 – Schematic of N biogeochemical cycl ing between various pools (Bowden 1986).

Nitrogen is, by far, the most well studied nutrie nt in springsheds. Of the 174 references on nutrient loading to springs synthe sized for this chapter over 50% dealt principally with sources80 and fate of N. Most of the rest address fl owpaths (dye trace studies, springshed delineation). With the exception of reports that provide r econnaissance data on nu merous water quality parameters (e.g., Scott et al. 2004, Stevenson et al. 2004), only 5 reports address possible effects of other nutrients, so the discussi on that follows focuses predominantly on N load (and the nitrate species of N in particular).

Nitrogen has several forms in an aqueous envi ronment; these include two oxidized inorganic ions (nitrate – NO 3 - , and nitrite – NO 2 - ), reduced inorganic species (ammonium - NH 4 + ), dissolved gases (nitrogen gas - N 2 and nitrous oxide – N 2 O), and myriad organic species of varying lability (collectively referred to as di ssolved organic N – DON a nd particulate organic N - PON). While all species can be found in spring rivers, nitrate dominates delive ry of N at spring vents (Figure 21). A small number of springs (Beecher, Santa Fe in Columbia County, Columbia, Treehouse springs and Holton Creek Rise) are dominate d by ammonium and DON. However, most springs, particular ly those that are severely enri ched, are dominated by nitrate. Temporal trends in other forms of N are unknown. Several propert ies of the nitrate ion make it particularly problematic as a groundwater pollutant . First, it is highly soluble (saturation of nitrate in water occurs at ~300 g/l) and extremely mobile, movi ng rapidly through soil profiles, due to low anion exchange capacity and low affin ity for nitrate; this rapid leaching leads to contaminated groundwater. Second, once nitrate has passed out of the soil/vadose zone, the absence of electron acceptors and anion binding sites in the mineral matrix of aquifers (karst or otherwise) markedly slows nitr ate reactivity; in fact, some au thors have shown that nitrate reactivity in groundwater is so slowed (0.013 μ mol N L -1 d -1 ) that its transport was comparable to a bromide tracer (Smith et al. 2004), and that nitrite production rates (0.036 μ mol N L -1 d -1 ) more than compensated for this attenuation rate. Othe rs have documented enhanced nitrification of dissolved organic nitrogen (DON) and ammonia in shallo w groundwater (where oxygen is present) (Miller et al. 1999), with important implications for N isotopic fractionation (if the process is rate limiting and incomplete ) and also for dissolved oxygen dynamics.

The reason for the high emphasis on N is that obse rved trends in nitrate-N are nearly ubiquitous in areas where anthropogenic loading to the spri ngshed land surface has occurred. For example, water quality monitoring of the major rivers in the Suwannee River Basin over two decades has indicated a statistically significan t increasing trend in the concentrations of nitrate-N (Ham and Hatzell 1996), which is primarily attributed to groundwater disc harges in the Middle and Lower Suwannee reaches (Pittman et al. 1997). Similar trends have b een observed for the Ocklawaha River, particularly where that system interacts most clos ely with groundwater (Mytyk and Delfino 2004). Figure 20 shows nitrate concentration trends for springs in the Suwannee Valley. Figure 21 shows a similar synthesi s across a total of 69 springs (Strong 2004). The x-axis is N concentration in the 1970s and the y-axis is cont emporary concentration. Note that these data are principally 2 nd and 3 rd magnitude springs; upward trends for 1 st magnitude are well established in Katz et al. (1999), Champion and Star ks (2001), Chelette et al. (2002), Scott et al. (2004), Munch et al. (2006), Upchurch et al. (2007), among numerous others. Points above the 1:1 line have seen enrichment; the mean increase was 0. 46 mg/L to 1.04 mg/L as N, but clearly there were some sites at which the increase was far gr eater, and some where concentrations declined. Indeed, Strong (2004) reports a trend of generally unknown cause where several springs increase through 1998 (Figure 21) and then decline to current concentr ations (e.g., Hornsby springs – Figure 22).81Figure 20– Nitrogen loading to major springs of Florida; shown are both nitrate and total N load. Note that for most springs, mostof the N (>75%) arrives as nitrate, but ammonia is important for some that have low nitrate concentrations.82 Figure 21 – Relationship between earl y time period (1907-1979) n itrate-nitrogen and recent time period (1990-2003) nitr ate-nitrogen, with a 1:1 line (- -) , for 69 Florida springs (Strong 2004).

Figure 22 – Nitrate trends in Hornsby Springs (Alachua County) since 1992 illustrating declines since 1998. Conc (mg

‐ N/L)
Name Early Late Wekiva 0.07 0.43 Madison Blue 0.01 1.25 Crystal 0.71 2.61
Owens 0.50 3.64 Ichetucknee 0.36 0.77 Fanning 0.35 3.82 Weeki Wachee 0.02 0.78 Hornsby 0.68 0.44

Overall (mean) 0.43 1.1383 Figure 23 – Trends in nitrate concentrations in springs of the SRWMD (Upchurch et al. 2007).

A synthesis of recent (1989-2006) records for springs (n = 68) in the SRWMD (Upchurch et al. 2007) suggests that signif icant trends occur in both direct ions (positive = 55% and negative = 25%), with some springs exhibiti ng no trend (20%); these are re ported spatially in Figure 23. The picture that emerges is complex, and suggest s that recent control efforts may be working in some places and not others. The prevalence of ne gative trends may also be driven hydrologically (see Figure 22) with recent drought and flood c onditions clearly manife st in the signal at Hornsby Springs. Toth and Katz (2006) obs erved a negative tr end between nitrate concentrations and age, suggesting that as spri ng vent flow became older (i.e., deeper source water during low flow conditions) nitrate concentrations declined; Einsiede l and Meyer (2006) observed the identical inverse re lationship for karst springs in Germany. This formal link between discharge and age is an important rese arch need because of its implications for regulatory minimum flows and levels.

Analysis of the concordance between nutrient lo ading and flow may be confounded by hysteretic (i.e., path dependent) behavior in the deliv ery mechanisms. Van Herpe and Troch (2000) observed counter-clockwise hysteresis in flow vs. nitrate concentration relationships in a Belgian river (Figure 24), suggesting that nitrates arri ve during the recessional phase of the storm hydrograph. Current monitoring data are insufficient to dete rmine if similar patterns are observed in Florida springs ; autocorrelation length in nitrate concentrations is apparently less than the sampling interval for both springs and rivers (Cohen, unpublished data). As such, the prevalence and incidence of occurrence and eff ect of pulses due to storm flows like those observed for dissolved organic carbon by Katz et al. (2004) in Wakulla Springs are unknown. Moreover, potentially non-linear ecosystem effects due to pulsed nutrient delivery are possible if storm flows are missed by cu rrent sampling frequency.84 What is clear from the time series record, at leas t at Wekiva Springs, is a strong cross-correlation of discharge and concentration with essentially no evidence of lag. This observation (Figure 25) confounds interpretation of water age dating techniques that assume a piston flow model. While flow could reasonably be expect ed to respond to up-gradient climatically-driven changes in potentiometric surfaces due to nearly instantane ous transmission of pote ntiometric pressure (Martin and Dean 2001), the concordant arrival of nitrates with extremely short lag means that travel times need to be far shorter than the years-to-decades estimates widely published (Katz et al. 2004). A binary mixture of young and old water with a nominal age of 20-25 years, but comprised of flow with markedly shorter reside nce times (Toth and Katz 2006) is more likely, and, as discussed above, supported by evidence from dye trace studies. The amplitude of the variability suggests that young wate r carries the bulk of the nitr ate, an observation partially confirmed by negative cova riance between age and nitr ate concentrations (Katz et al. 1999). It is also plausible that the positive cross-correlation be tween flow and nitrate concentration is due to proximate sources only; if this is true, the amplit ude of variability in Figure 25 would imply that these sources are a large fraction of the total load. Frydenborg (2006) observed that most of the load arriving at the Ichetuckne e was entrained by the sinkholes that capture confined region flow. Chelette et al. (2002) report concentrations in the Wakulla System, where nitrate concentrations increase dramati cally between Middle River Sink ( one of several key conduits to the vent) and the spring. Indeed , upward trends clearly observe d in the concentrations at Wakulla are not matched by those in the sink, sugges ting that either the nitr ate is delivered from diffuse sources between the sink and the spring, or that the vast majority of the nitrates are from other smaller conduits. Loper et al. (2005) cite that the load from the confined to the unconfined portion of the Wakulla springshed is 73,000 kg N (o f which some may be attenuated), compared with 270,000 kg N at the spring vent; this supports the cont ention that most of the N is introduced in the karstic short-time capture zones.

Figure 24 - Flow vs. nitrate hysteres is (van Herpe and Troch 2000).85 Figure 25 - Time series measurements of NO 3 -N and discharge in the Wekiva Springs.

Given the potential for disproportionate importa nce of local sources, th e observation (Upchurch and Champion 2003) in the Ichetucknee springshed of nearly monotonic increases in NO 3 concentrations with proximity to the trace and springs is important. In that study, distinct areas, mostly near the spring and tracer, were geosta tistically characterized as having UFA nitrate concentrations in excess of 2 mg/L, which is mark edly higher than levels observed in any of the springs in that complex.

Phosphorus

In most freshwater systems, P is limiting. Indeed Hutchinson (1957) suggested that for examining the impacts of nutrients on lakes, ecol ogists should focus efforts on understanding P. That prediction has been prescient across a wide array of aquatic systems, but most notably for lakes (Vollenweider 1976, Carpenter et al. 1998b). Odum (1957) f ound no association between P levels in springs and their primary production, concluding that the mechanisms that lead to P limitation in lakes (Schindler 1977 ) are not presen t in flowing water systems like springs. The same conclusion regarding eutrophication in flow ing waters was arrived at regarding nitrates (Odum 1957, Duarte and Canfield 1980), suggesting that nitrogen may not have significant ecological effects in springs; th is represents probably the mo st important unknown in springs ecosystems ecology.

The phosphorus cycle (Figure 26) is in some ways simpler than N (no gaseous phase, few valence states), but is made complex by the numerous geochemical reactions that involve phosphates. The principal source of P in the modern landscape, both in Florida and elsewhere, is mineral fertilizer. Notably, however, the source of P fertilizer is generally geologic, including the vast Hawthorn Formation deposits in Florida (e.g., Bone Valley depos its in central Florida supply ~30% of the worlds P). The presence of this massive and concentrated geologic source of P means that many ecological systems receiving Fl oridan Aquifer water are loaded heavily with P. Indeed the nominal concentrations of P in springs are between 50 and 100 ppb, which86 suggests that historic N:P ratios (m olar basis) were less than 1:1. Note, however, that N:P ratios in advective (flowing water systems) are only in formative regarding nutrient limitation if some other input (e.g., light) is not limiting.

Figure 26 – Generic P cycle. The P cycl e in Florida is characterized by low sorption in the Plio-Pleistocene sediments of the confined regi on, massive stores of geologic P in the Hawthorn Formation, and almost inexhaustible sorption capacity in the carbonate aquifer matrix.

There are several reasons related to loading that explain why P limitation may not persist in springs. First is the presence of apatite (CaPO 4 ) in the Hawthorn Formation which weathers relatively easily. P weathering is expected to be most pronounced at the Cody Escarpment, and streams flowing over the scarp and then into th e Floridan are generally P enriched (Hornsby 2007). The concentration of total P reaching Ichetucknee springs (60-80 ppb) is markedly less than the concentration in Rose Creek Sink (200-600 ppb), one of several principal accumulation points for runoff from the confin ed region of the springshed (Frydenbourg 2006). This suggests that most of the P is sorbed during passage th rough the Floridan Aquifer towards the spring, and that the P in the fertilizer applied to lands overlying the unconfined aquifer is not reaching the springs. The presence of geologic P somewhere along the flowpath is the likely explanation for the relatively high P levels observed in several springs compared, for example, with lakes and wetlands (Figure 27). Toth and Katz (2006) obser ved a significant negative relationship between water age and phosphate concentration, consiste nt with P enrichment by surface waters interacting with the Hawthorn and then followi ng shallow flow paths to the spring vent.

A second factor regulating P loadin g to the springs is the strong chemical sorption potential of the aquifer matrix for mineral P. While P br eakthrough is expected to occur under extreme P loading (Nair et al. 2004), the enormous sorption capacity of the calcitic soils and the carbonate aquifer are likely responsible fo r P loading trends that are star kly different than comparative N loading. Indeed, at many springs with strong upward trends of nitrate, the TP trends are stationary (Figure 28).87 Figure 27 – Summary of trends and nominal con centrations in 23 springs comparing observations in the 1907-1979 period and observa tions in 1990-2003. Nominal concentrations are low (~50 ppb) and no significant trend is observed (Strong 2004).

Figure 28 – Summary of TP concentrati ons in some of Florida’s 1 st magnitude springs, many of which have seen significant recent in creases in nitrate concentrations . TP concentrations are, in general, low and unchanged. (Data from Scott et al. 2004; figure from Loper et al. 2005)

Carbon

Organic matter dynamics in the subsurface are outs ide the scope of this review, but several key observations should be made. First, as in su rface streams, microbial processing and trophic88 concentration of energy occur; Simon (2000) s howed interesting similarities between surface stream organic matter processing and the processi ng that takes pl ace in karst c onduits, noting among the several differences the relative absenc e of nutrient limitation, presumably because of the ratio of flux to production. In addition to subsurface heterotr ophic respiration which mineralizes both particulate and dissolved orga nic carbon, there are geochemical processes that result in abiotic attenuation. It is worth pointin g out that the self-organizing character of the karst solution conduits occurs in part because of interactions between su rface sources of organic acidity and carbonate dissolution.

The carbon emerging in spring wate rs is, with few exceptions, almo st exclusively inorganic. Nominal organic carbon concentr ations across a broad survey of well samples in the Silver Springs springshed (Katz et al. 2001, Phelps 2004) are nearly below detection limit (~1 mg/L), though there is evidence of episodic breakthr ough of dissolved organic carbon in Wakulla Springs during periods of storm flow (Chelette et al. 2002, Katz et al. 2004). BOD levels are also extremely low; Phelps et al. (2006) are among the only studies to report this quantity, and observed no BOD levels higher th an 1.2 mg/L, with most less th an 0.5 mg/L. As such, any BOD introduced into the aquifer from wastewater sources is metabolized or geochemically bound during passage, so no trend in enrichment has b een observed. If BOD loads to the aquifer are significant, it will be manifest as DO changes.

The result of organic mineralizat ion (biotic or abiotic) is increases in so lution of carbonates. Consequently, inorganic carbon levels are uniformly high; dissolved CO 2 levels in 5 springs were between 9 and 12 mg/L (Phelps et al. 2006), and alkalinity levels are usually extremely high (50-150 mg/L). There is evidence that ha rdness (a proxy for dissol ved inorganic carbon in a carbonate system) has increased significantly (p < 0.05) since the 197 0’s (Strong 2004), with nominal increases of 25% (Figure 29). One sp eculative explanation for this is that as groundwater levels have declined with time due to drought and pumping, water emerging at the spring vent is older; longer subsurface residen ce times would affect the degree of calcite or dolomite saturation, and might also affect othe r water quality parameters (e.g., salts, dissolved oxygen and nitrate concentrations) th at are ecologically relevant. D eclines in rainfall pH with time might also be factors in these observed changes.89 Figure 29 – Relationship between early time period (1907-1979) total hardness and recent time period (1990-2003) total hardness, with a 1:1 line (- -), for 49 Florida springs (Strong 2004).

Other Water Quality Parameters

The role of nutrients in spring ecosystems is the subject of this work, but nutrient effects can be confounded by other water quality parameters. For example, many of the same effects of eutrophication on fauna (i.e., oxyge n stress) can be observed in areas where dissolved oxygen is naturally low. There is strong evidence of spatial variability in dissolved oxygen (DO) levels in springs (Phelps 2004 – Figure 30), and also str ong evidence of temporal declines (Scott et al. 2004 – Figure 31); effects of this variability are unknown. Of note is the observation (Andrews 2000) that DO concentrations in most shallow wells , even in the vicinity of dairy farms, were high, which suggests both short residence times a nd ultimately favors the pr ocess of nitrification of organic and ammonia N. Barrios and DeFosse t (2006) also report high levels of DO (> 3.5 mg/L) in the Upper Floridan, and a significant pos itive association (r = 0.5) between nitrates and DO, supporting the notion that shallow, short -residence time flow paths are the dominant conveyance of both. Stevenson et al. (2004) observed that DO in th e vent is the only variable significantly associated with al gal cover (the preferred metric of algal dominance – Quinlan et al.
in review ); increased algal cover was correlated with lower DO. Munch et al. (2006) report declines in DO in the Silver Springs system, and conclude that they are ecologically significant.

Identifying whether spatial and te mporal variability of DO is implicated in ecological change, and, if so, whether it responds to anthropogenic st ressors and management is a topic of future inquiry of immediate relevance to springs. Moreove r, if declines in DO are in some way related to increased algal dominance, low DO levels within algal mats due to microbial respiration may constitute a positive feedback that constrains a sy stem within an algal steady state even after the stressor that led to algal dominance in the fi rst place is removed (Dodds 2003). It is worth pointing out also that high rate s of P mineralization are obser ved in anoxic environments90 (Chambers and Odum 1990); consequently, declin ing DO in spring vents could make more P available for fueling plant or algal productivity.

Figure 30 – Map showing the significant variab ility in Upper Floridan Aquifer DO concentrations, with trend towards anoxic cond itions closer to the springs. Underlying the points are the Silver Springs springshe d and the geologic regions (Phelps 2004).91 Figure 31 – Historic (1972) vs. recent (2002) dissolved oxygen levels measured in the vents of 42 springs showing strongly signif icant declines (p < 0.001) (synth esized from Scott et al. 2004).

Other water quality metrics have also been disc ussed as possibly playing a role in ecological changes in springs. Among the numerous confounders of the nitrate-algae re lationship in springs is the presence of abundant algae at Silver Gl en and Alexander springs, both of which have nitrate concentrations at background levels (~30 ppb). However, both have comparatively high specific conductance (> 1000 μ S/cm); Florida Geological Survey scientists (H. Means, personal communication) have reasoned that salinity changes due to flow declines may have precipitated algal blooms. Cowell and Botts (1994) observed th at conductivity and alkalinity were the best predictors of algal abundance, while nitrate prov ided no predictive power. Formal survey and experimental consideration of this possibili ty should be among resear ch activities improving understanding of spring ecosystems.

Temperature of water is also often reported when examining trends in water quality in springs. Despite the relatively minor variation of water te mperature with time, temperaturehas been used as a metric of subsurface residence times and even more broadly as an indicator of source (Champion 2001, Martin and Dean 2001) . Temperature changes may be particularly relevant as a tracer of local source runoff.

Finally, there is evidence of contamination from the class of chemicals widely referred to as emerging contaminants (industrial chemicals, pe rsonal care products, pharmaceuticals, etc.). Phelps et al. (2006) report results from wells and springs in the St. Johns River Water Management District in which they examined concentrations of non-tr aditional contaminants (e.g., benzene, bisphenol A, caffeine, DEET, napht halene, atrazine, many others). The results are important, though not suggestive at this time, of a strong toxi city risk. Phelps (2004) found DEET in all 3 spring vents at Silver Springs, and in 27 of 35 wells; geology appeared to play a role in where it was found, with all sites where DEET was not observed at locations where Ocala limestone is covered by younger sediments. Other compounds that were found frequently92 include bisphenol-A (polycarbonat e resin), phenols (disinfectant), and triclosan (disinfectant common in antimicrobial soaps). None were measured in sufficient concentrations to pose known ecological or health risks. Few observations of caffeine, a relatively short-lived tracer of human contamination, suggests either long residence times or low loads from septic/wastewater.

SOURCES OF NUTRIENTS

Nitrogen

The vast preponderance of attentio n that has been paid to nitr ate loading has resulted in a relatively mature understanding of the sources of N to the landscape, with several notable exceptions. Delineating sources is made markedly less complex than it might be for P because there is no geologic store of N. As such, all of the nitrogen emerging in spring vents was deposited on the land surface.

The list of presumed loads is l ong, and includes atmospheric de position (wetfall and dryfall), imported fertilizer, imported food (for people and lives tock) which result in domestic wastewater discharges (centralized systems and septic tank s), and biological N fixation by plants, an often overlooked component of regional N budgets. Figure 32 shows a summary of source importance for watersheds in the northeastern US (Boyer et al. 2002, van Breeman et al. 2002), illustrating the kinds of results delivered for springsheds. On the omission of fixation from springshed source budgets, it is worth pointing out that under pre-development conditions biological N fixation represented over 90% of the terrestria l N budget, and is estimat ed to represent 30-35% of the contemporary budget; its omission leads to potentially enormous underestimation of the springshed nitrogen mass loading. The magnitude of the contemporary contribution in the sandy soils of Florida’s springshed is totally unknown.

Figure 32 – Summary of the nominal fractional importance of vari ous N inputs to 16 watersheds in the northeastern US (van Breeman et al. 2002)93 Figure 33 – Concordance between springs hed development and observed NO 3 -N concentrations in Silver Springs (Munch et al. 2006). The st rongest signal appears to be from urbanization since agricultural lands d ecline in importance over the period of record.94 What is clear from the data is that land use is principally responsible for increases in nitrate loads. Munch et al. (2006) report changes in land use in the Silver Springs watershed that are coincident with observed increases in nitrate concentrations emerging from the spring (Figure 33), and Stevenson et al. (2004) found a strong correlation betw een the quality of the local buffer and nitrate concentrations.

Natural Sources

Atmospheric deposition from wet and dry fall is estimated from long-term monitoring locations (National Atmospheric Deposition Program), of whic h there are several in Florida. The longest term site in North Florida is Bradford Forest, where wet fall measurements have been collected since 1977; other sites are used ba sed on proximity to the springshe d in question. Interestingly, trends vary quite substa ntially among sites; Bradford Forest ha s been relatively steady or slightly upwards for both ammonia and nitrate over the la st 20 years, while other stations (MACTEC 2007) have shown marked declines in nitrate, presumably in response to improved regional air quality. Spatial variability in atmospheric load is high, with urban areas exhibiting particularly high deposition rates. Wet fall measurement al one is used to repres ent deposition because dryfall measurement (i.e., dust) is often c onfounded by sample contamination (bird droppings, dead insects); standard practice is to take wet fall measurements (Figure 34) and multiply them by 2 to estimate total atmospheric load. Most st udies do not make this correction, and therefore likely underestimate atmospheric load.

Another source of N in springsheds is biological N fixation (BNF), a quantity rarely estimated in the efforts to understand catchment scale N budgets. Estimates of BNF for temperate environments are 100-200 kg N/ha/yr (Vitousek et al. 2002, Hubbell and Kidder 2003), and are regulated in part by ava ilability of N in soils, and the pr evalence of epiphytic cyanobacterial colonization of soil surfaces, tr ee trunks and leaves (Vitousek et al. 2002). Overall, BNF appears to be ubiquitous, and, though sensitive to pH, drought, temperature and inter-species competition, of enormous importance with respect to the overall magnitude of N fluxes. In an assessment of 16 northeastern watersheds, van Breeman et al. (2002) compute that fixation in forest and agricultural lands represents 28% of the total N budget; for so me basins, agricultural fixation (~2300-4500 kg/km 2 /yr) is larger than fert ilization. These values are 5-fold the expected total atmospheric deposition (~500 kg/km 2 /yr), which in some areas is a significant albeit diffuse source (e.g., Phelps 2004, Katz et al. 1999). While few studies quantify BNF for Florida landscapes, Hiers et al. (2003) report that native legumes in the understory of longleaf pine savannas exhibit high though variable levels of BNF, as expected for fire adapted ecosystems.95 Figure 34 – Time series of annual wet deposition in kg ha -1 yr -1 for ammonium (NH 4 ) and nitrate (NO 3 ) from Bradford Forest, North Central Florida between 1978 and 2005 (from National Atmospheric Deposition Program location at Bradford Forest http://nadp.sws.uiuc.edu/sites ). For load estimation both ammonium and nitrate s hould be included because of nitrification.

Fixation of nitrogen is an energe tically expensive process that re quires either strong electrical gradients (as made possible in lightning and in the Haber-Bosch process via which humans fix nearly all the mineral N used in fertilizer) or substantial biological energy. While it is frequently noted that only certain plants ar e capable of fixing N from the atmo sphere, in fact it is certain prokaryotes, such as microb ial symbionts (e.g., genus Rhizobium ), free-living bacteria (e.g.,
genera Clostridium and Azotobacter ), and blue-green algae (e.g., genera Anabaena and Microcystis ) that make this process possible. The r eason that so few organisms can perform this process is that conditions under which it ca n occur are highly redu cing, a setting created symbiotically in root nodules of N-fixing plants a nd in the cells of blue-g reen algae. Diazotrophs (microbes that fix N) use the enzyme nitrogena se, which is strongly sensitive to oxygen, and 8 moles of ATP (cellular energy ca rrier) per mole of ammonia fixe d to overcome the strong triplebonds that make atmospheric N biologically unavai lable. The ammonia re sulting from fixation is rapidly incorporated into certain amino acids , such as glutamine or alanine, which can be transferred to other N-containing compounds by a variety of commonly occurring amino-acid transfer reactions.

Inorganic Sources

Fertilizers are the most important anthropogenic s ource of nitrates on a mass loading basis (i.e., in contrast to the amount that actually reaches the UFA and the spring, which is discussed below) in nearly all N budget cases for springsheds. Assembling credible data for fertilizer inputs is relatively easy, but requires several assumptions. Authors that take the approach of using fertilizer sales for estimating fertiliz er inputs to springsheds (e.g., Katz et al. 1999) are forced to make assumptions about loads when county boundari es (for which sales da ta are compiled) do not correspond with springshed boundaries. Some authors (MACTEC 2007) use recommended96 fertilizer application ra tes (e.g., from UF/IFAS: www.edis.ifas.ufl.edu– Table 2) coupled with a detailed land use map to estimate fertilizer appl ied to the land. One pr oblem with both method is that attenuation is unaccounted for; that is, as N is absorbed by pl ants or denitrif ied within the field boundary, it will limit the loading rate to th e aquifer. The rate of uptake is not well known, and is expected to be quite va riable as a function of soil type, crop phenology, season, etc. The standard practice in springshed nutrient budgets (e.g., Jones et al. 1996, Katz et al. 1999, Chelette et al. 2002, MACTEC 2007) is to assume a la nd use loading rate. The values, summarized in Tables 3 and 4 are derived from a) long term stormwater data or b) model output.

Land use loading is an inexact measure of loadi ng; in Obreza (2002), N loading rates by land use in Florida are synthesized (Tables 3 and 4). Mo reover, the effects of BMPs on attenuation of N pollution is weak; Currens (2002) reports an incr ease in the median N load after implementation in the deeper karst aquifers in Kentucky. The da ta in Table 4 are more specific to loading rates

Table 2 . Recommended fertilizer application rate s for Florida (UF-IFAS EDIS publications)

Crop/Land Cover N Application Rates (kg N / ha / yr) Source Notes Residential Lawns 80 - 240 Trenholm et al. 2002 Landscape Plants 80 - 160 Trenholm et al. 2002 Athletic fields 200 - 280 Miller and Cisar 2005 for bermudagrass fields

Dairy cow pastures 240 - 360 Sollenberger 2006 est. from recommended monthly application rates Hay production 140 - 300 Staples et al. 2003 Corn Silage 50 - 300 Staples et al. 2003 Sorghum silage 60 - 300 Staples et al. 2003 Vegetable production 180 - 200 Cockx & Simonne 2003, Corn 150/210 Mylavarapu et al. 2002 Irrigated/non-irrigated Sugarcane 90 Mylavarapu et al. 2002 Wheat 80 Mylavarapu et al. 2002 Legumes (soy, peanut, alfalfa) 0 Mylavarapu et al. 2002

Deciduous fruit trees 140-200/200-400 Crocker and Rose 1999 mature fruit trees/mature nut trees, respectively Citrus 140 - 200 Morgan and Hanlon 2006 Plantation Forests (young) 45 - 55 Jokela and Long 1999 in the first year Plantation Forests (established) 160 - 220 Jokela and Long 1999 Applied every 6-8 years

than are the Event Mean Concentration estimates more widely used (Table 3), and they address one of the principal shortcomings of estimating a pplied load only (Table 2), which is that some of that applied load is harvested and only a frac tion of the load reaches waterbodies (surface or subsurface). Several observations can be made re garding this synthesis. First, the ranges of applied fertilizer are enormous (e.g., see range for urban land uses) which introduces some uncertainty when a single number is used; the rea lity is that these rates are unverifiable on a case by case basis, so it is hard to improve accura cy. Second, N export ranges are, in many cases, unknown. Finally, and most importantly, estimated vs. observed rates of actual surface water body loading (right-most column) are in dramatic disagreement (e.g., see row crops and pine tree nurseries) and the difference between imports and exports+loss is s uggestive of massive denitrification (e.g., recreational areas, corn, improved pasture). Further disagreement can be97 observed between estimates of row-crop loads a nd observed concentrations in Table 4 and in Albert (2002), where UFA nitrate levels were up to 130 mg/L.

Both methods (fertilizer app lication rates an d generic land use loading parameters) are questionable for high-recharge low-reactivit y karst landscapes and may dramatically underestimate the loads. By way of example, Ta ble 5 summarizes the concentrations in wells in the Silver Springs springshed (Phelps 2004), whic h indicate strong impact s of overlying land use on nitrate concentrations in the UFA, but at levels far higher than expected by stormwater data in

Table 3. Summary of annual land use loads and even t-mean concentrations nationally and in Florida. Data are for total N load, total N c oncentrations and nitrate concentrations. (after Harmel et al. 2006)
Land use

Location Urban – Low Intensity Urban – Med Intensity Urban – High Intensity Pasture Row Crops Citrus Prod. Forests Nationally 9 11 7.8 8.7 16.1 - 2.9

Reference Omeric Omeric Om eric Reckhow Reckhow - Reckhow

Florida 11.6 16.6 32.1 11.2 6.9 7.2 3.75Total N Loading(kg/ha)

Reference Harper Harper Harper Harper Harper Harper Korhnak

National 2.64 1.85 1.75 4.13 7.06 - 2.73
Reference multiple ‡ multiple ‡ multiple ‡ multiple ‡ multiple ‡ multiple ‡ multiple ‡

Florida 2.29 1.8 2.83 1.97 2.28 1.71 1.31Total N Conc.(mg/L)

Reference Harper Harper Harper Harper/ Graves Harper/ Graves Harper/ Graves Korhnak

National 0.72 0.56 0.57 0.81 4.71 - 0.32 Reference multiple ‡ multiple ‡ multiple ‡ multiple ‡ multiple ‡ multiple ‡ Binkley
Florida 0.28 0.68 0.19 0.80 † 0.80 † 0.80 † 0.08Nitrate Conc.(mg/L)

Reference McConnell McConnell McConne ll McConnell McConnell McConnell Korhnak

† - mixed land use. ‡ - References include Reckhow et al. (1980), Tarabe et al. (1997), McConnell et al. (1999), L.V. Kohrnak (unpublished data), Harper (no date), Graves et al. (2004), Binkley et al. (2004)98 Table 4. Summary land use loads and harvest exports for Florida (after Obreza 2002)

Land use N import (lbs/acre/yr unless specified) N export (lbs/acre/yr unless specified) Loss in Runoff/Leaching Measured|Modeled (lbs/acre/yr) Low Density Residential 0-264 0 6.4|19 High Density Residential 0-264 0 10.4|66 Commercial 0-264 0 28.7|- Recreation Areas 87-220 0 2.4|11 Golf Courses 87-348 0 18.8|- Improved Pasture 120-160 10-41 9.9|- Bahiagrass Pasture 50-180 8-27 9.9|11 Unimproved Pasture 0 0 2.4|- Native Range 0 0 2.4|- Corn 210 0.8 per bushel 6.2|38 Sorghum 150 1.7 per cwt 6.2|38 Cotton 60 0.03 per lb 6.2|38 Wheat 80 1.2 per bushel 6.2|38 Tobacco 80 1.7 per cwt 6.2|38 Peanuts/Soybeans 0 80 per ton 7.9|38 Tomato/Pepper 200 4 per ton 7.9|38 Beans 60-100 0.1 per bushel 7.9|38 Cucumbers/Melons 150 0.026 per cwt 7.9|38 Citrus 50-240 0.13 per box 6.4|36 Peaches 80-100 2.4 per ton 6.4|36 Cattle Feedlot 125 per 1000 lb cow unknown -|62 Dairy 250 per 1400 lb cow unknown -|62 Laying Hens 1.0 per 4 lb animal unknown -|43 Broiler Chickens 0.9 per 2 lb animal unknown -|43 St. Augustine grass sod 260 unknown -|16 Bahiagrass sod 180 unknown -|16 Leatherleaf ferns 100-350 52-164 -|188 Vineyards 100 2.2 per ton -|157 Pine tree nursery 200 125 7.9|157 Pine plantation† 200-250 19 4.9|- † - This level of fertilization is typically over a 20 year rotation, not each year.

Florida (Table 4). Andrews (1994, 2000) report s concentrations downgradient from row-crops and dairies in the Suwannee Valley that regularly exceed the 10 mg/L standard applied by the USEPA for protection of drinking water; notably, despite some wells with nearly 130 mg/L NO 3 - N, no evidence of phosphorus enrichment wa s observed. Such high levels of NO 3 contamination are emblematic of the intrinsic vulnerability of karst areas to over-fertilization. The presence of abundant dissolved oxygen and essentially no or ganic carbon suggest that opportunities for denitrification along the flow paths to the regi onal springs (Blue, Telford, Convict springs, all with dramatically elevated concentrations) are minimal.

Clearly, as nutrient budgeting matures as a scien ce, actual observations of vadose zone fluxes within different land uses are required to constr ain estimates of load. Studies of residential fertilization (e.g., Cisar et al. 2004) have shown strong covari ance between applied loads and99 resulting leachate concentrations (Morton et al. 1988), but with peak observed nitrate concentrations of 2.6 mg/L at nearly 250 kg/ha /yr application rates. Confounding factors in estimating loads are irrigation rate , which varies widely for urban lawns, and the quantity of soil organic matter (Park 2006). Park (2006) found only weak diffe rences in loading between standard mineral fertilizers and slow release varieties, contrary to findings in Cisar et al. (2001). Similar work in row crop fields (e.g., Hubbard and Sheridan (1989) for the SE Coastal Plain) supports the elevated concentra tions predicted by the modeling output in Table 4. Tree crops (peaches, citrus, pecans) have been studied for gr oundwater nutrient enrichment. Concentrations range from 5 to 16 mg/L downgradient of citrus farms (McNeal et al. 1995, Lamb et al. 1999, Crandall 2000), with relatively rapid dec lines in concentration with depth.

Table 5. Nitrate-N concentration (m g/L) in ground-water samples grouped by land use at well locations in the Silver Spri ngs springshed (From: Phelps 2004).
Land use N Minimum Maximum Median Urban 29 <0.02 5.9 1.15 Agriculture 13 0.05 12 1.7 Rangeland & forest 11 <0.02 2.2 0.09

Transportation and utilities 3 0.88 4.0 1.57

Organic Sources

Nitrate loads from organic sources (i.e., human and animal wastes wherein the organic and ammonia N has been nitrified) ar e also extremely important, and methods for estimation of loads are relatively well-developed. Most often, the assumption is that all of the load from septic tanks and wastewater sprayfields is r echarged directly to the Florid an Aquifer; MACTEC (2007) and Chelette et al. (2002) make the assumption that 30% of the load is attenuated; verification of this level of attenuation, and comparable attenuation rates for fertilizer loads, is a research need.

Embedded in Table 4 is an estimate of the load ing derived from confined animal operations, a source of significant nitrogen, BOD and other wa ter quality concerns. Andrews (1994) reports on concentrations in th e groundwater downgradient of dair y operations, and observes nitrate concentrations averaging 26 mg/L and ex ceeding 100 mg/L in some cases. Woodard et al. (2002) also observe massive enri chment downgradient of a dairy farm, with concentrations as high as 68 mg/L and averaging 18 mg/L in soil solutions; the relationship between soil solution chemistry and groundwater c oncentration is unknown, but inkarst areas with limited denitrification and rapid porewater turnover, a strong correspondence is likely. Poultry operations (Hatzell 1995) are similarly significant sour ces of N (mean groundwater concentrations of 13 mg/L), and have been shown to be sources of aerosolized ammonia as well (Lopez-Zamora et al. 2007).

The authors of Table 4 do not report whether the confined animal operations data in Table 4 are nitrate loads or total N, but assuming the latter, mo st of that N will be as ammonium or organic N. That form of N is far less mobile in the e nvironment than nitrate and thus is not observed in spring vents at concentrations of concern. Ho wever, the microbially mediated nitrification process converts ammonium to nitrate, using oxyge n as the terminal electron acceptor. As such,100 high loads of dairy waste can be converted to nitr ate within the oxic conditions persistent within the upper Floridan Aquifer. Associated effects on dissolved oxygen will be discussed below.

Also embedded in Table 4 is an estimate of N load s from septic tanks. The typical number for a standard On Site Treatment and Disposal System (OSTDS) is that it yiel ds 3-4 kg N per person per year; assuming that 70% (Anderson and Otis 2000) of that N reach es the groundwater, a typical household is expect ed to discharge 5-7 kg NO 3 -N/yr. Modern performance-based treatment systems can reduce N loads by 75%, which, despite the costs, is widely considered essential because of the intrinsic vulnerability of the Floridan Aquifer to nitrate enrichment.

Septic tank loads to aquifers ar e problems ubiquitous to areas with strong surface-groundwater connectivity. Reay (2004) presents data from the coastal zone in Virginia that may provide some parallel to the Florida experien ce (Figure 35). In particular, th at work shows that drainfield concentrations of both N and P are extremely high (3-4 orders of magnit ude higher than ambient

Figure 35 – Relation between nutrient load to groundwat er from septic tanks (after Reay 2004).

environment conditions) and that P and N attenua tion rates are markedly different. Moreover, the N:P ratio of the leachate and environment are roughly the same (~ 10:1). During passage from the drainfield to the site s where groundwater discharges to the surface there is typically a 2- 3 log (2 to 3 orders of magnit ude) decline in dissolved inorgani c P (DIP), and typically much101 more; the majority of samples along the flow pa th are within a factor of 10 of ambient concentrations (i.e., 3-log reduction). In contra st, flow path attenuation of N is smaller, as indicated by the 1-2 log reduction in concentrations of dissolved inorganic N and this results in an N:P ratio of ~50-5000:1.

Not included in Table 4 is an estimate of wast ewater discharges from centralized facilities (points source exports), which have been the subject of close scrutiny in areas like Lake City and Tallahassee. Loper et al. (2005) reviewed the evidence fo r nitrate enrichment from the Tallahassee wastewater sprayfield at Wakulla Springs, and concluded that it was the most significant (as high as 70% of the total load – Chelette et al. 2002) and most manageable of the various springshed sources. That sprayfield discharges approximately 20 million gallons per day to the land surface in the region of the springshed known as the Woodville Karst Plain where the aquifer is unconfined. Rapid and unattenuated movement of water and nitrates into the groundwater is assumed in the sandy unconf ined soils of the karst plain.

Isotope Tracing of N Sources

The identification of multiple sources of N leads to the management-relevant question of which sources are most important for meeting load reduction objectives. Whole springshed budgeting is relevant in this regard, since it allows comp arison of bulk loads, but th e substantial uncertainty about the fraction of load that actually reaches th e aquifer has led researchers to use the ratio of stable nitrogen isotopes ( 14 N and 15 N) in a sample to deduce the source (Hornsby 1994, Panno et al. 2001). This method has a long history of be ing used in surface flow tracing, and has been carefully explicated, including se veral important caveats, in Kenda ll (1998). The central premise is that animals preferentially excrete the lighter isotope in urine, resulting in enrichment in organic wastes; nitrification fo llowed by denitrification during st orage (e.g., pipes, septic tanks and waste lagoons) and transport (through soil and aquifer) furthe r fractionates organic N with resulting waste stream nitrate-enriched with 15 N (10-20‰). Since fertili zer has an isotopic ratio close to atmospheric (defined as 0‰), usually 3- 5‰, the isotopic ratio of nitrate-enriched source water will signal the principal source: low (3-8‰) for fertilizers and high (10-15‰) for organic sources (Figure 36 – Phelps 2004).

Bacchus and Barile (2005) used isotopes to conclude that loadi ng of N to Alexander Springs was from on-site sewage disposal; while high 15 N enrichment in the vegetation (~8-12‰; note that vegetation incorporation is typically a lightening process) is suggestive of unusual isotope loading, differences in flow betw een sewage (est. < 2 cfs) and the spring vent (> 100 cfs) suggest the need for further study before concluding that septic tank loads are important in Alexander Springs. This inference is further complicated by not having a reference value for the isotopic composition of the vent water. What is clear is that isotopic ratio al one (i.e., without an associated concentration signal) is not an indicator of pollution, since nitrate concentrations downstream of the main vent (after mixing with septic tank effluent in the groundwater prior to emergence at the spring) are still very low (~ 40 ppb, Mattson et al. 2006).102 Figure 36 – Summary of N isotopic measurements for groundwater samples in Silver Springs springshed (Phelps 2004), and their interpretation. Spring samples were generally higher (4- 9‰) and strongly variable in time.

This technique has been widely used for trac ing N sources in springs (e.g., Toth 1999, Champion and Starks 2001), and the results suggest that, in th e vast majority of cases, fertilizer sources are of principal importance (Figure 37 and Table 6). In some cases, however, wells and springs have been measured with isotopic ratios strongly indicative of organic so urces: Lafayette County wells/Poe Springs – Katz et al. (1999), Alexander Springs – Bacchus et al. (2003), DeLeon Springs – Phelps et al. (2006).

Table 6. Comparison of δ 15 N values measured in sp rings and groundwater basin

Location N δ 15 N values Nitrate sources References

Min Max Median
Rainbow springshed 19 0.5 7.7 2.4 Inorganic fertilizers Jones et al. (1996)
Wekiwa springshed 9 3.6 11.2 4.8 Inorganic fertilizers, organic sources Toth and Fortich (2002)
Woodville Karst Plain 13 1.7 13.8 6.8 Inorganic fertilizers, organic sources Katz et al. (2004)

Wakulla Springs (subset of the Woodville Plain) 3 5.3 6.1 5.8 Inorganic fertilizers Katz et al. (2004)

Suwannee Basin springs 26 2.7 10.6 5.4 Inorganic, organic

sources Katz et al. (1999)103 Figure 37 - Distribution of δ 15 N values from SWFWMD spring (Champion and Starks 2001).

It is also important to note that one of the arguments posited against the importance of animal wastes as a source is the isotopic composition of the nitrate that is discharged from the springs, which is comparatively light an d therefore indicative of inorga nic sources. This neglects fractionation of N isotopes during nitrification of organic and ammonia N to nitrate; during this process, the heavy isotope is preferentially exclude d, so the resulting nitrate is lighter. This will tend to mask the influence of organic sources unless the nitrifying bacteria oxidize all of the available Kjeldahl N.

Another confounder of isotopic infe rence relates to the omission of biological nitrogen fixation (BNF) among the sources. Gene rally speaking, the fixation of N 2 gas (defined as 0‰) from the atmosphere induces some kinetic fractionati on, preferentially uptaking the lighter ( 14 N) isotope. As such, the resulting fixed nitrogen will be -3‰. If this source is regionally significant, and there are few reasons to presume that it is not, then a mixture of nitrate sources from BNF and septic tanks may resemble nitrate of a fertilizer source. Because flowpaths delivering water to the springs are underground, and because they re present mixture sites for multiple nutrient sources, our mechanistic understand ing of flowpath changes in 15 N and subsequent mixing needs to be improved before isotopic information is used with authority. Moreover, there is a significant need to understand the te mporal variability of isotopic signals; Phelps (2004) reports values from Silver Springs that vary dramati cally (~3‰ vs. 8‰) between observations 6 months apart. Quantifying time-series variability and mechanisms for it are important for improving our allocation of load responsibility, and are likely to yield insights into the mechanisms for delivery (e.g., as a function of flow).

There are several unexplained discre pancies in the tracing of sources that are likely fruitful areas for further inquiry. Principal am ong these is the observation based on loading estim ates that the104 Tallahassee sprayfield is respons ible for over 70% of the total anthropogenic N enrichment in Wakulla Springs, while the nitrate isotopic measurem ents are consistent with fertilizer sources (6.3‰ – Chelette 2002). For reasons related to th e complexity of nitroge n fractionation, even in the subsurface where denitrification is presumed to be negligible, it is relevant to ensure that isotopic measurements are being interpreted corr ectly. Given the strong fractionation associated with nitrification and the potential influence of biological N fixation on isotopic measurements of nitrates in the water mixtures at spring vents, there is some poten tial that our use of this binary inference about sources is incorrect. Careful trac ing of isotopic fractionation from source to vent is necessary to establish the appropr iate manner in which to interpret δ 15 N data.

Synthesis of N Springshed Budgets

Most attention in springs protection has been on N, and most of the focus has been on identifying the sources of N to the spring ecosystems. This section selects a couple of key case studies.

Figure 38 shows the loads of N to groundwater after attenuation by soil in the Wekiva springshed, the most comprehensive and recen t example of a detailed springshed N budget (MACTEC 2007), wherein they concl ude that a) only a small fracti on (~21%) of the N applied to the land surface actually reaches the groundwater or surfacewater, and b) that despite a strong urban springshed landscape with septic systems, fertilizer use dominate the load. Though septic tanks represent nearly 25% of the total load; the isotopic ratio of the spring is relatively low (~4.8‰ – Table 6). A local aquifer vulnerability assessment has been done for this area as well, allowing more formal evaluation of links between particular locations and dynamics of water quantity and quality (Cichon et al. 2005).

Figure 38 – Summary of N loads to groundwater for the Wekiva springshed (MACTEC 2007).

A similar study of the Rainbow Springs catchme nt yielded a similar conclusion about the principal source of N. Fertilizer, principally fr om improved pastures, was estimated to contribute almost 50% of the total load (Table 7 and Figur e 39); horse farms and confined cattle operations105 were responsible for an additional 30% of the lo ad. Overall, atmospheric deposition was also important as a source (over 15% of the total load) while septic ta nks and sewage discharges were considered negligible. This is somewhat cons istent with the very low isotopic ratio (2.4‰).

Table 7 . Total N loadings (tons/yr) into groundwater in Rainbow springshed (Jones et al. 1996) Source Regions % of total loads Eastern Central Western Atmospheric 431 504 467 16.8 Septic tanks 22 17 24 0.7 Turf Fertilization 50 11 18 0.9 Golf courses 84 0 35 1.4 Sewage 13 3 1 0.2 Septage spreading 58 0 24 1.0 Row crops 0 0 44 0.5 Cattle production 410 439 407 14.7 Horse farms 991 510 < 1 17.5 Improve pasture 1,728 1,364 871 46.3

45%

17% 15% 18% Fertilized pasture Atmospheric deposition
Cattle
Horse farms Septic tanks
Turf fertilization
Golf courses Sewage Septage land disposal Row crops

Figure 39 – Relative contribution of N l oads different sources to the groundwater discharging from the Rainbow Springs (after Jones et al. 1996).106 Figure 40 – Estimates of a) surface loads to Wakulla Springs, b) annual delivery based on coarse hydrologic modeling, and c) actual loads by region. The 350 cfs of water at the vent (ca. 0.9 mg N/L) represents a flux of 273,000 kg, an a ttenuation of nearly 90% (Chelette et al. 2002).

Another effort in the Wakulla springshed (Chelette et al. 2002) where the Tallahassee sprayfield facility is in the unconfined re gion and demonstrably connected to the spring, suggests that the c) b) a)107 spray field is uncharacteristically important (40%) as a source of N (Figure 40b). This estimated importance arises only after the transport and delivery is considered; when only direct loading is considered, atmospheric depositio n dominates the load, and human wastes are but 25% of the load. It is also notable from this study that the consequences of loading on the confined region are a function of geologic regime, with nearly a ll the effluent loads occurring in the unconfined region, despite most of the fertil izer application occuring in th e less vulnerable confined and semi-confined regions; part of this may be e xplained by sinkhole-conduit systems that link large regional loads (e.g., Tallahassee’s sprayfield) to th e spring. The observation that the nearly 1.15 million kg N/yr surface loading yields an observed load of only 73,000 kg/yr (Chelette et al. 2002) suggests the large attenuation potential of those parts of th e state where su rface water and high-carbon soils and sediments co-exist.

Foster (2008) concludes that atmo spheric deposition is a large porti on of the load in Ichetucknee (Figure 41); the fate of atmospheric N is unknow n, and omission of biological N fixation further adds to the uncertainty. Moreover, because of how land use loading is computed, atmospheric deposition is very likely double counted.

Figure 41 – Synthesis of sources for the Ichetucknee Springshed (Foster 2008).

Phosphorus

Comparatively little is known a bout the magnitude of P sources in springsheds. While there have been numerous studies of P loading from biosolid app lication (Graetz and Nair 1995), septic tanks (Reay 2004), wa stewater sprayfields (Loper et al. 2005), and ferti lizers, there is essentially no consolidated information on the re lative magnitude of these sources in the fluxes observed at springs. Part of that is almost certa inly because no significant changes in P loading have been observed (e.g., Figure 28, which show s springs with marked N enrichment exhibiting108 P concentrations unchanged from inferred background levels). Moreover, mo st of the work on P loading in Florida has focused on areas where sandy soils and perched aquifers dominate (e.g., spodosols – Graetz and Nair 1995), and not where the carbonate sink is large and proximate to the load.

Loper et al. (2005) argue that this atte ntion on N is warranted by two observations: 1) P loading is not increasing in any systematic way, so us ing it to explain recent ecological changes in springs is problematic, and 2) geology is the prin cipal regulator of P loading, so loading is not management sensitive. Stevenson et al. (2007) echo this observation, suggesting that while modest correlations between al gal abundance and P were observed, the ability to regulate P in springs was limited, warranting attention to N alone as a management objective.

In contrast, however, Sigua and Tweedale (2003) report large gr oundwater P loads to the Indian River Lagoon. Indeed, the N:P ratio in measured groundwater discharges was just over 3.4:1 (84,900 kg N vs. 24,300 kg P), suggesting high P m obility; this area is sandy and wet, so the generality of these results to springs more broa dly may be confounded by characteristics of the flowpath. This finding is particularly su rprising because surface wa ter inputs are strongly dominated by N (i.e., N:P ~ 30:1 for atmospheric sources, and 8:1 for both point and non-point surface water discharges). Concordance betw een submarine groundwater discharge flowpaths and the karstic flowpaths to freshwater springs is unknown.

On the latter point regarding the geological contro ls on P there are two factors to consider, one pertaining to load generation and the other to load attenuation. Geologic loads of P are expected to be high (though variable) in this landscape be cause of the presence of the Miocene Hawthorn formation, previously described. Among the ma ny important features of this interbedded clay/sand layer is the abundan ce of apatite minerals (CaPO 4 ) that are the basis of Florida’s phosphate mining industry. Apatite is a highly weatherable material, and water percolating through, or more importantly running over the Ha wthorn is expected to dissolve significant quantities of P. If this weathering/erosion pro cess occurs proximate to the spring, it is quite likely that concentrations of P at the spring vent will be elevated. Indeed, Ichetucknee State Park contains a site historically us ed for mining phosphate, illustrating the potential significance of proximate P sources. No systematic effort to un derstand the links betwee n geologic setting and P concentrations has been undertaken, but it is cl ear that substantial a nd unexplained variability in P concentrations exists. Among the hypotheses that might be tested are th at P levels are high in springs with flowpaths that are sufficiently sh allow to interact with the Hawthorn, or in close enough proximity to the Cody Escarpment wher e surface water weathering processes and subsequent sinkhole capture lead to P enrichment. Given weak but significant evidence of links between P concentrations and the dominance of Vaucheria spp. (one of the pr incipal filamentous algal species of concern) (Stevenson et al. 2007), additional a ttention to P as a driver of springs ecology is warranted. As such, efforts to unde rstand the degree to which P in springs is a management sensitive parameter are also relevant.

Carbon (BOD)

The role of organic carbon in spring systems is ne gligible in all measurements except those made during extreme events when either a) stormflows dominate via conduits and/or surface runoff,109 bearing DOC, or b) tannin-stained water from downstream rivers cr eates a backwater event, also during extreme floods. The delivery of dissolv ed inorganic carbon, a necessary condition for driving high levels of submerged aquatic produc tion, is clearly sufficient. Calcite saturation indices and alkalinity, both of which measure the degree of cati on (Ca, Mg) satu ration, but also reflect the carbonate concentratio ns in the water, support the inference that rock-water interactions in the aquifer maintain high levels of DIC. Katz et al . (2004) reports bicarbonate concentrations averaging nearly 200 mg /L for large springs, and others (Katz et al. 2001) report similar levels for smaller springs.

Despite the relative paucity of organic C in the spring vents, there are well-known sources of organic C in the springshed. Natural sources ( DOC from wetlands and shallow-water table soils in the confined reaches) and anthropogenic sources (e.g., septic tanks, animal waste lagoons, waste water sprayfields) exist; given the potential BOD and NOD load from anthropogenic sources, it is notable that there is an inverse correlation between dissolved oxygen and δ 15 N (a tracer of organic N) among the springs at Ichetucknee (Katz et al. 2004). For example, given a cBOD+NBOD load of 200 mg/L in septic tank leach ate, and a flow rate of 400 L per day, one septic tank can deoxygenate over 6 million liters of aquifer water per year (assuming nominal DO of 5 mg/L) if the drainfield is in the groundwater. Mode rn systems (drainfields and sprayfields) allow greater oxygen ation of the water, so calculations above are likely to be reasonable primarily for older or illegal septic system. The Ichetucknee springshed has over 2,000 septic tanks (Foster 2008), while nearly 30,000 exist in the springshed of the Wakulla (Chelette et al. 2002), and 65,000 in the springshed a nd surface water basin of the Wekiva (MACTEC 2007). By extrapolatio n, the potential DO impacts of septic tanks alone are high; understanding the role of DO in springs, and then how DO management might be improved, appear to be avenues for important research, all despite the relative unim portance of BOD at the spring boil.

Other Water Quality Parameters

Mechanisms for the variability in the delivery of salts are unknown. There appears to be both a hydroclimatic and geological basis (Phelps 2004, Scott et al. 2004, Toth and Katz 2006). For example, strong inverse correlation between conductivity and flow was observed at Wakulla Springs (Loper et al. 2005), suggesting spring vent water is a mixture of low conductivity water varying in discharge over short time scales, a nd high conductivity water delivered more evenly. Extrapolating that relationship s uggests that, during periods of drought, conductivity levels rise. It is currently unknown what the implications of age-salinity relationships might have on springs, but episodic disturbance has been widely implicated in ecological transitions in lakes (Scheffer et al. 2001). The mechanisms for the role of salinity, fo r example, as a stressor of SAV or excluder of grazer populations that regulate filamentous algal population, is currently unknown and may require substantial research to clarify.

Similar uncertainty exists about mechanisms and importance of delivery of industrial contaminants (e.g., DEET). While it is clear th at the source is human waste waters, the time domain for delivery, the hydrogeochemical c ontrols on transport, and the ecological consequences are not so clear. Given the low concentrations generally observed (e.g., Phelps 2004), this is an area worthy of additional monito ring for temporal trends, but probably not yet110 for detailed accounting of sources. Analysis of ecological impacts, on the other hand, is extremely important because of evidence of ecot oxicological consequences of even relatively low concentrations.

SINKS AND TRANSFORMATIONS OF NUTRIENTS

Typical nutrient sinks in waters hed budgets include assimilation by biota (including harvest and export for crops, wood and animals), dissimilato ry biotic removal (denitrification, sulfate reduction), and both physical and chemical abiotic sinks (e.g., sedimentation and sorption of P). Van Breeman et al. (2002) consider the main pathways fo r loss of N (plant uptake, crop removal, denitrification, riverine export) and conclude that denitrificati on represents 50% of estimated losses, and exports (food, wood and riverine) account for 31%; the remainder is estimated to be changes in storage (soil, biomass).

Nitrogen removal is characterized by particular “hot spots” and “hot moments” (sensu McClain et al. 2002; hot moments might be periods during wh ich rivers flood and thus interact with organic rich riparian sediments) in the landscape. Wetlands in particular are well-known sites for both uptake/soil accretion and de nitrification. Table 8 summarizes the rates of N removal in wetlands and riparian areas in the Southeast. Figure 42 summarizes the same N cycle processes illustrating both the enormous potential for remova l and the significant heterogeneity in process
rates. Table 8. Nitrogen removal mechanisms in floodplain and wetland riparian forests in the southern United States (after Walbridge and Lockaby 1994) Mechanism Location Rate (kg N / ha / yr) Deposition Maryland 11.0
Louisiana 350.1 Little River, Georgia 31.5 Denitrification (potential) Coastal Plain, Georgia 224.0 Tar River, NC 130.0 Maryland 47.7 Denitrification (mass balance observations) Coastal Plain, Georgia 1.5 Tar River, NC 64.2 NH 4 + adsorption Maryland 0.8 Tar River, NC 16.2 Microbial Immobilization Coastal Plain, Georgia 87.0 Tar River, NC 15.5 Little River, Georgia 51.8 Plant Uptake Maryland 15.0

Denitrification, the most important long-term sink for N, is principally constrained by the availability of carbon (denitrifying bacteria ar e heterotrophic) and the absence of oxygen (when O 2 is present it provides a thermodynamically favorable electron acceptor); the process may also be nitrate limited (in low concentr ation settings), but is rarely constrained by the presence of the microorganisms. Indeed, Andrews (2000) obse rved that, while all other conditions for denitrification were unfavorable, counts of denitrifying bacteria were freque ntly greater than 10,000 per 100 mL.111 Figure 42 – Summary of measured N loss rates from selected wetland and ri parian zone studies (from Mitsch et al. 2001). To convert g N/m 2 /yr to kg N/ha/yr, multiply values in this table by 10.

Nutrient Sinks in Karst Landscapes

Nitrogen

As with non-karst regions, the two significant long- term sinks for N are in the stored organic matter and denitrification to the atmosphere. The former is expected to be finite, with mineralization eventually roughly balancing uptak e (agricultural harvest being one exception). The latter is rate limited by the availability of sites in the landscape with the combination of anoxia and labile organic matter to dr ive heterotrophic microbial metabolism.112 Evidence for denitrification once nitrate is in the gr oundwater is absent. Several lines of evidence appear to confirm that nitrates are effec tively conservative once in the UFA. First, as reported in Katz et al. (2004) and Phelps (2004), the partial pressure of N 2 gas in the water that emerges from the spring vents is consistent with expected equilibrium at recharge (Figure 43). Denitrification would be expected to increase N 2 concentrations; this condition was observed at only one of 14 locations in Suwannee and Lafeye tte counties (Katz and Bohlke 2000). Second, the Rayleigh distillation model for isotopic kine tics (Kendall 1998) in a closed system predicts that processes in which fractionati on occurs (i.e., pr eferential use of 14 N vs. 15 N during denitrification) would tend to in crease the “heaviness” of the remaining nitrogen (Einsiedl and Meyer 2006). As such, with denitr ification, we would expect to see δ 15 N of the groundwater steadily increasing with age. Wh ile there is no study of that pa rticular relationship, the large fractionation due to denitrification (~30‰ – Kendall 1998) and generally low values of δ 15 N in spring discharge indicates that this process is sma ll in magnitude if it occurs at all. Finally, the UFA is characterized by oxic conditions (Barri os and DeFosset 2006) and low levels of DOC (Katz et al. 2001, Phelps 2004) which suggests an envi ronment unconducive to denitrification.

Figure 43 – Chart comparing concentration of N 2 and Ar gas dissolved in spring water (Katz et al. 2004). The fact that observations cluster near the air equilibrium li ne and at a position that indicates a 20-25 degree C recharge temperat ure suggests the absence of denitrification.

Given that N is likely to be ne arly conservative in the subsurface, efforts to attenuate loads prior to transport through the soil to the aquifer are critical. Clear ly, land use planning figures prominently in reducing loads (see sections abov e). Best management practices can also, in some cases, provide attenuation, but the degree of ef ficacy is likely to be a function of the extent to which water remains at or near the surface wh ere BMPs can operate. In areas with rapid recharge (i.e., karst plains), oppor tunities to store and treat stor mwater are limited. Some of the conventional treatments for managing diffuse lo ads of nitrogen (denitr ification walls, lagoons,113 treatment wetlands) are limited by hydrogeologic constraints on surface water storage and residence times; as such, managi ng/reducing sources is likely to be far more effective than managing/enhancing sinks. Some nitrate removal in the riparian/hyporhe ic zone after spring emergence is likely (Mattson et al. 2006) and sensitive to land use practices (Hearne and Howard-Williams 1988), but also largely unquantifie d, and generally irrelevant for protecting headwater areas.

Phosphorus

P sinks in carbonate landscapes are overwhelmi ngly dominated by mineral adsorption. While in other parts of Florida, movement of P through th e sandy aquifer is rapid a nd largely unattenuated (Campbell et al. 1995), in the karst regions where carbonate sorption occurs, P breakthrough has not occurred. The magnitude of the carbonate sink is unknown. It is plausible, given strongly preferential flowpaths in the subsurface, that sa turation of P sorption capacity can occur within management-relevant time spans (Wang et al. 1995). The implications for this saturation are potentially profound given a) the relatively str onger association of ecological change with P rather than N, and b) the time domain of wate r quality recovery afte r breakthrough (Mansell et al. 1995, Wang et al. 1995), even with complete reduction in loading. As such, this should be an area of active research, with the primary objectiv e to determine the degree of sorption saturation along principal flowpaths, a nd the timing and implications of future breakthrough.

In a carbonate aquifer system in the Florida Keys, Dillon et al. (2003) used radio-labeled 32 PO 4 to demonstrate rapid adsorption onto the Key Lar go limestone. That research established an equilibrium phosphate concentration in solution after extended interaction with phosphate-rich water of ~ 26 μ M, which is markedly higher than concentrations observed in springs. Desorption isotherms, during which the sample that had b een exposed to high levels was exposed to phosphate-free water, showed convergence to the sa me concentration, sugges ting that this is the equilibrium P concentration of that limestone s ubstrate. While infere nce from this distant experiment is perilous, it does s uggest that the capacity for limestone to sorb P is extremely large. The kinetics of this uptake is not well-know n and is a key factor in understanding the functional capacity of the matrix to sorb P loaded from the landscape.

In a study of P loading in calcare ous soils in South Florida, Jaber et al. (2006) report extremely low P levels (0.04 mg/L) in groundwater beneat h lands to which organic amendments were added. These concentrations support the contention th at P sorption is large; further results in that work underscore the differences observed with sandy soil where P concentrations were nearly two orders of magnitude higher un der the same treatment. Reed et al. (2006) study on P movement in calcareous soils in South Florida suggests remarkably little transport despite high application rates.

Under reducing conditions, the mobility of P can be changed dramatically. Kroeger et al. (2007) observed high levels of phosphate in submarine groundwater discharge in the Tampa Bay area draining regions of high populat ion density and under reducing c onditions; the resu lts of that study , however, are not applicable to most of the Floridan Aqui fer where N is transported as nitrate and oxic conditions prevail.114 Carbon

There is no sink in the subsurface for inorganic carbon except abiotic gas exchange. Since gas exchange is expected to be st rongly limited in the Floridan Aquifer, most aquifer water is saturated to super-saturated with respect to CO 2 , a condition that leads to considerable degassing of CO 2 at spring vents. One of the consequences of this is proximate de clines in pH, and the super-saturation of groundwater wi th respect to calcite in spri ng river water (with important implications for karst channel development).

Aquifer flowpath attenuation of organic carbon is n early complete, based on typical low DOC conditions observed in springs. While episodic pulse s of DOC rich water stains the vent flow at Wakulla Springs, among others, DOC levels in sp ring vents are ~ 2 mg/L or less. Numerous pathways for organic carbon attenu ation exist. Most prominent is bacterial respiration, a process made possible by high levels of dissolved oxygen in much of the Upper Floridan Aquifer and high bacterial counts throughout the aquifer por osity (see Simon 2000 for karst aquifers generally). Other possible sinks include abio tic processing (carbonate adsorption – Zimmerman et al. 2006), though the relative importance of this pa thways is, as yet, poorly quantified; areas where surface water interacts most with groun dwater are dominated by microbial processes, while areas with relatively lower levels of connectivity exhibit DOC breakdown products indicative of abiotic attenuation pa thways. The relative importance of the different pathways is relevant to the karstificati on of the carbonate matrix.

Other Water Quality Parameters

Much of the literature on othe r water quality parameters focu ses on the presence of emerging contaminants (e.g., DEET). Surprisingly little is published about sinks for DO, despite the obvious merits of understanding beha vior in the subsurface. What can be inferred from existing data is that natural a ttenuation is small (high DO is observe d over large areas of the UFA – e.g., Katz and Bohlke 2000), but the effects of anthropogenic oxygen-consuming wastes at the springshed-scale are unknown.

Work on Florida’s karst suggests that some of the emerging contaminants may be attenuated via biotic mineralization and.or biotic sorption as a function of time spent in the subsurface (Phelps 2004) while others (e.g., DEET) appear more cons ervative. Overall, however, there is limited information on the extent of attenuation, nor on th e factors that might le ad to variability in attenuation rates.

Springshed nutrient budget – Synthesis of Sinks

Numerous springshed nitrogen budgets exist and, de spite subtle differences in assumptions and data quality, offer an opportuni ty to synthesize conditions. The best exemplars of nitrogen budgeting work (Boyer 2002, Van Breeman et al. 2002) are for surface drainage basins in the northeast; in those efforts the s ources (previously presented) a nd sinks are quan tified with the metric of success being the degree to which the N budget can be closed. Figure 44 shows estimates of the mean importance of the key si nk/transformation pathways across 16 watersheds, and illustrates a level of detail not matched in Fl orida’s springshed examples. It is important to115 note that Boyer et al. ’s study sites are all surface water sy stems, and springs are subsurface basins; intrinsic differences in processing capacity and delivery rates may limit the general utility of the observed N partitioning. Moreover, in no cas es that were reviewed for this document were exports in food or changes in storage considered explicitly; the principal reason for this is that the objective of the springshed examples is to identify key inputs for management, not close the N budget. However, future efforts to more expl icitly account for N sources and sinks would aid in understanding some of the curious results obtaine d, particularly with rega rd to the degree to which N loads are attenuated in springsheds de spite substantial agreement that N sinks are largely absent.

Most of the sinks are of variable interest. While it is likely important to close the N budget to know the magnitude of food exports and changes in storage, thes e may be less important when allocation of load reduction respons ibility is the pr incipal objective.

One striking thing about the N fate pie-chart (Figure 44) for surface water systems is the magnitude of riverine export. On average, rive rs transport 20% of the N deposited on land to the sea; the range is from ~12% to more than 35%. Given the arguments that most Florida springshed budgets neglect the important cont ribution of biological N fixation, and, more importantly, that a karst landscape is intrinsically vulnerable to N transport, the expectation would be that springsheds deliver more N (i.e., >> 20%) than other watersheds where denitrification and storag e potentials are higher.

Figure 44 – Summary of the nominal fractional importance of vari ous N outputs and sinks for 16 watersheds in the northeastern US (van Breeman et al. 2002).116 A synthesis of 5 springshed loading budgets in Fl orida reveals the opposite result (Figure 45). For 4 of 5 springsheds (Ichetucknee is the excep tion), the spring discharges 10% or less of the total estimated load. In the Wekiva basin the load is 1,800 tons N/yr and spring discharge 230 tons N/yr. This result is after estimating the load attenuation between surface and groundwaters, estimated to be 7,600 tons N/yr (MACTEC 2007). In the Silver Springs system, the landscape attenuation is nearly 95% of the estimated load (Phelps 2004). Jones and Upchurch (1994) quantified the N load to the Kings Bay system, a nd observe large attenua tion (2757 tons/yr load vs. 335 tons/yr spring flux); they suggest, based on a survey of wells, that proximate sources dominate the loading. They also posit that most of the N in th at spring is of or ganic origin (e.g., wastewater sources) despite low δ 15 N levels.

There are several plausible explanations for the obs erved attenuation in springs heds. First is that intrinsic time lags (Katz et al. 2001, Toth and Katz 2006) make comparison of contemporary spring fluxes with contemporary load ings problematic. That is, if water recharge occured ca. 20 years before present, the nitrate concentrations in the spring vent are not representative of observed landscape loads. Water that has residence times longer than 50 years (expected to be as much as 70% of the flow – Toth and Katz 2006) w ould be expected to have concentrations near background, confounding budgeting effo rts. While this is certainly possible, it is likely, as previously discussed, that most of the N is de livered via flowpaths with short residence times.

Figure 45 - Estimated N attenuation fr om budgeting work computed by comparing estimated landscape load to fluxes out of springs. (Chelette et al . 2002, Foster 2008, MACTEC 2007, Phelps 2004, Jones et al. 1996). Note: values reported for Wekiva are after estimated adjustment for load attenuation betw een the surface and subsurface ; levels of to tal estimated attenuation are thus far higher.117 In a similar argument, it is also plausible that the effective springshed is markedly smaller than the delineated boundary. Phelps (2004) provides 10- yr and 500-yr capture zones, both of which are substantially smaller than the delineated sp ringshed; as such, the area over which active N loading occurs is likely small, and estimates of total springshed load are too high. Moreover, much of the delineated springshed for systems like Ichetucknee is within the confined region, and much of the water drains via surface flowpath s to other catchments. Developing nutrient and water budgets for effective springsheds is likely to yield a far clearer picture of source importance and help realistically target load reduction strategies.

It is further plausible that the subsurface conduit structure delivers only a portion of the water to the spring vent; Toth and Fortich (2002) speculate that this is o ccurring in the Wekiva system, with the remaining water (and nitrogen) delivered directly to the St. Johns River via subaqueous or diffuse discharge.

Finally, it is important to consider if the obse rved attenuation is the result of biological processes. While it would be surprising to observe accelerated denitr ification in the karst regions, denitrification in the confined and semi -confined regions is cl early very rapid (Cohen et al. 2007). The general finding that most of the N loading occurs in the unconfined region and that denitrification in the subsurface is negligible sugge st, however, that these levels of attenuation are confounded.

Regardless of the mechanism for the observed a ttenuation, it is clear th at understanding why it occurs is relevant for management. If it is due to time lags, then forecasting changes in N loading due to contemporary management need to incorporate these lags . If it is due to differences between effective and delineated sp ringshed size, then improving our understanding of N processes within the effective catchment are important. If it is due to flow by-pass, then larger scale N budgets (e.g., for the St. Johns) need to account for the displaced load. It is also worth pointing out that regardless of the wate r quality parameter of choice, understanding the springshed and flowpaths is a cr ucial baseline. Further work refining these concepts should remain a priority.118 UNKNOWNS AND RESEA RCH PRIORITIES

The preceding synthesis of nutrient loading, and, pe rhaps more importantly, how load is affected by hydrogeologic processes, identifies several key research gaps that need to be addressed in the coming years. I have broken these by theme and offer a brief rationale for each.

Hydrologic and Geochemical

1. Characterizing temporal dynam ics of water age estimates . It is clear from the literature on this topic (Katz et al. 2001, Toth and Katz 2006 to cite but a few examples) that there remain large uncertainties about the age mixtur e of water discharging from spring vents. Given the discordance between observed dye tr acer residence times and nominal ages, a research program to better understand wa ter age over time seems relevant. Decomposition, auto-correlation and cross-corr elation analyses of resulting time series should enable improved understand ing of flow regimes and al so links with other water quality variables that are str ongly relevant to springs ecol ogy (nutrients, salts, dissolved oxygen). Specifically, if binary mixing models are useful simplifications of spring vent hydrology, time series will allow decompositi on of flows into two components, with insight into the relative importance of each, a nd changes in mixing with regional climatic trends. These methods need to be applied co gnizant that different springs are likely to have different hydrogeologic prope rties, and inference from one spring may not apply in others.

2. Spring Flow variability and vulnerability. Ongoing efforts to set minimum flows and levels (MFLs) should consider the role of fl ow in what are genera lly considered to be water quality response variables. For example, changes in nitrate th at are implicated in springs ecological change have occurred along with changes in fish access, vent discharge, and salinity, among othe r things. Moreover, in adv ective systems, flux is often a better predictor of ecologi cal enrichment than is con centration, and with declining flows, fluxes are affected. Understanding the ro le of flow both direc tly and indirectly in springs ecosystems, and linking that to how wa ter has been managed in the springshed is likely to be high value research.

Nutrient Loading

1. High resolution temporal dynamics – Among the key implications of the literature is that pollutant delivery in karst systems can be highly variable, though less so in eogenetic karst. Pulses of nutrients and other wate r quality parameters (dissolved oxygen, salts, toxins) have been invoked for catastrophic shifts in other systems, particularly when ecosystem drivers have broadly changed. Th e presence of such pulses (even long duration anomalies might be considered pu lses) in spring system s is unknown, and the implications on springs ecology equally uncertain.

2. Improved N isotope tracing – There is significant uncertain ty about the interpretation of bulk stable isotope measurements in complex ka rst hydrologic systems. First, there is a119 broad omission of N fixation fr om N budgets (which would te nd to lighten the isotopic ratio of overall N delivery). Second, nitrific ation processes are poorly understood: since that process results in large kinetic fracti onation (lightening) an d denitrification is negligible in the aquifer, the potential impact on the overall N isotope ratio is both unknown and of crucial importance for understanding flow paths. This is also relevant for questions beyond nitrogen de livery, since ther e is a plausible connection between loading of organic wastes and declines in dissolved oxygen. Constraining the isotopic kinetics in springsheds by observations along flowpaths will likely improve understanding of water a nd pollutant sources.

3. Water Quality beyond nitrogen. The synthesis of the literature on springs is dominated by information about nitrogen, but essentially de void of similarly deta iled studies of other water quality parameters. Where surveys ha ve been done across a broad-spectrum of water quality attributes (Stevenson et al. 2004, for example), attributes such as salinity, flow, dissolved oxygen and phosphorus were at l east as predictive of ecosystem change as nitrogen, despite the dramatic enrichment of N that has been observed. There is a clear and present need to understand how variability in these attri butes occurs, a nd the extent to which that variability is in response to anthropogenic activities.

4. BOD Loading. Septic tanks, animal waste dispos al systems and sprayfields generate enormous quantities of biological and ni trogenous oxygen demand if discharged unoxygenated into the groundwater. Indee d, one septic tank emits enough BOD and NOD to deoxygenate over 6 million L of 5 ppm of dissolved oxygen in UFA water per year. Since dissolved oxygen concentrations are strongly spatially variable, have significantly declined across a br oad spectrum of springs (Scott et al. 2004), and were the only parameter significantly associat ed with algal cover in Stevenson et al. (2004), understanding the roles of hydrogeologic a nd anthropogenic processes in low DO at spring vents is a first-or der research unknown.

Subsurface Nutrient Processing

1. Explanation of observed assimilative capacity. Efforts analogous to the springshed nutrient budgeting for watersheds in the northea stern US yield an estimate of 20% for the fraction of terrestrial load that is transported by rivers to the sea. In a karst environment where N movement is essentially unimped ed, conservative estimates of landscape attenuation are as high or higher (70-90%), without accounting for N fixation. Even where efforts were made to estimate th e load reaching the aquifer (MACTEC 2007), those estimates are 6 fold higher than what emerges at the spring vent. Elaborating the reasons for this discordance, and how it might be affected by development and enhanced by management, should be a first-order priorit y. Among the issues that will need to be dealt with are the intrinsic lags due to aqui fer residence times, and the relative importance of surface vs. vadose zone vs. aquifer processes.

2. Saturation of assimilative capacity. As land development increases, there may be loading thresholds that exceed the assimilative capacity of the landscape and lead to elevated load120 breakthrough. The potential for observing this effect is low for nitrogen pollution, but may be high for P. Ensuring that P satura tion fronts are not a dvancing towards spring vents is important for forecasting.

3. Management responsiveness – As we begin to understand the links between springshed activities and responses in sp ring ecosystems, evaluating the extent to which parameters are management-sensitive is relevant. Th is is relatively well understood for N (see Obreza 2002 for an assessment of BMPs on N delivery), but not at all for other water quality metrics in springsheds. If othe r parameters (P, DO, salts, flow) figure prominently in holistic springshed management, understanding how to best manage them is a clear priority.121 REFERENCES

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# 3

Biogeochemical Processes and Imp lications for Nutrient Cycling

Patrick W. Inglett, Kanika S. Inglett, and K. Ramesh Reddy Department of Soil and Water Science University of Florida

Summary The availability and cycling of nutrients is dete rmined by an interaction of physical, chemical, and biological processes in an ecosystem. This interaction of processe s, collectively known as biogeochemistry, is important as it determines th e forms, transformations, and ultimate fate of nutrients in a given system. This chapter fo cuses on biogeochemical processes in springs and spring runs with an examination of the role of springs as a potential ‘source’ or ‘sink’ for nutrients and pollutants.

The Literature: What is Known

• The effect of biogeochemical processes on spring water composition begins immediately upon emergence from the boil and continues thr oughout the entire length of a spring run. In this way, the biogeochemical character of a spring and its biota depend greatly on the chemical nature of the groundwater source.

• Key interfaces involved in elemental cycling in springs environments are biofilms, algal mats and hyporheic sediment zones. These are areas of high biological activity and as such, have the potential to di splay indicators of major proce sses and nutrient effects (e.g., through enzyme activities).

• A variety of processes are known to oc cur in springs including organic matter decomposition, nitrogen fixation, nitrificati on, denitrification, iron reduction, sulfate reduction, and associated pro cesses involving phosphorus. The relative importance of each process depends greatly on the conditions of individual springs.

• Current data have increased our understand ing of nutrient sources, and have provided indications of some nutrien t transformations such as overall system productivity, potential biotic uptake and tran sformation of inorganic nutrien ts, and potential losses of nutrients from the system (e.g., gase ous losses or sediment storage).

• Elevated levels of nitrate in Florida springs are currently a major concern, especially due to the suggested correlation of nitrate with increased algal growth. There is clear evidence of elevated nitrat e levels in many springs; howev er, there is a high likelihood that the effect of increased nitrate may be coupled to other nutrient levels (e.g., phosphorus or metals) or physico-chemical parameters (e.g., dissolved oxygen).136 The Literature: What is Not Known

• Despite our understanding of bioge ochemistry in other aquatic systems, there is little information regarding biogeochemical processe s in spring systems where the focus has mainly centered on water column nutrients/wa ter quality indicators. Consequently, our current biogeochemical models of Florida sp ring systems are more of a black-box with emphasis on water column nutrient levels a nd little understanding of individual processes or differences between systems of various types.

• There is little information to ascertain th e involvement of proce sses/cycles controlling nitrate levels (e.g., plant uptake versus denitrif ication) or alternativ e pathways affecting nitrate conversion and loss (e.g., iron a nd sulfur involvement in lithotrophic denitrification).

• Among the studies which could advance our un derstanding of springs as sinks sources and transformers of nutrients are:

o Expanding water quality data collection to include additional geochemical parameters (e.g., sulfide) could improve our ability to e xplain and predict individual spring responses to nutrients.

o Increasing the frequency of water quality data collection (e.g., diel patterns) will offer additional insight into our ability to id entify pathways involved in spring processing/cycling of nutrients.

o Basing the collection of water quality data on stream segments will improve our ability to understand processes and sp ring run characteristics contributing to nutrient uptake and transformation.

o Characterization of sediments (e.g., nutrien t storages, organic matter distributions, mineral composition) would enable better assessments of benthic and hyporheic processes affecting both existing and future water quality (sediment memory).

o Aquatic plant management efforts are likely to result in significant alteration of the spring environment through episodic de position of organic matter, however, as yet there are no studies to document the potential for these activities to alter biogeochemical conditions (i.e., crea ting anoxic zones) and cycles (e.g., increasing denitrification loss of nitrat e or increasing releas e of ammonium and phosphorus).

o More studies are needed to document the involvement of mechanisms in observed declines in nitrate with distance in spring runs (i.e., dilution vs . biotic uptake vs. denitrification).

o More exploratory work is necessary to determine the existence and importance of novel pathways of nitrate conversion (e.g., anammox and lithotrophic nitrate reduction).137 INTRODUCTION

In combination with light, nutri ent availability is a dominant factor controlling aquatic productivity. For this reason, nutrient levels, transformations, and fates are key factors in determining the health of aquatic resources such as spring ecosystems. Increased levels of nutrients in many spring systems are also a grow ing concern, particularly in Florida where a significant portion of the groundwat er interacts with karst geol ogy. This interaction gives a unique character to Florida springs as ecosystems, and has important implications for the ways in which these systems respond to, process and cycle nutrients.

The study of biogeochemistry deals with the inte raction of biological systems with geochemical reactions which determines the overall cycling of elements within ecosystems. Both abiotic and biotically-controlled geochemical reactions can a ffect nutrient and mineral solubility, as well as reduction/oxidation (redox) chem istry, which can affect the co mposition and activity of the biotic system which in turn affects the water ch emistry. Whether a spring system is a source or sink for nutrients and pollutants in the landscape is de termined by the coupled biogeochemical processes occurring in its water column, se diments, biota, and adjacent floodplains.

The potential biogeochemical reactions and path ways occurring in spri ngs are numerous, and have been studied in varying levels of detail. Therefore the purpose of this review is 1) to describe the biogeochemical processes related to nutrient cycling in springs or similar systems, 2) summarize the available literature for Florida spring systems, and 3) highlight the needs and directions for future spring research. The major literature sources for this review were derived from published journal article s and government/agency reports . These documents included geochemical studies, biological/microbiological studies, and ecosystem/ ecology studies focused on spring systems including those from Florida and other springs around the world. An overall lack of spring-specific data on many processes for ced the use of literature pertaining to similar systems, such as chalk and other limestone-based streams. Finally, general stream literature was also used where appropriate to describe simila r processes occurring in lo tic (flowing) systems.

SPRING ENVIRONMENTS/ORGANISM GROUPS

Spring water chemistry

Water chemistry is one of the most important ch aracteristics of a spring system which directly impacts its biological functioning. Because the primary source of spring water is groundwater, the characteristics of the different aquifers such as mineralogy and residence time will determine the overall chemical character of the spring receiving this outflow. Slack and Rosenau (1979 classified Florida springs into four types, in order of occurrence: 1) calcium-magnesium bicarbonate, 2) sodium chloride, 3) mixed (no dominant cation or an ion) and, 4) calcium sulfate. This review will focus on the dominant spring wate r type in Florida - the hard water type, which has high calcium-magnesium-bicarbonate concentra tions resulting from the interaction with limestone and dolostone host rock. However, other chemical constituents characterizing the other spring water types will also be addressed as they relate to specific processes.138 Groundwater confined within the aquifer for extended periods of time undergoes significant changes as it emerges at the spri ng vent or boil and is directly introduced into conditions at the earth surface. Groundwater (at mo re or less equilibrium with subs urface temperature/ pressure/ mineral compositions) changes dramatically wh en introduced into the high oxygen/reduced pressure environment at the boil/vent/or seep 

### ( Figure 1 

### ) . In general, soil/aquifer respiration processes increase the partial pressure of CO 2 in ground waters, which in limestone systems is responsible for the dissolution of the aquifer be drock to form karst features. The equilibrium reaction for this process is given by the equation:

CaCO 3 + CO 2 + H 2 O Æ Ca 2+ + 2HCO 3 -

Dissolution of limestone and dolos tone results in high levels of alkalinity (carbonate buffering) and dissolved constituents (e.g., ca lcium, magnesium, chloride, sulf ate) which give rise to the chemical composition of most sp ring waters in Florida.

Figure 1 – Depiction of degassing and oxidative c onversion of chemically reduced species (Red.) to oxidized (Ox.) forms following emerge nce of groundwater in a spring vent/boil.

Emergence of groundwater at a boil or vent resu lts in an immediate disequilibrium of the water with respect to CO 2 , and as a result, degassing occurs. Loss of excess CO 2 shifts the dissolution equation toward pr ecipitation of CaCO 3 which results in supersat uration of groundwater with respect to CaCO 3 (Dreybrodt et al . 1992). Other gaseous constituen ts of groundwater have the potential to degas upon emergence including N 2 (produced through aquifer (denitrification), the greenhouse gas N 2 O from (from incomplete nitrification/ denitrification proc esses), and caustic H 2 S from sulfate reduction (evidenced by a rotten egg smell).

Supersaturation of CaCO 3 may further result in CaCO 3 precipitation and mineral formations near boil or seep areas (Drysdale et al . 2002; Liu et al 2006). This phenomenon is not common in Florida, but the process of CaCO 3 precipitation can result in mineral encrustation in localized139 areas of some springs, especially where proc esses such as photosynthesis further enhance supersaturation through uptake of CO 2 . This process can result in mild calcifications on surfaces of submersed aquatic vegetation (SAV ) or algae filame nts (Berrie 1992).

If sufficient amounts of organic materials are pr esent in the groundwater, respiration processes within confined aquifers resu lt in a depletion of available oxygen potentially creating anaerobic conditions. Many springs report dissolved O 2 levels <1 mg L -1 (Scott et al. 2004). Under anaerobic conditions, other oxidized species (e.g., NO 3 - , Fe(III), and SO 4 2- ) can become chemically reduced (forming e.g., NH 4 + , Fe(II), S 2- ). In extreme cases, the highly reduced gas, methane (CH 4 ) can even be formed. As a result, many groundwaters can contain a number of reduced species under low-oxidizing conditions. These reduced species are highly reactive, however, they can also persist in groundwaters (even under aerobic conditions) for significant periods of time (Darling and Gooddy 2006).

At the boil, chemically re duced spring waters encounter atmospheric levels of O 2 leading to a suite of chemical oxidation reactions:

S 2- + O 2 + 2H 2 O Æ SO 4 2- + 4H +

NH 4 + + O 2 + H 2 O Æ NO 3 - + 6H +
4Fe(II) 2+ + O 2 + 10H 2 O Æ 4Fe(III)(OH) 3 + 8H +

CH 4 + O 2 Æ CO 2 + 4H +

Two important consequences of each of these oxid ation reactions include: 1) consumption of O 2 and 2) the production of H + ions. Depending on the redox status of the groundwater, consumption of oxygen may continue even afte r springwater emergence. As groundwater species are oxidized, the resulting acidification al so increases the potential for dissolution of CaCO 3 (in systems supplied with highly re ducing groundwater) while degassing and photosynthesis (reducing CO 2 levels) results in precipitation of CaCO 3 (Guasch et al . 1998). The balance between these two extremes determin es the extent of calcification processes which is an important determinant of vegetative specie s composition (particularly algal assemblages) in spring systems (Freytet and Verrecchia 1998).

Reactive Zones

The high energy, lotic characteris tics of springs create a vari ety of different zones where biogeochemical reactions can occur. The most im portant zones are the interfaces of the water column with sediments, epilithic/periphytic biofilm s, and benthic mats of macroalgae (Figure 2). These interfaces are critical to nutrient cycling processes because of the sharp gradients they create in physical and chem ical parameters (e.g., pH, O 2 , redox). In addition, solutes are actively produced and consumed creating concentrati on gradients which further enhance solute movement and transport.140 Biofilm

In aquatic systems, development of structural microbial assemblages, sometimes referred to as biofilms, is influenced by the habitats which ma y be present. In a stream or a spring run, the presence of rocks and stones, li tter, detrital material, logs, plan ts and other submerged surfaces provide substrates for the microbial biofilms to establish. In well-developed stages, aquatic biofilms are known by names such as periphyton or aufwuchs.

Figure 2 – Depiction of biogeochemical exchanges o ccurring at the biofilm, epiphyte, and benthic algal mat surfaces in a spring system.

Biofilm forming algae and bacteria actively exude substantial quantities of exopolymeric substances (EPS) into the environment th at are primarily organic carbon (Hoagland et al. 1993; Costerton et al. 1995; Goto et al. 1999). This exudation of carbon can constitute a large proportion of carbon acquired by bot h bacteria and algae (Goto et al. 1999) and result in the accumulation of non-living organic carbon on surfaces in aquatic environments (Wetzel, Ward & Stock 1997; Frost & Elser 2002).

In aquatic systems like springs, biofilms are domin ated by algae, diatoms, bacteria, and fungi. Exuded C provides the basis for other heterotrophi c organisms to colonize, and as the biofilm develops, respiration can consume available oxygen to create anaerobic zones within the biofilm. The creation of anaerobic microsit es and layers allow a myriad of other bacterial groups to become established. These groups represent important classes of biogeochemical processes141 including denitrifiers, other anaerobic (iron reducers, sulfat e reducers), and other aerobic organisms (nitrifiers, iron /sulfide oxidizers).

Benthic Algal Mats

Conspicuous mats of benthic algae are beco ming a common feature of many spring systems (Stevenson et al . 2007). These macroscopic algal growth s, similar to biof ilms, are actually communities of microorganisms incl uding: algae, bacteria, and fungi . The initial formation of mats is usually based on the structure provi ded by filamentous algae, which provide the framework for accretion of other organisms (e.g., diatoms, bacteria) and additional organic and inorganic materials (Stal 2000). The two most co mmon algal taxa forming these mats in Florida systems are Lyngbya and Vaucheria , however a variety of other f ilamentous or colonial species can initiate mat formation (Stevenson et al . 2007).

As the algal filaments develop, accretion progresses, and secondary organisms begin to dominate and alter the mat’s physical and chemical characteri stics. Physical gradie nts (light, temperature) form as light penetration to deeper mat layers becomes limited (Jorgensen et al . 1983). In response to the physical gradients, photosynthetic mat organisms b ecome structured (or in some species, migrate vertically) to maximize their locati on in the presence of li ght and temperature. (Stal 2000). With the localiza tion of C fixation, nutrient grad ients develop in response to increased autotrophic and subse quent heterotrophic growth. Gr adients of reduced and oxidized compounds also form and drive mat structure a nd metabolism (Figure 3) (Pinckney and Paerl 1997; Stevenson et al . 2007). These gradients of nutrients and redox can occur vertically or in response to diel fluctuations of photosynthetically-induced o xygen, pH and C production. As a result, the character (i.e., species composition and physical characteris tics) and associated functions of an algal mat can be drastically diff erent both vertically and temporally within the
mat.142 Figure 3 – Illustration of the sampling device and examp le of the measured profiles of nutrients from the study of benthic macroalgal ma ts by Stevenson, et al. (2007).

Sediments

In addition to a dissolved load, springs from ca rbonate aquifers have also been shown to be significant sources of sediments (Herman et al . 2006). In aquifers based in marine deposits (such as in Florida) these particles may be de rived from the direct su rface connections with underground conduits, or from part icles release through erosion or dissolution of the aquifer material itself. As a result, sediments in spring flows can vary in composition from small organic materials, to clay particles, to sands and carbonate minerals (Herman et al . 2006; Mahler and Lynch 1999). The direct transport of surface derived materials also results in a strong dependence of sediment flux at the boil on the springshed characteristics and relationships between rainfall, runoff, recharge, and discharg e. For example some springs, normally clear under low flow conditions, become turbid after storm events.

Not much is known about the composition and distri bution of sediment types in Florida springs and spring runs, including informa tion regarding particle size, mineralogy, or organic content. To our knowledge, only one study has attempted to characteri ze sediments/subaqueous soils from a spring system (Saunders 2007). In other stream environments, flow velocity is the most important determinant of sediment distribution pa tterns, therefore, it is likely the same would hold true for springs with flow velocities ranging from ca 1 to 100 cm s -1 . In general, particle sizes of sediment should decreas e with passage downstream as channel widths widen, current velocity decrease, and progressively finer size classes are allowed to settle out. Saunders (2007) confirmed this pattern for the upper Chassahowitz ka River where the areas near the boil were dominated by sands low in organi c matter, and areas further downstream consisted of finer sands with higher organic content.143 The main biogeochemical feature of spring run se diments is the presence or absence of redox gradients. In sediments, microbial re spiration frequently consumes available O 2 leading to the formation of anaerobic zones at depth. Under these conditions O 2 contained in the water will diffuse downward to the anoxic si tes, creating a gradient of O 2 and the presence of both aerobic and anaerobic sediment zones. In aerobic zones, microbes are involved in biogeochemical reactions and use oxygen as the primary electron accep tor, while in anaerobic layers, facultative and obligate anaerobes function using alternate electron acceptors (e.g., NO 3 - , Fe(III), and SO 4 2- ) in their respiration.

One feature of sediments in high energy lotic syst ems is the high percentage of coarse particle size classes (e.g., sands) which greatly enhances a dvective processes (Figure 4). Sediments of high energy spring runs are also more likely to include varying st ratigraphies with layers of different composition (e.g., organi c matter) juxtaposed. In a manner analogous to that in biofilms or benthic mats, these types of sediment distributions serve to enhance the potential for varying reactions where reacta nts and products are more easi ly exchanged between reactive zones to greatly increase reacti on rates (Harvey and Wagner 2000).

Characteristics of each of the se diment habitats in spring syst ems give rise to a variety of functional groups of organisms. Varying redox a nd nutrient conditions in sediments of springs likely result in the coexistence of both aerobic an d anaerobic genera. This increased diversity of organisms and metabolisms should enhance the potential biogeochemical cycling of nutrients such as carbon, nitrogen, and phos phorus. Bacteria, in particul ar, are constantly responding and adjusting to carbon and nutrient av ailability, and thus, their activit ies can indicate conditions of availability. One of the ways microorganisms react to environmental conditions is through expression of enzymes which pl ay a pivotal role in catalyzing the rate limiting steps of biogeochemical processes. For this reason, enzyme assays can serve as excellent indicators of microbial activities and environm ental conditions (e.g., nutrient/conta minant impacts) of aquatic systems (Chrost 1991; Sinsabaugh et al . 1997).

Figure 4 – Diagram of characteristics and hyporheic fl ow processes related to biogeochemical functions in a hypothetical spri ng run sediment profile. H, hy porheic zone; OM, organic matter concentration; GW, groundwater movement.144 BIOGEOCHEMICAL PROCESSES

Carbon

As the major component of all life, carbon is arguably the most important of all biological elements in ecosystems. The cycling of C in spring systems has important implications for productivity, food-web dynamics, and metabolism of all spring biota. Consequently, the biogeochemical cycle of carbon involves complex interactions between and within organic and inorganic C reservoirs and invol ves numerous metabolic pathways and organism groups. These interactions have been widely studied by scientists in various disciplines.

Carbon Sources

The main sources of C to a spring system occur as 1) C sources in the groundwater; 2) C from direct or diffuse inputs of surface or groundwater along the spring run; 3) C inputs derived from riparian vegetation (litterfall) ; and 4) C fixed into biomass of spring autotrophs (primary producers). Of these, the C derived from auto trophic production within the spring run would be considered autochthonous (within-s ystem) sources while the remai nder are C sources external to (or allochthonous to) the spring run.

In aquatic systems, carbon is predominantly pr esent as dissolved inorganic carbon (DIC), dissolved organic carbon (DOC), and as particulate organic (P OC) and particulate inorganic (PIC) carbon forms. DIC is the primary form of C in hard water (limestone-based) springs, and is the primary form of C available for photosynthes is. For this reason, DIC is rarely limiting to autotrophic production in springs such as those in Florida. In contrast, ground waters in karst systems generally have low concentrations of DOC, and thus, much of the secondary production in springs is likely derived from autochthonous sources of C either directly, or following decomposition (Berrie 1992).

Perhaps one of the largest sour ces of C to spring systems is through the connections of many spring runs with extensive floodplains with fo rested, high-C soils (Whiting and Stamm 1995). This becomes significant during high flow periods which could result in significant exports of floodplain particulate C or leaching of C and subsequent increases in spring DOC concentrations. Hydrologic fluctuations (i.e., varia tions in rainfall and spring flow/d ischarge) can also lead to the introduction of variable amounts of high DOC su rface water into groundwater systems (Pronk et al. 2006). Under high flow conditi ons there are also greater input s of C from connected streams and discharges of surface or drainage waters.

As they are known to actively exude C compounds , both algae and SAV ar e likely important as autochthonous C sources in springs. These e xuded compounds can range in size from low molecular weight (LMW) compounds such as sugars, proteins, amino acids, enzymes, phenolics, glycolate, and alkaloids (Sondergaard 1983) to high molecular weight (HMW) compounds (>10,000 Daltons, where a dalton ≈ the mass of one hydrogen atom) including primarily polysaccharides which can appear as slimes or algal sheathes (De Philippis and Vincenzini 1998; Stal 1994; Wetzel and Penhale1979). The com position and production of extracellular C is145 specific to both organism type and growth conditions includ ing culture age, N source, temperature, salinity, P, and potassium (reviewed by De Philippis and Vincenzini 1998). Net release rates of 5-10% of total carbon fixatio n in macrophytes have b een observed (Wetzel and Penhale 1979), while in some species of cya nobacteria, extracellular C production can represent up to 50-70% of dry weight (Bertocchi et al. 1990). Production of EPS has also been shown to increase with increasing irradiance in algal cultures (Otero and Vincen zini 2003; Renstrom and Bergman 1989). For this reason, lig ht availability, and rates and pa tterns of algal photosynthesis/ photorespiration, can be a significant determinan t of the quantity and quality of C compounds released within a spring system.

When present, EPS is considered a bioavailable C source that can stimulate bacterial respiration and productivity (Neely a nd Wetzel 1995; Espeland et al . 2001). Consequently, microbial associations have been shown to alter the co mposition of products excreted by the source plant (Sondergaard 1983). Overall, exuded C sources are viewed as highly available, particularly for the microbes associated with the algal cells or SAV leaves. In order to utilize exuded C substrates, bacteria mu st hydrolyze complex polymeric OM in to simple utilizable forms by producing extracellular enzymes, in particular glucosidases (Chrost 1989; 1991; Eivazi and Tatabai 1988; Arsnoti 2003).

Decomposition

In the absence of direct exudation of C com pounds, secondary production is based on C derived through decomposition. In most aquatic systems, th is is the dominant path way of C flow, and is the primary influence on the me tabolic activity and diversity of microbes conducting many other biogeochemical processes. Springs are not known as accumulators of organic matter, perhaps due to their high flow velocity and (in most cases) continuous supply of well-oxygenated water. Temperature is seasonally controlled (20-24 oC), and so decomposition should occur at a moderate rate throughout the year. Also, the well -buffered hard waters of most spring systems should prevent low pH conditions from inhibiti ng microbial decomposition activity as in acid bogs.

As in other lotic systems, organic matter in springs tends to accumulate in dense vegetation stands, debris dams, and at the bottoms of la rge pools. In a slow moving spring run (upper Chassahowitzka River), Saunders (2007) observed la rge (>5 cm) organic material horizons with a patchy distribution. Similarly, in chalk streams, seasonal changes in beds of aquatic vegetation were shown to strongly affect flow veloci ties and rates of sediment and organic matter accumulation (Cotton et al ., 2006). For this reason, seasonal ity and management of SAV can have a direct impact on distribution of zones of organic matter accumulation and the biogeochemistry of a spring system. Aquatic vegetation management, including spraying of herbicides (e.g., Wakulla Springs State Park) can also contribute a significant amount of C (in large pulses) to spring runs.

Decomposition (or C mineralization) is a co mplex process beginning with the physical breakdown of macroscopic plant/algae tissues. Th e major constituents of this detrital plant matter are soluble substances (such as sugars , fatty acids, and amino acids), structural compounds (such as cellulose, hemi-cellulose, li gnin and associated pr oteins), other water-146 insoluble components (including fats, waxes and re sins), and mineral matte r (ash). The relative proportion of these constituents varies with type and source of detrital matter, degree of decomposition, and the age of the material. While SAV in springs contai n more soluble and less structural C compounds than vegeta tion in the riparian zone, algae contain even fewer structural compounds. Approximate ranges in values for these constituents in plant ma terial, peat, and soil organic matter are shown in Table 1 (Reddy and DeLaune 2008).

The initial step in the decompos ition process is leaching of wate r soluble organic and inorganic compounds as a result of cell autolysis. Water soluble components include soluble carbohydrates, nucleotide bases, fatty acids, a nd amino acids. These components are readily bioavailable and serve as energy and nutrien t sources to bacterial and fungal communities colonized on detrital plant tissue. Physical leac hing of soluble components is generally complete within a few weeks of plant se nescence, depending on structural components, temperature, and water flow. Leaching can be a significant f actor in decomposition of non-lignocellulose materials such as algal mats and SAV (>50% mass loss). However, in detrital tissue containing lignocellulose materials (such as tree leaves and woody materials) only approximately 10-20% is lost within four weeks after imme rsion in the water column (Benner et al . 1985).

Table 1. Major carbon components of plant ti ssue, peat, and soil organic matter.

Component Plants Peat SOM Monomer Linkage
% composition

Cellulose 15-60 15-30 2-10 Glucose β -1,4 ether Hemi-cellulose 10-30 10-40 0-2 5C & 6C sugars β -1, 4 & β -1, 3 ether Lignin 5-40 20-60 30-50 Phenyl propane Aryl ether & C-C Water soluble fatty acids, sugars, amino acids 10 <1 NA NA

Ether extractable lipids, oils, waxes, etc. 1-8 2-10 fatty acids, sugars, phosphate Esters & C-C

Proteins 2-20 4-20 2-20 amino acids Peptide Mineral matter 1-10 5-25

After initial leaching and fragmentation, breakdown of plant detritus depends on the amount of major polymers, including lignin, cellulose, hemi -cellulose, lipids and waxes, and proteins. These structural components are mainly particulat e in nature and must be broken down. In other aquatic systems, macroinvertebra te colonization is a key factor affecting observed litter decomposition rates; however, studies of their e ffect in springs are few and inconclusive. For example, in a study of four spri ngs in Minnesota, Bortodziej and Perry (1990) found similar rates of decomposition of watercress ( Nasturtium officianale ) in different spring types despite significant differences in macroinve rtebrate colonization. In the same study, they also found that macroinvertebrate colonization was correl ated with decay rates of Boxelder ( Acer negundo ). Results such as these may indicate that macr oinvertebrate shredder and grazer function may147 depend on plant biomass type/quality (e.g., high vers us low contents of structural compounds) in high energy spring environments.

Decomposition of particulate matter into smaller particles is an im portant process in determining C availability for secondary production and biogeoch emical metabolism in spring systems. Plant detritus (either attached or de tached from the plan t) undergoes physical fr agmentation through the action of waves and currents, UV expos ure, and through gr azing activities of macroinvertebrates (Dawson 1980). Fragmentation of plant detritus into fine particulate matter (FPOM) fraction (<1 mm) results in increased surface area, and accelerated microscale processes (such as enzymatic hydrolysis and catabolic activiti es). Coarse particulate matter (CPOM) being reduced into smaller fractions greatly enhances the surface area which fungi and bacteria can attack. In lotic environments, the action of macr oinvertebrate shredders results in a transport of C particulates from the headwate r reaches and a steady accumulati on of fine particulate organic matter (FPOM) at lower stream zones (Webster et al . 1999). As it travels downstream, this FPOM material becomes increasingly more bi oavailable through continued decomposition and the incorporation of added nutrien ts and algal derived particulates (Allan and Castillo 2003).

Even FPOM must first be converted into smaller metabolizable units before it can be assimilated by microbes as energy and nutrient sources. Extracellular enzymes secreted by microbes (bacteria and fungi) aid in hydrolysis of high mo lecular weight organic compounds associated with plant detritus and soil organic matter, which in their original form ar e too large for transport into the periplasmic sp ace of the cell (Weiss et al . 1991). Microorganisms in cluding bacteria and fungi are the dominant producers of extracellula r enzymes. Selected examples of enzymes involved in degradation of cellulose, hemi- cellulose, and lignin are shown in Table 2. Breakdown of these structural polymers may re quire a multicomponent enzyme system including a wide range of microbial specie s. Rarely can a single micr oorganism produce all the enzymes required for breakdown of organic substrates.

Extracellular enzymes are generally considered as the rate limiting step in the overall degradation of organic matter. Kinetic limita tions to enzymatic br eakdown include slowing diffusion of enzymes and substrates, adsorption of enzymes on solid phases, and complexation of enzymes to humic substances (Sinsabaugh 1 993; Wetzel 1991). Co mplexation of humic substances with enzymes may not be a significant issue in hard water springs as major cations (e.g., Ca 2+ ) react with humic acids and reduce the precip itation and inactivation enzymes. In soft water with low levels of base cations, the interference of humic substances on enzyme activities may be greater than hard water with Ca 2+ and Mg 2+ concentration in the range of 40-60 and 15- 25 mg l -1 , respectively (Wetzel 1991).148 Table 2. Enzymes involved in the degradation of polymeric C compounds (Sinsabaugh et al. 1991). EC number = En zyme commission number. Process Enzymes EC number Substrates
Starch degradation α,β,γ -amylase 3.2.1.1,2,3 Amylose Cellulose degradation Exocellulase Endocellulase 3.2.1.4 3.2.1.9.1 Cellulose Cellulose β -glucosidase 3.2.1.2.1 Cellobiose

Hemi-cellulose degradation Exoxylanase Endoxylanase 3.2.1.8 Xylan Xylan β -xylosidase 3.2.1.37 Xylobiose

Lignin degradation Phenol oxidase 1.10.3.12 Lignin Peroxidase 1.11.1.7 Lignin

Because enzyme activity is the rate limiting step in microbially mediated decomposition of plant detritus in wetlands and aquatic systems, many researchers have studied enzyme activities over scales ranging from the molecular to the ecosystem level (Sinsabaugh et al . 1993). Activities of β -Glucosidase, endocellulase and exocellulase have been measur ed as indicators of organic matter decomposition in wetlands and aquatic system s. Activity should be on a relative basis, when comparing the results from various ecosystems. Despite their use in other aquatic systems, there has not been any published report of car bon degrading enzyme expression in Florida springs.

As decomposition progresses, soluble or easily hydrolyzed compounds are selectively degraded yielding increased recalcitrance with age of detritu s. Decomposer biomass is also incorporated as populations wax and wane. In this way, or ganic matter in waters and sediments can be viewed as a complex of plant, microbial, and animal products in various stages of decomposition. Continued chemi cal alteration of the C mixtur e ultimately results in the formation of highly reca lcitrant C classes including humic and fulvic acids (Stevenson 1994). This process of stable organic matter formation results in aquatic systems serving as accumulators of C, and in springs, this process can be accentuated by increased deposition and burial of organic matter by sediments.

Nitrogen

Along with P, the high abundance of N in living biomass makes it one of the most highly demanded nutrients in aquatic sy stems. Nitrogen levels in ma ny springs statewide have shown steady and dramatic increases over the past ha lf century. In many cases, this excess N could result in significant changes in productivity; ho wever, not all springs have exhibited changes with increasing N concentrations. The high level of N in some springs systems is also cause for significant concern (even if not related to w ithin-spring impacts) from toxicological and landscape perspectives.149 Numerous transformations and a suite of enzymes are involved in the transformation of N in aquatic systems (Table 3, Figure 5) . Nitrogen is present in aquatic systems in organic forms and as inorganic forms such as dinitrogen gas (N 2 ), nitrate (NO 3 - ), nitrite (NO 2 - ), ammonium (NH 4 + ), and trace gases such as N 2 O. Organic N is derived primarily from amino

- bonded C (-NH 2 ) such as amino acids and proteins, an d nucleic acids, amino sugars

. Generally, the sum of NO 3 - /NO 2 -
, and NH 4 + measured after filtration is taken to represent the dissolv ed inorganic N (DIN). The difference between the total dissolved N of a system and its DIN is then defined as the dissolved organic N (DON).

Table 3 . Enzymes involved in various nitrogen cy cle processes. Fr om Reddy and DeLaune (2008).

Process Enzymes EC number Substrates

N 2 fixation Nitrogenase 1.18.6.1 N 2 gas Amino acid degradation Amidase/ deaminase 3.5.1.4 monocarboxylic acid amide Amino acid degradation L-Glutaminase 3.5.1.2 L-glutamine
Urea degradation Urease 3.5.1.5 Urea Nitrate reduction Nitrate reductas e 1.7.99 nitrogenous substrates Protein degradation Peptidase 3.4.21 Amides

Biological Nitrogen Fixation

Nitrogen is unavailable in the gaseous N 2 form and must be first conve rted (or fixed) to become bioavailable. One of the most studied enzymes (nitrogenase) is responsible for this conversion of N 2 gas to NH 4 + , and is possessed by only select prokary otic organisms including bacteria and cyanobacteria. Both of these groups are presen t in the biofilms, sediments and benthic algal mats of spring systems, however the potential for their expression of nitrogenase is uncertain.

A variety of factors can influence expressi on of nitrogenase enzymes including species composition, light availability, ava ilability of P and me tallic co-factors (Fe, Mo), oxygen levels, and NH 4 + concentrations. In the case of the springs, NH 4 + levels are genera lly not high, and should not be sufficient to suppress nitrogenase. Metal availability (in pa rticular trace elements required for nitrogenase function) may be limiting in some springs; however, there are little data to support this conclusion. The presence of cya nobacteria in spring systems indicates that Nfixers may have a competitive advantage for N ac quisition. It is true that many springs are Nlimited, therefore, it is also likely that active N 2 fixation may be occurring. This has been documented in a number of stream systems; howev er, no data exist to ascertain the significance of this process for Florida springs. Furthermore, it is also uncertai n what the fate of this fixed N would be in the spring environment.150 Nitrogen uptake

In streams and springs with high biomass and produc tivity, nitrogen required for growth that is not provided by nitrogen fixation must be supplie d by N uptake and assimilation from the water column. High production of algae and SAV in Florid a spring runs should also contribute to high N assimilation rates especially in areas where gr owth is N limited (Notestein et al., 2003; Inglett et al., 2007). Of the two inorga nic dissolved N forms, both NO 3 - and NH 4 + are bioavailable and have been shown to be rapidly taken up by biota in headwater str eams. Much of this headwater uptake is likely due to the high a bundance of organic matter (leaf litt er) in these systems. In one study of a low organic matter, brackish portion of the Chassahowitzka Ri ver, Saunders (2007) calculated rates of N loss (as NO 3 - ) from the water column as high as 5 mg m -2 d -1 , however; it was not determined what proportio n of this N loss was due to biot ic uptake, denitrification, or dilution.

Figure 5 – Idealized diagram of the nitrogen cycle occurring in a typical spring run system.

In springs it is unclear whethe r the abundance of N in the water column is the main source of N to the plant/epiphyte communities, or if SAV derive the majority of their N through root uptake. Several studies have noted that both pathways ma y be equally important for productivity in SAV (compare Madsden and Cedergreen 2002 and Xie et al . 2005). Both NO 3 - and NH 4 + are considered bioavailable; however, NO 3 - must first be reduced to NH 4 + before it can be assimilated into biomass. The enzyme, nitrate reductase is functional during this process, and thus, can be used to show active incorporation of NO 3 - in vegetation (Melzer and Kaiser 1986). At least one study has used this enzyme activity to assess potential NO 3 - use by biota in springs151 with the findings that 1) higher water column NO 3 - levels resulted in higher levels of NO 3 - reductase activity and 2) enzyme expression in biofilms differed depe nding on substrate type (Inglett et al . 2007) (Figure 6).

In addition to inorganic N forms, it has also been noted that organic forms of N, such as amino acids and other small molecular weight N compounds (e.g., urea) can be a si gnificant source of N to both microbial/algal communities (e.g., Berman and Bronk 2003; Linares and Sundback, 2006) and wetland macrophytes. This is important not only because of its role in nutrition, but also because in many aquatic systems, the presen ce of organic forms of N (in particular, urea) has been shown to effect algal species compositi on (e.g., Heil et al. 2007). Thus, both quality as well as quantity of N may be important factors determining the impact or fate of N in springs.

Figure 6 – Nitrate reductase activity ( μ

mol NO 2 - produced g -1 dw h -1 ) in sediment and algal biofilms from various substrates (S, sedi ment; FB, filamentous benthic algae; BF, brown filamentous; E, epiphytic algae; FG, floating gr een algae; RW, rock wall biofilm; D, detrital material; SB, submerged bark biofilm) at four lo cations along the Wekiva River. (from Inglett et al. 2007).

Nitrogen mineralization

The biological transfor mation of organically combined ni trogen to ammonium nitrogen during the degradation of organic matter (ammonification) is one of the key mechanisms regulating the fate of nitrogen in aquatic systems. This pro cess is strongly related to organic C decomposition, since a major portion of N in soil organic matter and plant detritus is bonded to C. As organic matter is decomposed, organic N is mineralized and released as ammonium. This process is achieved by a series of microbially-produced en zymatic processes (e.g. cellulose, cellulobioase,152 glucosidase, aminopeptidase, protease etc.). Therefore, similar to C decomposition, the regulators of organic nitrogen mi neralization include substrate quality, microbial biomass and enzyme activity, temperature, redox potenti al and hydrology. In a spring run, sites of ammonification would coexist with sites of orga nic matter/detrital matter accumulation or where organic matter has a rapid turnover as in the biofilms or the algal mats.

Depending on microbial N requirements, minera lized ammonium may be released, rapidly utilized by the microorganisms (immobilized), or taken up by plants. During decomposition, the C:N ratio reflects the mineralization-immobilizati on pattern of nitrogen, where a high C:N ratio promotes immobilization, while a lower ratio fa vors mineralization. During immobilization nitrogen is considered to be limiting to decomposer organisms. For this reason, initial N content of organic matter is a strong determinant of d ecomposition rates. Thus, organic matter with a low C:N ratio (such as algae or SAV) is consider ed labile, while those with high C:N ratios (such as wood, or senescent riparian vegetation) are considered more recalcitrant. Based on these factors, the production or consumption of NH 4 + in the water column in a spring can be highly variable.

Ammonium in the water column is derived fro m (1) decomposition of organic matter in the water column; (2) mineralization of organic nitrogen in the aerobic soil/sediment layer; (3) mineralization of organic nitrogen in the biofil ms; and (4) diffusion and mass flow of ammonium from anaerobic soil layer to aerobic soil/sediment layer and water colu mn. Saunders (2007) found that sediments in the Chassahowitzka River were reducing at depth, while Stevenson et al. (2007) measured porewater NH 4 + levels >10 mg L -1 under mats of Vaucheria sp. Both studies indicate that sediments in a spring run system ca n act as a significant source of ammonium to the overlying aerobic soil layer and water column. Flux of ammonium is accomplished by diffusion, advection, bioturbation, and mi xing at and near the soil-f loodwater interface.

Once NH 4 + enters the water column, a variety of processes can affect its fate in a spring run/system. Ammonium is bioavailable, a nd therefore can be rapidly assimilated by phytoplankton, biofilm organisms, or SAV. Alte rnatively, if pH is high (>8) a significant proportion of NH 4 + will be in the form of gaseous NH 3 which can volatilize (flux out of the water column). Volatilization of NH 3 can be significant in highly productive systems; however, in flowing waters with a high buffering capacity, it is unlikely that losses via this pathway are significant.

By far, the most likely fate of NH 4 + in the well-oxygenated water co lumn of a spring run will be nitrification where microorganisms oxidize NH 4 + to NO 3 - . Autotrophic bacter ia are dominantly responsible for the process of n itrification, and therefore these organisms do not require organic matter to function. Nitrifying populations of bacter ia are likely to be f ound in all components of spring systems, and should be particularly conc entrated at the aerobic/anaerobic interfaces where NH 4 + and O 2 will both be available.

Denitrification

Nitrate has a high oxidation poten tial (similar to oxygen) and can be used by a variety of organism groups as an electron acceptor (beco ming chemically reduced). One of the most153 significant and highly studied proc esses reducing nitrate is the pr ocess of denitrification where organisms couple organic C oxidation to nitrate re duction. This processes is almost exclusively anaerobic (a few aerobic exam ples have been found). Many of the organisms capable of denitrification are facultative and are capable of surviving in transitional (moderately anaerobic) or fluctuating aerobic/anaerobic conditions. In fact, in many aquatic systems denitrification zones occur in close proximity to the aerobic zones of NO 3 - production (nitrification).

Denitrifying organisms are predominantly hete rotrophic. Therefore high organic matter C deposits (sediments, debris dams, biofilms, and al gal mats) are the likely zones for occurrence of denitrification in spring systems. Alternatively, turbulent flow in many springs would likely be sufficient to advect significant quantities of water column NO 3 - into anaerobic sediments in the hyporheic zone (Triska et al . 1993; Duff et al . 1998). Hyporheic zones include regions beneath the stream channel and regions adjacent to the channel where stream water is exchanging with interstitial waters at rates and scales larger than those limited by diffusion. The majority of springs are too well oxygenated for denitrification to proceed at any great rate in the water column; however, dense SAV or heavy biofi lm growths possess enough C and can sufficiently reduce water flow to indu ce anaerobic conditions either spatially (e.g., in microsites) and/or temporally (short-term transient events or diel cy cles). This would be pa rticularly important at night in the absence of photosynthesis.

To date, the only reported measurem ents of denitrification potentia l in Florida springs have been conducted in Wekiwa Springs /Wekiva River by Inglett et al . (2007). In that study, potential denitrification rates were measured in sediments and algal biofilms at four sites ranging from the headspring boil to 14.8 miles downstream (Figure 7) . Observed rates ranged from 10 nmol N 2 O produced gdw -1 h -1 in surface sediments (0-10 cm) to 170 nmol N 2 O produced g -1 dw h -1 in biofilms collected from bark samples from the headspring. The ra tes measured in the Wekiwa/Wekiva system are similar to DEA ra tes reported in wetla nds (0.5 to 20 mg N 2 O kg -1
soil h -1 ; White and Reddy 2003), estuarin e sediments (15-130 nmol N cm -3 d -1 ; Joye et al . 1996), and in periphyton mats (1.9 μ mol N 2 O m -2 h -1 ; Triska and Oremland, 1981). Based on their results, Inglett et al. (2007) c oncluded that despite a conclusive trend in DEA downstream (along a depletion gradient of surface water NO 3 - ) denitrification appeared to be a significant pathway for N removal from the system.

Alternative N pathways

Conversion of NO 3 - can occur by additional pathways. Some of these have been identified for a number of years while others have emerged more recently (reviewed by Burgin and Hamilton 2007) (Figure 8). Many of these pathways have yet to be identified in springs or streams; however, the potential exists for these proce sses to account for significant reduction of NO 3 - in aquatic systems. For this reason, we will discuss them briefly here.

The high energetic po tential of using NO 3 - as an electron acceptor is not solely possessed by heterotrophic bacteria. Seve ral pathways have also been identified that can reduce NO 3 - through the oxidation of other reduced compounds. Identification of these new mechanisms has challenged the classical view of denitrifica tion in aquatic system s. Among these, NO 3 - reduction can be coupled to the oxidation of eith er reduced iron (Fe(II)) or sulfur (S 0 , S 2- ). The importance154 of these lithotrophic pathways to overall rates of den itrification has not ye t been determined. Likewise, in springs and lotic systems, Fe(II) and S 2- -based denitrification has received little attention. It is not unlikely that these pathways could be signifi cant given the presence of iron and sulfur in many spring systems and the myriad of reactions occurri ng at the oxic/anoxic interfaces of microbial mats , biofilms, or sediments.

Another pathway for NO 3 - reduction can convert NO 3 - into NH 4 + . This process, called dissimilatory nitrate reduction to ammonium (DNRA), is different from the assimilatory process in that NH 4 + produced is not directly incorporated in to biomass. Fermentative DNRA couples organic matter oxidation to NO 3 - reduction by fermentation reactions (Tiedje 1988).

Figure 7 – Denitrifying potential in sediments, and algal biofilms collected from various substrata. S, sediment; FB, filamentous bent hic algae; BF, Brown filamentous algae; E, epiphytic biofilms; RW, Biofilms attached to rock wall; D, detrital material; SB, algal biofilm on submerged bark. (From Inglett et al. 2007)

Another form of DNRA is chemolithautotrophic and couples NO 3 - reduction to oxidation of reduced S (e.g., S 2- ). Strongly reducing conditions with high contents of organic matter or high concentrations of reduced Fe and S are required for DNRA. As stated for chemotrophic denitrification, conditions may exist for DNRA act ivity in springs; howeve r, these areas would be highly localized to accumulations of organi c matter and/or slow or no flow conditions.

Most recently, a novel pathway for N conversion to N 2 has been identified which adds even more uncertainty to the aquatic N cycle. The process known as anaerobic ammonium oxidation (Anammox) has now been identif ied in a number of freshwat er systems including lakes and155 wetlands. This process occurs in strictly an aerobic conditions, and couples the reduction of NO 2 - with the oxidation of NH 4 + to produce N 2 gas. Annamox bacteria a ppear to be au totrophic, and the presence of organic matter slows the annamox process. Another important requirement for the anammox process is the presence of NO 2 - which is rarely observed in most systems.

Figure 8 – Hypothesized controls on predomin ant dissimilatory pathways of NO 3 - removal. C inputs refer to labile organic carbon available to microbes. Sulfidic refers to the presence of either free sulfide (S 2–

), elemental S (S 0 ). Of these S forms, only free sulfide inhibits denitrification and thus promotes DNRA. C:N ratios refer to the ratio of labile organic carbon to nitrate. Respir = respiratory; denitrif = denitr ification; DNRA = dissimilatory nitrate reduction to ammonium; ferment = fermentati ve. (From Burgin and Hamilton 2007).

Nitrite accumulation is possible however, and several conditions of spring systems may be adequate for NO 2 - accumulation to occur. The formation of NO 2 - is largely a byproduct of incomplete denitrification or pa rtial ammonium oxidation. For NO 2 - accumulation to occur in the denitrification a nd DNRA processes, NO 3 - reduction must occur at a gr eater rate than that of NO 2 - reduction. One way in which incomplete denitr ification can occur is the lack of suitable organic substrates, or hi gh concentrations of NO 3 - . Similarly, in the nitrification pathway, elevation of NO 2 - concentrations can occur only if the oxidation of NH 4 + is greater than that of NO 2 - . Both of these scenarios have been reported to occur in th e aquatic environment (Herbert 1982; Smith et al . 1997). Nitrobacter spp., which are responsible for the oxidation of NO 2 - in156 the nitrification process, may be pa rtially inhibited by free ammonia (NH 3 ) (Anthonsieu et al. 1976; Smith et al . 1997), leading to NO 2 - accumulation. Likewise, NO 2 - accumulation is a common trait of DNRA (Cole 1988 ) resulting from either inhibitory effects of NO 3 - on NO 2 - reductase (Smith, 1982) or repression of this enzyme (Paul and Beauchamp 1989).

Higher than expected NO 2 - levels (100-200 μ g N l -1 ) have been reported in some European rivers, particularly in waters heds receiving larg e inputs of nitroge nous pollution (Smith et al

. 1995; Kelso et al . 1997). In one study, Kelso et al . (1997) found that NO 2 - accumulation in six Irish rivers corresponded with NO 3 - declines via DNRA activity. This process was occurring in areas with high organic matter contents (Kelso et al . 1999), leading these au thors to conclude that high NO 3 - concentrations, free ammonia, or high fl ow velocity created conditions suitable for NO 2 - accumulation to occur. All of these studies suggest NO 2 - may be much more abundant than previously thought. Th erefore, it is likely that NO 2 - levels may be sufficient to support alternative N cycling pathways (such as anammox) in more aquatic ecosystems than previously thought.

Overall N Cycle

It is uncertain what role spring systems play as sinks, sources, or transformers of N. Several studies have reported dec lining inorganic N concentrations (primarily NO 3 - ) with distance down spring run/river (e.g., DeBrabandere et al . 2007; Inglett et al . 2007; Mytyk and Delfino 2004) (Figure 7). In general lotic systems should resu lt in a transformation of N, including conversion of DIN to DON through biotic uptake and senesc ence. Increases of DON with distance through a run have also been observed (Inglett et al ., 2007; Wetland Solutions, Inc., 2006). The study by Wetland Solutions, Inc. (2006) also indicates th at DON production in spring runs is likely the result of excess N, where Wekiva and Rock Springs Run (both NO 3 - -impacted) show much higher production of DON in specific river segmen ts, while Alexander and Juniper spring runs (reference conditions) show smaller producti on (or even net reduction) of DON.

It is also quite likely that spri ngs are active zones of denitrifica tion. Evidence of denitrification in springs comes from at least one study (Inglett et al . 2007), and the conditions exist for other unexplored areas where denitrificat ion could occur (Saunders 2007). Areas with direct inputs of reduced groundwater into sediments with low organic matter content, or water column NO 3 - being advected into anaerobic sediments, are al so key conditions for alte rnative denitrification pathways to occur in springs (Grimaldi and Chap lot 2000). In systems with significant quantities of iron or sulfur, additional li thotrophic N loss mechanisms ma y also be operating. Nitrogen cycle studies and budgets of springs are not available at this time. For this reason it is difficult to attribute the several ex amples of declining NO 3 - to either the process of biotic uptake or denitrification; however, this question is of cr itical importance for management of springs as nutrient sources.

Phosphorus

Phosphorus (P) is one of the ma jor nutrients limiting productivity in freshwater ecosystems. During the last five decades, many studies have been conducted to determine the fate of P in freshwater systems, including springs, streams and other aquatic ecosystems. At the landscape157 level, streams and wetlands form a critical in terface between uplands and adjacent water bodies, as all of these ecosystems are hydr ologically linked. Water and asso ciated nutrients (such as P) are transported from uplands either by subsurf ace or surface flow. An idealized P cycle for spring systems is shown in Figure 9.

Both biotic and abiotic mechanisms regulate re lative pool sizes and transformations of P compounds within the water column and sediment. Alterations in these fractions can occur during flow in streams and wetlands that de pend on the physical, chemical, and biological characteristics of the systems (Meyer and Li kens 1979). Thus, when evaluating P fate and transport in springs/spring runs, both biotic and abiotic processes must be considered. Biotic processes include assimilation by vegetation, biof ilm/periphyton and sediment microorganisms. Abiotic processes include sedimentation, adso rption by sediments and organic/inorganic substrates, precipitation, and exchange proce sses between sediment and the overlying water column.

Figure 9 – Idealized diagram of the phosphorus cycle o ccurring in a typical spring run system. PIP=particulate inorganic P, POP=particulat e organic P, DIP=dissolved inorganic P, DOP=dissolved organic P, Fe-P=iron bound P, Ca-P=calcium bound P.

Forms of phosphorus in water column and sediments

Phosphorus entering a spring system is typically present in both organic and inorganic forms. The relative proportion of each form depends upon the geology, soil, vegetation and land use158 characteristics of the aquifer, or springshed. To trace the transport and transformations of P, it is convenient to classify forms of P entering these sy stems as (1) dissolved inorganic P (DIP); (2) dissolved organic P (DOP); (3) particulate inorganic P (PIP), and (44) particulate organic P (POP) (Figure. 9). The particulate and soluble organic fractions may be further separated into labile and refractory components. Dissolved i norganic P is considered bioavailable, whereas organic and particulate P forms generally must undergo transformations to inorganic forms before being considered bioavailable.

Inorganic P forms in sediments are characterized ba sed on their differential solubilities in various chemical extractants. The fractionation schemes the includes following pools: (1) exchangeable P;(2) Fe and Al-P or non-apatite P; (3) Ca- a nd Mg-bound P or apatite P; and (4) residual P or nonreactive P (Hieltjes and Lijklema 1980; van Eck 1982; Psenner et al . 1988; Cooke et al

. 1992; Olila et al . 1995; Ruttenberg 1992; Reddy et al . 1998). The example in Figure 10 shows labile and non-labile pools of P in selected stream sedi ments of south Florida (Reddy et al

. 1995a). In this example, the in organic P extracted with neutra l salt such as KCl represents loosely absorbed P, which usually accounts for less than 2% of th e total P. This pool of P is considered bioavailable as it is sufficiently labile to enter por ewater for uptake of plants or potential flux into the overlying water column.

The NaOH-P i represents inorganic P associated with Fe and Al, and represents P not readily available. The HCl-P i represents P associated with Ca/Mg, is also relatively stable and not readily available. Alkali extractable orga nic P includes both readil y available organic P (microbial biomass P) and slowly available organic P (P associated with fulvic and humic acids). Residual P represents highly resistant organic P or unavailable mineral bound P not extracted either with alkali or acid. The Ca-bound P such as apatite has been found to be unavailable (Pettersson 1986; Gunatilaka 1988), while the re dox-sensitive Fe-bound P may become available under anaerobic conditions (Figure 9) (Wildung et al . 1977; Furumai and Ohgaki 1982; Hosomi et al . 1981).

Although these categories allow some evaluation of properties, mobility, and quantitation, the extent of bioavailability among each of the organic P fractions can be obscure and variable. Bioassays with microbes (bacteria, algae) can yi eld insights on instantaneous bioavailability, but clearly sorption and desorption reactions can shift the solubilities and av ailabilities rapidly. These exchange reactions are markedly influe nced by microbially-indu ced redox shifts and hence can change rapidly in water and especially in sediments.

Organic P forms may dominate the water column, with the majority of P occurring as particulates in biomass of algae and macrophytes. Some of the organic P may be associated with dissolved organic phases. The relative proporti on of organic P in sedime nts of springs, spring runs and streams may be smaller as compared to riparian zones. In many streams accumulation of organic matter may be lower compared with riparian zones, thus resulting in lower accumulation of organic P. In organic matter, P is a major constituent, as nucleic acids (DNA and RNA), in phospholipids of membranes, a nd as monoesters of a variety of compounds, particularly those involved in bi ochemical pathways. In microor ganisms and algae, more than half of the organic P is in nucleic acids, wh ereas phospholipids and monoe sters constitute the remainder in varying proportions (Magid et al . 1996). Many organisms can store orthophosphate159 or polyphosphates (Cembella et al . 1983, 1984). In plants, inos itol hexaphosphate (IHP) can form a major storage compound for P, particularly in seeds.

The epiphytic community living on submerged surfaces acquires P from both the water column and from the supporting ‘host’ substrate. Dissolu tion of inorganic substrat es such as minerals can potentially provide some essential nutrient s to algae and microbes colonizing on them. Although relatively little of the to tal P pool within actively growi ng macrophytes is released, this released P can be important to certain epiphy tic species that grow adnate to the macrophyte tissue (Moeller et al .1988). Even when P concentration in the overlying water is high, some nutrients are obtained from the macrophyte simply because diffusion within the complex epiphytic community is too slow to meet dema nds (Wetzel 1993a). The periphyton, rather than the macrophytes, functions as the primary scavenger for limiting nutrients such as P from the water column.

Figure 10 – Fractionation of sediment phosphorus in South Florida streams dominated by calcium (DL Stream) and iron (Rucks Stream). (From Reddy and DeLaune 2008).

Mineralization of organic phosphorus

Active cycling of organic P in sediments and wa ter column is largely mediated by microbial metabolism. In order to understand cycling of organic P it is essential to evaluate the organically-bound P within compounds of the mi crobial biomass and the organismal dynamics, particularly the bacteria, as they influence or ganic P dynamics. Microbial metabolism influences rates of mineralization of organic P in two prim ary ways: (1) by direct biochemical mobilization using extracellular or periplas mic hydrolases, particularly from inositol phosphates, and (2)160 through alteration of ionic composition, redox, pH, and other conditions that in turn can alter the efficacy of P binding to particulate organic matte r, clays, and other soil particles, as well as solubility of organic P and its potential enzymatic hydrolysis (Reddy et al . 2005).

The availability of phosphate monoesters requ ires enzymatic cleavage of the ester linkage joining the inorganic P group to the organic moiety. Such hydrolysis is achieved by phosphomonoesterases (phosphatase s) bound to or within the cell membrane or by dissolved enzymes released extracellularly within the interstit ial solution adjacent to the cells. There is a general lack of specificity fo r the organic moiety that incl ude, beyond phosphate monoesters, substrates such as diesters and phosphoanhydrides such as pyrophosphates, ADP, and ATP. The phosphatases of bacteria, algae, fungi, and higher plants are constitutive i norganic P-irrepressible enzymes, which tend to be intracellular and function in intermediary metabolism, or are phosphate ester-induced inorgani c P-repressible enzymes, which are usually membrane-bound and function in extracellular P cleavage (Cembella et al . 1983).

Bacteria and microalgae growing on soil particles and within interstitial water of sediments of aquatic ecosystems are able to utilize exog enous organic P compounds through enzymatic hydrolysis of terminal phosphate groups. A br oad spectrum of dissolved organic P compounds can sustain growth of these organisms, including glycerophosphate, D-glucose 6-phosphate (G-6-P), adenosine 5’-monophosphate (AMP), cytidine 5’-monophosphate (CMP), guanosine 5’-monophosphate (GP), adenylic acids, and phosphonate compounds (Cembella et al . 1983, 1984). Because of the high reactivity of solubili zed P, the distance between sites of enzymatic hydrolysis of organic P and sites of uptake must be short, often in immediate juxtaposition (Tarafdar and Jungk 1987). Inactivation by reaction with the soil matrix increases with time after release and with greater available surface areas associated with decreasing size of soil aggregates (Linquist et al . 1997a).

Expectedly, a positive correlati on has been found between bact erial biomass and phosphatase activity in surficial lake sediments (Reichardt 1978). Correl ations between enzy me activities and bacterial biomass were less robust several centi meters below the sediment-water interface. A higher affinity of phosphatases was found in se diments where a low mobility of sediment P occurred (Pettersson unpub lished; cited in Boström et al . 1982). The lowest substrate affinity (highest half-saturation constant s) was found in organic sediments with low chemical adsorption capacities that release P under anae robic conditions. Low availabil ity of P in sediments often induces production of phospha tase activity with high affinity for substrates.

Biotic and Abiotic Ca Interactions

In calcium dominated streams, the presence of high Ca 2+ can result in formation of complex calcium phosphate compounds of varying solubi lities such as calcium phosphate, dicalcium phosphate, beta-tricalcium phosphate, octacalciu m phosphate, and hydroxyapa tite. Under these conditions, the phosphorus concentra tion in the pore water of calc areous sediments is a function of Ca 2+ activity. Solubility of these compounds decr eases with an increase in Ca content. Insoluble beta-tricalcium phosphate is more lik ely to found at a high pH. Thermodynamically, apatite is the most stable compound, while at re latively high phosphate concentrations, dicalcium phosphate or octacalcium phosphate may form and slowly transform to the more stable phase,161 hydroxyapatite. These precipitation reactions can occur on surfaces of calcite. The amount of exposed surface will determine the amount of phosphor us precipitated. In Ca-saturated clay, calcium phosphate will precipitate as a separate phase above pH 6.5; however, this depends on the concentration of phosphate ions in sediment pore water.

Algal communities can play a majo r role in regulating P concentrations of the water column. Algae can assimilate both organic an d inorganic forms of P (Bentzen et al . 1992), and can induce marked changes in pH and dissolved oxygen concen tration of water column and soil-floodwater interface (Carlton and Wetzel 1988). On a diurnal basis these processes can increase pH to as high as 10, depending upon the buffering capacity of the water column. These changes can potentially influence the solubility of P, esp ecially in streams. Some studies (e.g., Hartley et al

. 1997) have demonstrated that this process can result in significant co-p recipitation of phosphorus with calcite in the presence of photosynthesizing algae, like Chlorococcum sp

.
Retention of inorganic p hosphorus by precipitation will be signi ficant in spring wa ters with high Ca 2+ and alkalinity. In cal cium-dominated streams (e.g., chalk streams) a significant portion of water column phosphorus can co-precipitate with CaCO 3 where dissolved Ca 2+ is available. House (1990) attributed only 6% of the overall p hosphorus removal to co-precipitation, while the remainder was due to biological uptake. However, Salinger et al . (1993) found metastable calcium phosphate species to be a predominant form of phosphor us transported in the River Jordan. Several studies showed inhibition of calcite growth in the presence of high phosphorus concentrations (Avnimelech 1980; Dove a nd Hochella 1993) while at low phosphorus concentrations, coprecipitati on of phosphorus with calcite has been observed (House and Donaldson 1986; Giannimaras and Koutsoukos 1987; Kleiner 1988). Approximately 75-90% of the phosphorus precipitated was solubilized when pH levels decreased to below 8 as a result of an increase in carbon dioxide levels (Diaz et al . 1994).

Biotic and Abiotic Fe Interactions

In spring and stream sediments dominated by iron minerals, reduction of ferric (Fe(III)) hydroxides and complexes results in ferrous Fe (F e(II)), and adsorbed phosphate to be mobilized into soil pore water (Figure 9). Such conditi ons are of particular importance in areas where organic matter (and reducing conditions) can be transient, or when ground water varies temporally between aerobic an d anaerobic states. Under str ongly reducing conditions where sulfate is reduced to hy drogen sulfide, some S 2- can react with iron co mpounds (both Fe(III) and Fe(II)) resulting in the precipitation of Fe-S co mpounds. In this manner, sulfide formation coupled to sulfate reduction can displace iron-bound P to enhan ce the release of P into the interstitial waters (Wetzel 2001) . The relative P release from sediments can be significantly higher as sulfate concentrations increase from natural or anthr opogenic sources, particularly in oligotrophic, softwater lakes and rivers. In highly reduced soils with low sulfate content, microbial reduction generates mostly Fe(II) co mpounds from microbial reduction of Fe(III) oxide. PO 4 -3 can be retained with Fe(II) compounds, but Fe-associated PO 4 -3 is released when amorphous Fe(III) oxide is converted to iron su lfides (Roden and Edmonds 1997). Thus, the conversion of reactive soil Fe compounds to iron sulf ides by sulfate-reducing bacteria can lead to a more efficient release of Fe-associated PO 4 -3 than does direct microbial Fe(III) oxide reduction.162 Overall P Cycle

Springs with high N availability should respond with higher rates of P uptake and subsequent release of DOP following senescence of biomass. Many springs do show such an increase of DOP with distance from th e boil (e.g., Wekiva; Inglett et al . 2007). In these cases, springs should act as transformers of P or act as a small to moderate sink. In other cases, springs may actually serve as sources of P. This could occur through accumulation and mineralization of riparian vegetation deposited in the spring run, or solubilizati on of P contained in sediments emerging at the boil.

The process of P exchanges across the sediment -water interface is regulated by mechanisms associated with mineral-water equilibria, sorption processes (p articularly ion exchange), oxygen and other electron acceptor-dependent redox inte ractions, and the physiol ogical and behavioral activities of many biota from bacteria, algae to other living communities. Exchange rates between various deposits of P and the inters titial water of the sediments depend on local adsorption and diffusion coefficients and their alteration by enzyme-media ted reactions of the microbiota. The sink/source func tion of springs could be transi ent, where there are periods of storage/transformation followed by periods of flux/release. For this reason, whether individual springs act as sinks, sources, or transformers of P is highly dependent on the spring, and its physical, chemical, and biological characteristics.

Sulfur

The idealized sulfur cycle for spring systems is shown in Figure 11. The predominant forms of sulfur are organic-S (existing mo stly as protein/AA-S), sulfide (S 2- ), and sulfate (SO 4 2- ).

Sulfide Oxidation

Emerging groundwater potentially contains quantities of S 2- that become oxidized when contacting atmospheric and photosynthetically-produced O 2 near the boil/vent. This reaction is thermodynamically favorable, and several groups of bacteria (e.g., Thiobacillus spp.) can utilize this energy yield to fix atmospheric C (lithoaut otrophy), or respire or ganic C (heterotrophy). Data from several springs indicate this process ma y be occurring. Sulfide is rarely determined in routine monitoring, but some springs show post-boil increases in SO 4 2- (Wekiva, e.g.; Wetlands Solutions, Inc. (2005 )) indicating that S 2- oxidation could be occurring.

Springs with high SO 4 2- content may also host this process, but it likely oc curs pre-boil (e.g., Juniper Springs). Other evidence for sulfur oxidizing pathways in spring systems comes from a study by Caldwell et al . (1975) who found large numbers of S-oxidizing bacteria including the
genera Thiothrix and Thiocystis in a sulfur spring in the Florida Keys (Flowing Well) and several springs associated with lakes in Michigan. These bacteria were found attached as filaments or mucilage-encased microcolonies on aerobic su rfaces of plants, interspersed with other filamentous algae, or on decomposing organic matte r. Gradient culture techniques revealed that these S-oxidizing genera were culturable even at the lowest S levels tested (55 μ mol S l -1 ).163 Sulfate Reduction

Sulfate present in the water column of spring runs can be utilized as an electron acceptor by select bacteria decomposing simple organic substr ates (dissimilatory sulfate reduction). This process where SO 4 2- becomes reduced to S 2- only occurs under highly reducing conditions with redox potentials less than 200 mV. For this reason, measurable sulf ate reduction is likely only possible in slow-moving spring runs or areas w ith high organic matter sediments (algal mats, SAV beds, litter accumulations).

Sulfate reducers are obligate anaerobes that co uple oxidation of organic substrates to carbon dioxide with the reduction of the terminal elec tron acceptor sulfate to sulfides (Widdel 1998). Sulfate reducing bacteria cannot synthesize hydrolytic enzyme s, thus they cannot hydrolyze polymers such as polysaccharides, and many groups cannot use monomers such as monosaccharides (e.g., glucose) as substrates for energy. As a result, sulfate reducers are dependent on fermenting bacteria to oxidize mono mers to simple readily utilizable organic compounds.

Figure 11 – Idealized diagram of the sulfur cycle o ccurring in a typical spring run system.164 Overall S Cycle

Because of the prevalence of oxidizing conditions in the water column, most spring runs should result in a net export of oxidized S species (S 2- oxidized post boil). W ith the exception of the initial degassing at the boil of some high S springs, most Florida springs should not emit significant amounts of S as H 2 S. The only other significant quantity of S compounds should be through the transport (loss) of organic- S in organic matter carried downstream.

EFFECTS OF INCREASED NUTRIENTS

Carbon/Organic Matter Impacts

Common ecological theory would pr edict that addition of nutrients to a nutrient-limited system will enhance productivity and lead to greater ove rall cycling of C through the organic pool. However, a study of four spring runs by Wetland Solutions, Inc. (2005) found that spring runs with higher nutrient levels (Wekiva and Rock Springs) had lower ecosystem productivity estimated through total system metabolism (DO gain/loss) than the reference spring runs (Alexander and Juniper). Increased nutrients leading to algal dom inance and growth (even if not resulting in an overall increase d productivity) would still reflec t a change in the quality of organic matter in the system and could have cascading effects thro ugh other heterotrophic communities both locally and further downstream.

Other aspects of the C cycle in springs could be altered through th e addition of C. Increased C inputs could occur through increased sewage/pollu tant discharges which would act to raise the concentration of solids C or DOC . Management of exotics/inva sives could also have profound impacts on nutrient dynamics through increased a nd episodic deposition of C following herbicide application. The effects of larg e discharges of C into streams are well known and most likely the same effects apply to springs, spring runs and sp ringsheds. In partic ular, the introduction of large C quantities increase respiration losses of oxygen and other electron acceptors and increase production of significant quantities of reduced compounds potentially affecting many of the nutrient cycles, including N (increased denitrific ation), or P (increased release of iron-bound P).

Nitrogen Impacts

In the natural state, most Florida springs cont ain high levels of inorganic P, and based on N:P ratios, are N-limited. In the N-limited state, spri ngs would be considered highly susceptible to N pollution. One of the key response s to N in these cases would be increased biomass or shifts in dominant species of all groups, including algae, macr ophytes, fungi and bacteria. One example of this would be that under N-limiting condition s there should be a dom inance of particular Nfixing microbiota. With the shifts in speci es composition, changes in C abundance and quality can also be expected. For example, increased algal production would re sult in lower amounts of structural plant C compounds (cellulose, lignin) and an overall, more available C supply for aquatic heterotrophs.

Nitrate is by far the most commonly reported N pollutant, and has incr eased dramatically, in Florida’s major springs during the last 30 years (Scott et al , 2004). With this trend, many springs165 are increasingly becoming P limited (Figure 12) (Inglett et al . 2007; Frazer et al . 2001). Aside from effects of incomplete denitrification (producing NO 2 - and potentially N 2 O), it is currently uncertain if high levels of NO 3 - can have any other adverse impact s to biogeochemical processes.

Phosphorus Impacts

In general, many springs already have high amounts of bioavaila ble P with SRP levels >80 ppb. High levels of bioavailable P lead to the general observation of low water column N:P ratios in most springs (i.e., those unaffected by NO 3 - pollution) (Figure 12). As in other aquatic systems, elevated P levels should result in higher overall producti vity, or if limited by some other nutrient (or light) a higher potentia l productivity level. Fo r this reason, P availability may determine the overall threshold response of a given spring system to another nutrient, such as N. This response may be visible only through increa sed productivity of certain sp ring species; howev er, in this manner, high-P springs may demonstrate a great er response (either through greater overall production or that of individual species) or be more susceptible to increased N.

Figure 12 – Ratios of dissolved inorganic N:P and total N:P (mass basis) in the Wekiva river in relation to theoretical boundary between conditions of N limitation and conditions co-limited by N and P. (From Inglett et al. 2007)

Sediments in spring systems have the capability to store P as Ca-P or Fe-P complexes, and thus could be a reservoir of P that becomes mobilized upon increased leve ls of N. Fluctuations in water column P levels will determine if the sediments act as a sink or source, as low water166 column P levels will favor flux of sedimentary P. The response of some systems to additions of P has been observed as a decreas e in macrophytes chlorophyll leve ls. This may reflect changes in the microbial composition or the stoichiometric balance of N:P and C:P which could shift the dominance of autotrophs to more heterotrophic communities . Increased heterotrophy would have cascading type effects on th e balance of aerobic/anaerobic c onditions, and thus, would also effect the cycling of redo x-sensitive elements such as N, P, Fe, and S.

Sulfur Impacts

It is uncertain what role S may be playing in th e biogeochemistry of Florid a springs. Sulfur can be significant as an electron acceptor (as SO 4 2- ) with a toxic byproduct of S 2- . Also, S 2- can become a powerful electron donor for particular organism groups and can diversify the pathways for NO 3 - removal through denitrificat ion. Furthermore, the impacts of sulfur may not be immediately visible from data collected at the boil/vent. The oxidation pathways and processes may develop slowly following gr oundwater emergence and may be temporally variable. These effects may also be spatially localized to zones of significant S input or transformation such as the mixing of spring water with co ntrasting water high in S (e.g., th at observed when tributaries meet).

SUMMARY AND CONCLUSIONS

Based on the above discussion, springs are clearl y dynamic systems with a myriad of potential biogeochemical processes/reacti ons and interactions among el emental cycles. Although the majority of these basic biogeochemical processes ha ve been identified in other aquatic systems, many may not be universally applied to spring syst ems because of their unique features including karst-derived water chemistry, and relatively c onstant temperature a nd flow conditions. A diversity of conditions within spring systems al so gives rise to numerous microbially active zones where conditions may exist for as ye t unexplored processes and pathways.

Nutrient cycles in spring systems are dominated by the conversion of inorganic to organic forms. Organic matter (dissolved and pa rticulate forms) is therefore an important component of the spring system as it is both an internal supply of nutrients and a form of nutrient export. Carbon is also the basis of many other anaerobic biogeoc hemical reactions, and is thus an important determinant in the cyc ling of other nutrients including nitr ogen, phosphorus, and metals. Despite this importance, organic matter distributions, c oncentrations, and quality types within spring systems are relatively unstudied, but may have an importance for how individual springs process and store nutrients. Also, aquatic plant manage ment practices can be a contribution of organic matter deposition in spring runs (a ffecting internal nutrient cyclin g); however, the importance of the process is undetermined.

Nitrogen has been a nutrient of great interest in the Florida springs due to its trend of increasing concentrations in the springs which are associated to changes in spring appearance, especially increases in algae. In additiona l to external loads of N through NO 3 - , nitrogen fixation is an important processes contributing to N, but is also currently unst udied in Florida springs. There are several pathways which can act to convert NO 3 - (reducing its concentration), with the dominant most likely being heterotrophic denitrif ication or plant assimilation. The importance167 of these two pathways will largely determine if a spring is a sink or source of N; however, current studies have not attempted to partiti on N flow through each pathway. High flow rates and patchy distributions of organic matter also indicate that conditions in many springs could support several alte rnative (novel) pathways of NO 3 - reduction, but again, these have received no attention.

Phosphorus may also play a key role in regul ating species composition or productivity in a spring system. A variety of biot ic and abiotic processes act to determine the biogeochemical cycling of phosphorus in spring systems. The ge nerally high calcium cont ent of many spring waters leads to the potential for accumulation of Ca -P compounds associated with biotic growth and mineral forms in sediments. Organic matter and associated enzymatic hydrolysis of organic P compounds is also another key mechanism regul ating both accumulation and bioavailability of phosphorus in springs. Despite the significance of these processes; however, the general lack of studies focusing on sediment properties currently limits our understanding of P cycling, or its importance as an internal nutrient supply mechanism.

Several other elements could have importance for nutrient cycling in spring systems including sulfur and iron. Iron participates in redox reactions that determ ine both its form and solubility/mobility and its pa rticular interaction with PO 4 3- (affecting P availability). Iron is also significant as a potential limit to algal growth in certain highly alkaline systems. Like iron, sulfur can undergo dynamic transformations through redox processe s in alte rnating aerobic/anaerobic systems like springs. Presence of sulfide is likely only a problem (toxicity) in certain sulfur spring systems; however, there is a possibility that sulfide-driven denitrification may be a significant process reducing NO 3 - levels in some aquifers and in some spring runs.

Management and Policy Implications

Nutrient enrichment of Florid a springs is major concern. The steady increase in nutrient concentrations observed in many Florida spring s during the past four decades threatens the economic and ecological value of these springs . Nonpoint sources of nutrients dominate eutrophication processes of many Florida springs. Thus, most recommendations involve the use of alternative land management practices in the wa tershed in an effort to reduce the overall load to these springs. The key management questions often asked are: (1) Will springs respond to external nutrient load reduction?, and (2) If so, how long will it take for these systems to recover and reach their background condition?

The question of system impact is important fo r management of spring systems. Nutrient limitation and biogeochemical proce sses are a very inter active to determine the biotic response of a given spring to external nutrient loads. The chemistry of spring source water is crucial to determine the types of metabolic pathways and diversity of organisms that will populate/dominate a spring ecosystem. In pa rticular, microbial co mmunities will likely demonstrate responses to changes in nutrient co nditions immediately through changes in species or functional groups and through activities of en zymes which are produced in response to their external chemical environment. In this way, monitoring of biogeochemical processes may serve as an early warning indicator of spring health.168 Much of the nutrients added to springs can be retained within the system through the accumulation of large reserves in detrital tissue and/or reactive forms in sediments. These nutrient stores can serve as a nutrient source for l ong period of time, even after external loads are reduced. As discussed in this paper, various biogeochemical processes in sediment and water column regulate the mobility and reactivity of stored nutrients. The cycling of nutrients can extend the time required for springs to reach thei r original condition, with a lag time from a few years to several decades, depending on the extent of internal nutrient supply. This lag time for recovery should be considered in developing management strate gies and monitoring efforts, where environmental managers must recognize that the recovery process is not immediate, even despite successful reductions in external loads.

To determine the role of internal nutrient supp ly from sediments, it is important to document storages and fluxes of nutrients between sedime nt, vegetation, and water column. At this time little or no information is available on internal cycling of nutrients a nd its overall role in regulating eutrophication process. When considering internal cyc ling of nutrients in springs, several issues need be addressed. These issues ar e: (1) what is the relative bioavailability of various nutrient forms stored in sediments and water column? (2) how stable are the stored nutrients and under what conditions could they be released back into the water column? and (3) how long could the stored nutrients serve as a source to the wate r column to maintain eutrophic conditions, once external loads ar e reduced? Thus, monitoring in ternal nutrient storages on a regular basis is equally as important as monitoring external nutrient loads. Such information will be useful to the state’s Department of Environmental Protection as it attempts to establish nutrient criteria and Total Maximum Daily Loads (TMDLs) for springs.

RESEARCH NEEDS

Biogeochemistry is one of the least studied as pects of spring systems. The complexity of biogeochemical pathways, the complexity of th e high energy/lotic spring environment, and the diversity of spring water types ma kes it difficult to form any wi de-ranging genera lizations. For this reason, biogeochemistry is also one of th e aspects of spring systems requiring additional effort to increase our understanding of spring nutrient impacts. Based on the review of literature presented, the following are our recommendations for research related to the effects and consequences of nutrients in calcareous spring systems.

Specific Research Recommendations

Biogeochemical Characterization

There is general lack of data on characterizati on of spring systems for biogeochemical purposes. Characterization data and long term measurements of water chemistry ar e available; however, the addition of some parameters to this monitoring are warranted, including sulfide (S 2 ) and reduced iron (Fe(II)). Also, ther e is a large uncertainty in the function of sediments for storage and cycling of nutrients. Identification of biogeochemically active zones will depend on knowledge of sediment mineralogy, organic ma tter contents and qualities, and stratigraphic morphology. For this reason, data should also be collected to characterize the biogeochemical properties of spring run sedime nts including basic nutrient chem istry, mineralogy, and potential169 storages of nutrients in the sediments of most springs/runs. Better mapping of spring run sediment characteristics can also help improve modeling of nutrient dynamics within a given spring run (including sediment functi on as sink or source of nutrients).

Source and Fate of N

Research is needed to understand the fate of N within the spring system. This would include understanding of biotic uptake processes (algal/macrophyte uptake) , importance of different N sources (NH 4 + , NO 3 - , DON), and the occurrence, rates and importance of biological N 2 -fixation. Whole spring run N budgets should be performed to better isolat e the processes responsible for N decline, and these data put in the context of nutrient limita tion of systems of differing geology, channel morphology, water chemistry, and biotic productivity. Give n the potential for structuring aquatic microbial communities, the presence and effects of potential DON sources (including urea) should be investigated in rela tion to algal composition, microbial activity, and community nutrient limitation. Potential pathways of NO 3 - loss should also be investigated, including identification of hotspots of denitrification as well as the potential occurrence and conditions regulating any novel N loss pathways (e.g., lithotrophic Fe(II)/S 2- oxidation coupled to denitrification, or anammox).

A variety of techniques can be employed to help answer some of these questions regarding N cycling and fate in spring systems. Determinat ion of system N uptake can be done in small batch or whole stream segments (Sa unders 2007; Wetland Solutions, Inc. 2005). Relative estimates of enzyme activities are simple to conduct and provide information regarding spatial or temporal distribution of processes (Inglett et al . 2007). Other enzyme assays can be used to provide quantitative estimates of N processes such as the determination of denitrification using the traditional acetylene inhibition assay (DEA technique; Inglett et al. 2007) or N 2 fixation assays (Inglett et al. 2004). It is well known that natural a bundance level stable N isotopic methods can be used to assess sources of N in sp ring systems; however, the analysis of δ 15 N can also provide information on rates of N uptake or denitrification (De Brabandere et al . 2007). Experiments utilizing tracer levels of 15 N can also provide useful identi fication of N pathways (e.g., novel NO 3 - reduction pathways) as well as quantitative information of pathways and flows (e.g., N 2 fixation, N uptake, and denitrification).

Carbon Dynamics

Carbon sources and distribution pl ay a key role in structuring th e biogeochemistry of the spring system, therefore better understa nding is needed regarding the sources and processing of C in spring systems. This would include identific ation and quantification of autochthonous and allochthonous C sources (quantitie s and qualities), availability a nd cycling of different C types, and investigation of the importance of differe nt C sources for respiration processes and denitrification. The role of aquatic plant manage ment in the distribution and fate of C within particular systems is also of si gnificant concern as it constitutes a significant source of C and nutrients within the system.170 Sulfur Distribution and Function

Little attention is given to sulfur contents of ground waters, except wher e smell or taste issues are concerned; however, S may be a key elem ent regulating several biogeochemical cycles including N and P. For this reason, more em phasis should be given to S as it relates to biogeochemical processes includin g its distribution and redox state in spring environments (i.e., locations of oxidized and reduced S accumulations). One of the more important aspects of the uncertain role of sulfur is its potential involvement in regul ating N and P biogeochemistry, therefore, studies linking redox processes with denitrification or P flux from sediments are potentially useful for elucidati ng the functioning of S in some nut rient-impacted spring systems.

Coupled Biogeochemical Cycles

Identifying and understanding biogeochemical path ways is of critical importance; however, as should be evident from this review, biogeochemical processes do not act independently in natural systems. From the basic concept of organic ma tter and biotic productivit y, it can be seen that elements (and thus elemental cycles) are tied toge ther (or coupled) so that the pathways of one cycle inherently affect the pa thways of another element.

Currently, there have been no attempts to understa nd the coupled nature of cycles (N and P; S, N, and P; S and trace metals) in spring systems. Not only could this have important implications for understanding both the localized effects of nutrients on the biot a (e.g., questions of N versus P limitation), but it could also have application to modeling th e function of these systems as sinks, sources or transformers of nutrients and C at the landscape scale. For this reason, there should be increased focus on asse ssing the potential fo r linkages between biogeochemical cycles (C, N, P, and S) as they relate to landscape patterns of geomorphology, species composition/productivity, land use, and nutrient impact . In addition, data se ts (both current and future) should incorporate the concepts of coupl ed biogeochemical cycles when collecting data to assess spatial and temporal scales.171 REFERENCES

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# 4

Nutrient Effects on Spring Flora and Fauna

Charles A. Jacoby, Thomas K. Frazer and Edward J. Phlips Department of Fisheries and Aquatic Sciences University of Florida

Summary Many of Florida’s springs and spring runs ar e enriched in nitrate due to broad-scale contamination of groundwater supplies. This ob servation fosters two primary and interrelated concerns regarding the effects of nutrients on flora and fauna in spring systems. High nitrate concentrations can affect fauna directly through t oxicity. In addition, high nitrate concentrations can promote eutrophication or an increase in orga nic matter, which drives a diverse array of responses and feedbacks in aqua tic systems that comprise a eutrophication progression scheme.

The Literature: What is Known Florida’s springs and spring runs are enriched in nitrate due to broad-scale contamination of groundwater supplies. As a result, the potential for eutrophication becomes a major concern. A eutrophication progression scheme suggests that increased nutrient deliver y to aquatic systems favors production of fast-growing algae that ulti mately outcompete and di splace native vascular plants (Duarte 1995). In some of Florida’s spring systems, both the apparent proliferation of nuisance algae and the apparent decline of native vascular plants are c onsistent with such a progression and a cue that mana gement actions are needed.

In general, nutrients can affect the faunal assemblages in aquatic systems in three primary ways: 1) toxicity, 2) changes to trophic webs and 3) ch anges in habitats. Elevated concentrations of ammonia and nitrate can lead to increased mortality and sublethal effects. At this time, lethal effects are not an overwhelming concern in most spring system s. However, the potential for sublethal effects should be cons idered during discussions of ma nagement actions, especially given the uncertainty factors suggested for most criteria. The potential for changes in trophic webs and habitats, along with flow-on effects on faunal assemblages does exist. The form andmagnitude of any changes are determined by complex interactions among bottom-up and topdown processes, such as grazing pressure, habita t use and predation pressure. The outcomes of such interactions clearly relate to su stainable management of spring systems.

The Literature: What is Not Known There are few quantitative data that definiti vely link changes comprising the eutrophication progression scheme to increased nitrate loads in spring systems. In fact, the eutrophication progression scheme predicts that nonlinear responses may have “decoupled” these systems from180 contemporary concentrations of nutrients in the water column. Currently, patterns in the distribution and abundance of plan ts and algae do correlate with other physical and chemical parameters, such as light availability. However, little is known about the individual or combined effects of the various forcing factors or the form and strength of potentia l feedback mechanisms, including those linked to nutrient concentrations.

At this time, there are no data to support rigorous evaluation of bottom-up and top-down influences on faunal communities. For example, changes in faunal assemblages have not been linked to changes in vegetated communities and changes in grazer abundance have not been linked to changes in flor al assemblages. In addition, sublethal effects of nitrate may not translate to effects on populations, assemblages or ecosystems.

Recommendations for Future Research Management of nutrients in spring ecosystems w ould benefit from compilation of inventories, collection of baseline data, and completion of diagnostic studies focused on bottom-up and topdown interactions. Ultimately, unf oreseeable, emergent properties that arise when management is implemented at real-world, spatiotemporal s cales will create a need to manage adaptively.

Efficient and effective research to improve ma nagement of Florida’s springs and spring-fed systems requires input from multiple stakeholde rs. Initial suggestions for discussion include:

• inventories and baselines designed to:

− measure nutrients, oxygen, carbon, light, flow , substrates, grazing, algae, plants, invertebrates, fish, and other key driver s and valued components of spring systems chosen with the best availa ble information and, eventually, with guidance from the results of diagnostic studies

− target systems where effectiveness of management can be assessed (e.g., systems with total maximum daily loads, minimum flows and levels, pollution load reduction goals, surface water improvement plans or basin management action plans)

− sample synoptically or at spatial and tem poral scales that su pport rigorous analyses of interactions (note that these inventor ies and baselines form the platform for adaptive management)

− balance breadth with depth (e.g., opti mize taxonomic detail and replication)

• diagnostic studies that determine assimilation of nutrients by flora as key elements in nutrient budgets for spring systems, includi ng contributions by epiphytes, macroalgae, vascular plants and microphytobentho s or microalgae found in sediments

• diagnostic studies that determine how nutrien ts and other drivers affect overgrowth, shading and other relationships among periphyt on, macroalgae and vascular plants in an effort to determine the need for management actions, including aquatic plant management and restoration of native species

• diagnostic studies designed to:

− focus on species of algae and plants that are hypothesized or known to play major roles in nutrient assimilation181 − extend beyond nitrogen and phosphorus to in clude other potentially limiting factors and interactions among limiting factors that can significantly influence the outcomes of management (e.g., micronutrients, iron, oxygen, carbon, light, flow and substrate type)

− explore relationships and consequences across multiple levels of biological and ecological complexity, includi ng uptake and other physiolo gical responses, growth and reproduction of individual organisms, and ecological changes in populations or assemblages to promote the success of management actions at the system level (note that some of the key, large-scale relatio nships form the core of an adaptive management approach)

− elucidate the direction, magnitude and form of limitations (e.g., nonlinear and nonreversible) to identify suitable targ ets and goals for management actions

• diagnostic studies to identify and elucidate changes in comp osition and function of faunal assemblages related to changes in habitats in an effort to identify habita ts to be protected or restored by management actions

• diagnostic studies to determine palatability of flora to grazers, rates of grazing, and the form and magnitude of predation by primary c onsumers in an effort to identify flora, grazers and trophic links to be protected or restored by management actions

• diagnostic studies to identify and elucidate th e toxic effects of ammonia and nitrate beyond physiological and individual levels of or ganization (e.g., population and assemblage effects) in an effort to determine thresh old levels, safety fact ors and the need for management actions182 INTRODUCTION

Anthropogenic activities have increased the delivery of nutrients in aquatic systems of all types resulting in concerns about the bi otic integrity of these systems. In fact, excessive growth of primary producers in water bodies due to high inputs of nutrients, es pecially phosphorus and nitrogen, has been cited as the most important form of pollution in lentic and lotic systems (Mainstone and Parr 2002; Hilton et al . 2006). For example, high levels of nutrients, especially nitrogen and phosphorus, are report ed to threaten wadeable stre ams, the shallow, freshwater systems that represent about 90% of the length of flowing waters in the United States (Figure 1; USEPA 2006). Florida’s spring-fed systems share ke y features with these and other lotic systems because they are often relatively shallow and they rely on primary production that arises from periphyton, algae and vascular plants rather than phytoplankton. Thus, an understanding of the effects of nutrients on shallow, freshwater ecosy stems represents a key concern in Florida and beyond.

Figure 1 – Spatial extent of stressors and associated relative risks or relative likelihood of degradation from biological reference conditions in wadeable streams of the United States (USEPA 2006).

Signs of stress from increased nutrients in aquatic systems have led to development of narrative standards or numeric criteria as important manage ment responses. Nutrient standards and criteria have been set in respon se to: 1) adverse eff ects on humans and domestic animals, 2) aesthetic impairment, 3) interference with human use, 4) negative impacts on aquatic ecology, and 5) excessive nutrient transpor t to downstream systems (Dodds and Welch 2000). These five concerns interact. For example, systems exhib iting negative ecological imp acts probably fail to be aesthetically pleasing or support a full rang e of human uses. In fact , aesthetics and direct human use typically relate to the health of aquatic flora and fauna, e.g., a balance between macrophytes and filamentous algae, the absence of algal blooms, and the presence of healthy fisheries. Flora and fauna also influence the transport of nutrients to downstream systems through processes such as uptake and cycling. Thus , an understanding of nutrient effects on flora and subsequent effects on fauna become important in managing nutrients for almost any purpose.

In Florida, past and present land uses have resu lted in broad-scale nutrient contamination of groundwater (Cohen et al . Chapter 2). This nutrient-laden groundwater exists within a very183 permeable, karst setting, with myriad pathways between aquifers and surf ace waters (Knight and Notestein Chapter 1; Cohen et al . Chapter 2). In fact, freshwater springs function as conduits that discharge contaminated groundwater into surface waters. Once nutri ents, such as nitrogen and phosphorus, are mixed into spring-fed, surface wate rs, they attain the greatest potential to negatively alter the ecology of aquatic ecosystems. The effect s of increased nitrogen in the form of nitrate raise particular concerns for flora and fauna associat ed with Florida’s springs and their downstream receiving waters.

Nitrate as a Compound of Concern

Why the focus on nitrate? Nitrate is a nitrogen compound that occurs naturally in aquatic systems, including Florida’s springs. Aquatic plants and algae assimilate nitrate as a way to gain nitrogen, which is one of the elements that co ntributes to their growth and production. Concerns arise when nitrate concentrations become toxic or an increase in the supply of nitrate or another limiting factor stimulates excess plant and alga l growth creating unde sirable aesthetic and ecological consequences. Such an increase in plan t or algal growth is commonly referred to as eutrophication (c.f. Nixon 1995). Th e potential eutrophication of Fl orida’s spring systems is a major concern for many resource managers in the state (Knight and Notestein Chapter 1). In fact, the current view among many (if not all) scientists and resource ma nagers is that plant and algal populations in Florida’s spring-dominated ecosy stems are undergoing major structural and functional changes, due, in large pa rt, to increases in anthropogeni c enrichment of nutrient levels in groundwater and the consequent nu trification of spring discharges.

This widely held view is not without merit. Th ere is, in fact, a growing body of literature on the chemistry of spring discharges that supports th e conclusion that nutrien t concentrations in springs, and nitrate concentrations in particular, are increasing (Cohen et al . Chapter 2). The perception that increased algal biomass has led to the displacement of native macrophytes, on the other hand, appears to be based largely on anec dotal accounts and qualita tive observations (see Frazer et al . 2006b). Nevertheless, the prevalence of such reports is a compelling reason for concern. It is fair to say, however , that there is little direct evidence that either definitively links the observed increases in algae to increased nitr ate delivery to spring systems or demonstrates that the purported changes have compromised the ecological h ealth and integrity of these systems. Our goal is to articulate what we know about the factors that affect submersed plants and algae in Florida’s spring systems, the rela tionships among nutrients and spring flora, the habitat associations of spring fauna, and the potential toxic e ffects of some nitrogen compounds. From this synthesis, we aim to draw guidan ce or recommendations for policy and management. Where sufficient information is lacking, our goal is to suggest potential remedies.

PRIMARY PRODUCERS IN SPRING EC OSYSTEMS AND FACTORS AFFECTING THEIR ABUNDANCE AND DISTRIBUTION

Florida’s springs often are perceived as stab le-state ecosystems in which major chemical constituents, temperature and discharge characte ristics of individual springs are essentially constant (Odum 1957b). In fact, there is substa ntial variability in chemical and physical attributes among the state’s 700+ springs (Odum 1957a; Jones et al . 1997; Scott et al . 2002; Strong 2004), and, in some cases, variability wi thin spring systems (O dum 1957b). Patterns in184 the abundance and distribution of spring-associated flora and fauna should be expected to vary accordingly. We would argue that spring-dominated ecosystems, like all other ecosystems, are, in fact, dynamic in nature and exist (on some scale) as heter ogeneous environments in which various chemical, physical and biological processes interact to determine ecological structure and function. Thus, the inherent complexities in sp ring-dominated ecosystems must be recognized, and any broad generalizations re lated to observed patterns and presumed processes should be evaluated objectively.

For example, it is important to recognize that nitrate availability is only one of many factors with the potential to directly influence the abundance and distribution of plants and algae in Florida’s spring systems. Other e ssential nutrients , particularly phosphorus, also are likely to be important determinants of plant and algal abundance (Frazer et al . 2001a; Notestein et al . 2003; Kurz et al . 2004; Frazer et al . 2006b; Stevenson et al . 2007). In addition, we would be remiss if we failed to acknowledge that ot her, naturally variable, chemi cal properties of groundwaters and surface waters influence th e vegetative characteristics of Florida’s springs and their downstream receiving waters. For example, concentrations of dissolved gases, such as oxygen and carbon dioxide, exert a profound influen ce on production of submersed aquatic plants and algae (Sand-Jensen 1989, Sand-Jensen and Borum 1991). In addition, vegetative assemblages in several of Florida’s sp ring systems have been shown to va ry with specific conductance and salinity (Whitf ord 1956; Frazer et al . 2001b; Hoyer et al . 2004; Frazer et al . 2006a). One also must consider physical attributes of various spring systems in any attempt to ascertain the relative importance of factors affecting broad-s cale or fine-scale pattern s in plant and algal abundance. For example, water depth, light availa bility, stream velocity, flow regimes, and characteristics of substrates all represent im portant considerations (Butcher 1933, Hynes 1972).

Beyond chemical and physical influences, biologi cal interactions and ecological processes may influence patterns of plant and algal abundance. Grazing on aquatic plants and algae can be significant, and it may serve to moderate th e potential negative influences of nutrient over-enrichment (Duarte 1995; Hauxwell et al . 1998). In addition, competition for space, light and/or nutrients can alter the sp ecies composition of algal and plant assemblages. Changes in these patterns may, in turn, affect higher-order organisms present in the systems, particularly those that use submersed aquatic vegetation when foraging or taking refuge. Thus, the indirect effects of nutrients represent a fo cus of this chapter and a logical complement to the discussion of direct impacts on sp ring-associated flora.

Light

Light availability is arguably among the most important factors directly affecting the abundance and distribution of aquatic plants and algae. Without sufficient light, plants and algae cannot utilize nutrients effectively. Light transmittance in aquatic systems is affected primarily by suspended solids and dissolv ed organic substances (Kirk 1994). Suspended solids include algal cells (Canfield and Hodgson 1983) , nonvolatile suspended solids (Canfield and Bachmann 1981; Hoyer and Jones 1983) a nd detrital material (Buiteveld 1995).

All of the “in-water” constituents that reduce light availability are generally in low concentrations in groundwater serv ing as the origin of flow for Florida’s springs and spring runs. For example, Florida springs with discharges of 5 m 3 s -1 or higher (1 st magnitude springs) had the185 lowest concentrations of dissolved or ganic matter yet reported (13 ± 1.6 μM C L -1 ), and springs with lower discharges exhibited onl y somewhat higher values (Duarte et al . 2007). Thus, Florida’s springs are not surprisingly among the clearest waters in the world (Odum 1957b; Rosenau et al . 1977; Duarte and Canfield 1990a).

Nevertheless, light availability is likely to li mit the distribution and production of submersed aquatic vegetation in Florida’s spring systems in many instances. For example, Duarte and Canfield (1990b) investigated 31 springs and spri ng runs in interior centr al and northern Florida and concluded that shading by riparian vege tation was a primary determinant of primary productivity and macrophyte standing crops in these systems. It is interes ting to note that these investigators reported negative rates for daily maximum production in seve ral spring runs devoid of “vegetation,” and it was assumed that biochemical oxygen demand exceeded any production by benthic and/or pseudopl anktonic algae in these systems. It was further suggested that oxygen concentrations in groundwater em anating as spring discharge may, in some cases, be too low to support the respiratory demands of aquatic m acrophytes (see additional discussion below).

In a separate study of five spring-fed rivers along the Gulf coast, Frazer et al . (2001a) suggested that light availability was gene rally sufficient to support net phot osynthesis of submersed aquatic vegetation, but they also noted that terrestrial canopy coverage along some shoreline areas may reduce incident light to a level that restricts the abundance and di stribution of aqua tic plants and algae. Correlations with the pe rcentage of shoreline canopy cove r in the Weeki Wachee River, for example, accounted for 26% of the variation in total vegetative biomass, 21% of the variation in filamentous macroalgal biomass, and 8% of the variation in macrophyte biomass. As the percent canopy cover increas ed, the total vegetative biomass decreased, which led these authors to suggest that canopy cover in the Weeki Wa chee River does influence the distribution and abundance of aquatic plants.

There are few other quantitative studies of the relationship between light and submersed aquatic vegetation in Florida’s spring systems, and we did not find experimental investigations that might indicate the magnitude of light limitation in those springs subject to shading by riparian vegetation. This is somewhat surprising as Od um suggested in his seminal work (Odum 1957b) that primary production in Silver Springs (and other spring syst ems in Florida; Odum 1957a) is linearly related to light intensity.

In addition to shading by riparian vegetation, there is a widely held perception that the accumulation of algae on macrophytes reduces the photosynthetic efficiency of the host plant due to shading. This scenario is , in fact, a cruc ial element of the argument that nutrification of Florida’s spring systems has led to a proliferatio n of nuisance algae and concomitant decline of native macrophytes. A key to effective management of Florida’s springs will be to understand the factors and processes that interact to either accelerate or decelerate this eutrophication progression scheme (see Duar te 1995 and also Valiela et al . 1997). At this time, there is no direct evidence that accumulations of periphyton on macr ophytes have reduced their growth rates in Florida’s spring systems. The potential for such an effect has been inferred from laboratory studies with portions of plan ts (Sand-Jensen 1977; Chen et al. 2007) and relatively few, empirical field studies in estuarine and coastal marine systems (Deegan et al . 2002; see also Heck et al . 2000). It is likely, however, that the diffusion of carbon dioxide (CO 2 ) or active uptake of bicarbonate (HCO 3 - ) is hindered by increased boundary layer effects created by a layer186 of algae. Thus, the effects of changes in light c limate will be mediated by stream velocity and the supply of carbon, especially disso lved inorganic carbon (DIC). Ther e is a need to quantify the relationships among light climate, DIC availabi lity, stream velocity , algal abundance and macrophyte production because such relationships have profound implications for management of spring systems.

Carbon and Oxygen Balance in Macrophytes

At saturating light levels, the maximum rate of photosynthetic carbon fixation by submersed vascular plants exceeds the respiratory carbon demand of leaves and belowground tissues (Hemminga and Duarte 2000). This outcome holds only if DIC is in ample supply. Because concentrations of DIC in Florida’s springs ar e naturally high due to microbially mediated remineralization of organic matter in the groundw ater system, this situation typically applies (Inglett et al . Chapter 3).

Many springs (particularly 1 st magnitude springs) are supersatur ated with DIC and there is, in fact, a net flux of gaseous CO 2 from the springs to the atmosphe re at the point of discharge and for some distance downstream during the equilibration process (Duarte et al . 2007). Thus, the supply of inorganic carbon is not likely to limit pr imary production in spring systems, at least in those areas in close proximity to a groundwater source. Mediation by biological (algal mat formation) or physical (water ve locity) processes, however, may re sult in localized depletions of DIC. At such fine scales, the potential exists for DIC limitation, but we have insufficient data to evaluate the existence or ultimate effect of such limitation.

Oxygen is a byproduct of photosynthesis and a ke y chemical constituent of all systems that support biological communities. Any factor or co mbination of factors that reduces primary production in spring systems will necessarily re duce the availability of dissolved oxygen. The potential consequences of redu ced oxygen availability on faunal assemblages may be intuitively obvious, but the consequences of low oxygen concentrations on flora are less clear.

Rooted aquatic plants require di ssolved oxygen to support aerobi c respiration, especially in belowground tissues. Reduced photosynthetic ca pacity compromises the supply of oxygen, and such a change may negatively affect plants by disrupting their carbon ba lance (see Hemminga and Duarte 2000). Oxygen deficits might be expected to occu r where macrophytes are highly epiphytized or covered by extensiv e macroalgal growth because of increased respiration and/or boundary layer effects on diffusion. We reiterate, how ever, that there are no experimental studies clearly demonstrating a negative effect of epiphytes or macr oalgae on vascular plants in Florida’s spring systems, although there are a number of empiri cal investigations that are suggestive of such a cause and effect relationship (e.g., Frazer et al . 2006b; Jacoby et al . 2007). In those spring systems that exhibit chronic hypoxi a, one might expect rooted macrophytes to be more susceptible to the potential effects of al gal overgrowth. It is in teresting to note that Stevenson et al . (2004) reported a significant negative relationship between dissolved oxygen concentrations and algal cover base d on a broad survey of springs in Florida. The nature of this relationship merits further invest igation, especially in light of recent information that suggests a broad-scale decline in dissolv ed oxygen concentra tions in Florida’s springs (Heffernan et al

. unpublished data).187 Stream Velocity and Substrate

Water movement can affect the gas exchange, phys iology and growth form of plants and algae, particularly periphyton. In additi on, stream velocity can determ ine the characteristics of the substrate, which, in turn, influence the viability of plants and algae.

For example, a stream/river bed consisting of bare rock or large stones th at are continually being rolled or scoured will have li ttle submersed aquatic vegetati on, while a river bed that is comprised largely of mud, silt and sand has the pot ential to support abundant aquatic vegetation (Butcher 1933; Allan 1995). Systems with flow rates greater than 0.60 m sec -1 tend to have heavy rock substrates, and system s with progressively slower flow have light gravel, sand, silt and mud substrates that are more favorable for aquatic plant growth.

All of Florida’s spring runs are considered low gradient streams, and they seldom exhibit flow rates in excess of 0.60 m sec -1 . Therefore, substrates within these systems generally are conducive to the colonization and subsequent production of many types of submersed aquatic vegetation (see Frazer et al . 2001a) if nutrients, light and other conditions are suitable. However, sections of spring systems may not sup port all types of submer sed aquatic vegetation. For example, exposed limestone is prevalent in the immediate vicinity of spring vents, and outcrops of limestone are distri buted patchily throughout many spring runs. Rooted macrophytes will not colonize these areas, although algae does grow successfully near spring vents and limestone outcrops are colonized by al gae and bryophytes (Whitford 1956; Frazer et al . 2001a; Kurz et al . 2003 and 2004; Hoyer et al . 2004; Stevenson et al . 2004 and 2007).

In the context of this review, st ream velocity will be a concern pr imarily because of its effects on the periphyton assemblages associated with ma crophytes. Although submersed vascular plants and large macroalgae in sp rings draw considerable attention, epiphytic microalgae also are an important component of spring systems. In fact, epiphytic algae and benthic mats can be major, or even dominant, elements of the primary pr oducer community, in terms of both productivity and biomass (Quinlan et al . 2008). The composition of algal communities, however, varies considerably between different spring ecosystems (Whitford 1956), and it might be expected to vary temporally due to seasonal differences in light availability (see above) or more stochastic variations in flow regimes that occur as a conseque nce of climate variab ility or anthropogenic demands on source wa ter (Odum 1957b).

In systems with relatively high flow rates, periphyton associat ed with macrophytes or other substrates are comprised mostly of low-lying mucilaginous diatoms th at are resistant to sloughing even at high velocities (Biggs et al . 1998). Long, filamentous algal species are much less resistant to high velocity environm ents. Indeed, for such a species, Biggs et al . (1998) reported a linear decrease in ash- free dry mass of periphyton as a function of near-bed velocity. To our knowledge, there are no rigorous experiment al studies in Florida spring systems that quantify the effects of flow on periphyton biom ass or species composition. There are, however, several empirical investigations in which both flow and periphyton abundance on macrophytes have been measured. For example, Frazer et al . (2001a) found no significant relationship between flow rates and periphyt on associated with macrophytes in any of the five spring-fed, coastal rivers they investigat ed. In a subsequent study of these same systems, however, statistically significant decreas es in periphyton associated w ith macrophytes in the Weeki188 Wachee River coincided with an increase in flow (Frazer et al. 2006b). Kurz et al . (2004) found no significant relationship betw een periphyton abundance on macr ophytes and flow in the main stem of the Ichetucknee River. It should be noted, however, that periphyton abundance on macrophytes was significantly higher in seve ral feeder springs wh ere flow rates were significantly less (Kurz et al . 2004; see also Dormsjo 2007; Politano unpublished data).

These findings are counter-intuitive because re duced flows are expected to increase boundary layer thickness. Boundary layers determine, in part, the availability of nutrients (or other chemicals such as dissolved oxygen and free di ssolved inorganic carbon) to the periphyton. When boundary layers are thic k, as in areas with slow-movi ng water, nutrients and other chemicals must diffuse or be act ively transported across large di stances. Thinner boundary layers facilitate contact between cells and nutrients or other chemicals. The steepness of the concentration gradient across the boundary layer determines transport rate, per Fick’s law:

∂

φ J = – D ∂χ

where J is the diffusion flux in, for example, moles m -2 s -1

, D is the diffusion coefficient or diffusivity in m2 s -1 (determined by temperature and the properties of both materials – here, water and cell cytoplasm), φ is the concentration gradient per unit length in moles m -3 , and χ is the length in m of the boundary layer. Depletion of nutrients in a slow -moving system decreases the concentration gradient, ∂ φ

/

∂χ , so periphyton come to have lim ited access to nutrients. Even minute increases in stream velocity (< 0.1 cm/s ) have been shown to increase photosynthesis of aquatic macrophytes (Westlake 1967), presumably by maintaining a steep concentration gradient of nutrients near the plants’ surfaces or by re ducing the concentration of metabolic by-products that might impair photosynthesis. In general, pe riphyton biomass increases as velocity increases until shear reaches a level that initiates sloughing (Biggs and Hickey 1994; Biggs et al . 1998). Clearly, the relationship between water velocity and periphyton dynamics in Florida’s spring systems merits further investigation.

Nutrients

Although modified and influenced by other factors, the availab ility of nutrients plays a major role in governing productivity in natural systems. In freshwaters, phosphorus generally is considered the nutrient that most frequently limits the production of plants and algae (Wiebe 1931; Schindler 1971). Nitrogen, however , can and does limit primary production in many instances (see Vitousek and Howarth 1991). Other macronutrients and micronutrients (e.g., potassium, calcium, magnesium, sulfur, iron, manganese, copper, zinc, molybdenum, sodium, cobalt, chlorine, bromine, silicon, boron, and iodine) also can limit plant and algal production depending on their relative abundanc es and the particular physiologi es of the taxa present in the system. In some situations, plants and algae may be limited by the avai lability of dissolved inorganic carbon, although there is li ttle indication that this might be the case in any of Florida’s springs, except perhaps in algal mats. Due to th eir import, our focus here is on nitrogen and phosphorus and their potential to limit the production of subm ersed aquatic vegetation in Florida’s springs and spring runs.189 Rooted vascular plants, e.g., Vallisneria americana and Sagittaria kurziana , are generally assumed to acquire nitrogen and phosphorus larg ely from sediment porewaters (Chambers

et al . 1989), although some research suggests that many plants (som e of which are present in Florida’s spring systems) are capable of nutri ent uptake through their aboveground tissues (see DeBrabandere et al . 2007 and references therein). There has been no experimental work to demonstrate unequivocally that any submersed vascul ar plant in a Florida, spring-fed system is limited by either nitrogen or phosph orus. Moreover, empirical data generated in a large number of spring systems suggests that there is no sign ificant linkage between surface water nutrients and production of macrophytes.

For example, Canfield and Hoyer (1988a and b) carried out an intensive investigation of 17 inland streams (many of which were spring -fed) to determine if the abundance of aquatic macrophytes (composite category including both root ed vegetation and macroalgae) was related to nutrient enrichment. Standing cr ops of macrophytes in these syst ems were not correlated with in-stream concentrations of total phosphorus or total nitrogen. Furthermore, in a subsequent investigation, Duarte and Canfield (1990b) sa mpled 31 spring runs and similarly found that standing crops of submerged vegetation were not significantly correlated with either total nitrogen or total phosphorus.

Frazer et al . (2001a) investigated the re lationship between nutrient co ncentrations in surface waters and characteristics of submersed aquati c vegetation in five spring-fed coastal rivers subjected to increasing nitrate delivery. They found that statis tically signific ant, positive relationships between nitrate and standing crops of either macrophytes or filamentous macroalgae were evident only in the Chassahow itzka and Homosassa Rivers (Table 1). Mean nitrate concentrations declined markedly at downstream sites in both rivers, which is indicative of rapid nitrate remova l. Subsequent research in the Chassahowitzka River suggests, however, that the vast majority of the nitrate re moval occurs as a conse quence of denitrification (Saunders 2007; see also DeBrandere et al . 2007) rather than uptake and assimilation of nitrate by submersed aquatic vegetation. Denitrification also may be operating in the Homosassa River where nitrate concentrations decline marked ly along the river’s length even though little vegetation remains in the system (Frazer et al . 2006b). It is noteworthy that Frazer et al . (2001a) reported that soluble reactive phosphorus was correlated with macrophyte and macroalgal biomass in the Chassahowitzka and Homosassa ri vers, and they suggested that phosphorus might be a more important determinant of vegetative biomass than nitrate in these two rivers. Notestein

et al . (2003) subsequently determined that phosphorus does in fact limit algal growth in the Chassahowitzka River. The addition of soluble reactive phosphorus resulted in significantly higher periphyton accumulation on artificial substr ates relative to nitrat e amended treatments or controls (no nutrient addition). Frazer et al . (2001a) reported that soluble reactive phosphorus was correlated with macroalgal biomass in the W eeki Wachee River as well, so phosphorus also may limit the growth of algae in this river.190 Table 1. Pairwise Pearson Product-Moment correlations between log 10 –transformed annual mean nutrient concentrations and log 10 –transformed summer mean biomass for macroalgae, all vegetation, macrophytes and periphyton from ten transects in four coastal rivers during 1998, 1999 and 2000 (n = 30). Data for the Withlacoochee River were not shown because plants were largely absent from the system below the dam at Inglis where the substrate (largely rock) was unfavorable (from Frazer et al . 2001a).

Nutrient Biomass Correlati on coefficient Probability C HASSAHOWITZKA

Ammonium Macroalgae -0.31 0.09 Nitrate Macroalgae 0.63 < 0.05 Soluble reactive phosphorus Macroalgae 0.55 < 0.05 Total nitrogen Macroalgae 0.21 0.27 Total phosphorus Macroalgae -0.19 0.31 Ammonium All vegetation -0.46 < 0.05 Nitrate All vegetation 0.79 < 0.05 Soluble reactive phosphorus All vegetation 0.68 < 0.05 Total nitrogen All vegetation 0.29 0.12 Total phosphorus All vegetation -0.31 0.10 Ammonium Macrophytes -0.38 < 0.05 Nitrate Macrophytes 0.70 < 0.05 Soluble reactive phosphorus Macrophytes 0.61 < 0.05 Total nitrogen Macrophytes 0.33 < 0.05 Total phosphorus Macrophytes -0.20 0.29 Ammonium Periphyton -0.34 0.07 Nitrate Periphyton 0.28 0.14 Soluble reactive phosphorus Periphyton 0.23 0.23 Total nitrogen Periphyton 0.36 0.06 Total phosphorus Periphyton -0.42 0.83 C RYSTAL

Ammonium Macroalgae -0.22 0.24 Nitrate Macroalgae 0.38 < 0.05 Soluble reactive phosphorus Macroalgae -0.46 < 0.05 Total nitrogen Macroalgae -0.48 < 0.05 Total phosphorus Macroalgae -0.43 < 0.05 Ammonium All vegetation -0.13 0.48 Nitrate All vegetation 0.23 0.22 Soluble reactive phosphorus All vegetation -0.33 0.07 Total nitrogen All vegetation -0.70 < 0.05

C RYSTAL (continued)

Total phosphorus All vegetation -0.28 0.14 Ammonium Macrophytes -0.12 0.52 Nitrate Macrophytes 0.21 0.26 Soluble reactive phosphorus Macrophytes -0.32 0.09 Total nitrogen Macrophytes -0.69 < 0.05 Total phosphorus Macrophytes -0.26 0.16 Ammonium Periphyton -0.44 < 0.05 Nitrate Periphyton 0.18 0.40 Soluble reactive phosphorus Periphyton 0.10 0.63 Total nitrogen Periphyton -0.34 0.11191 Nutrient Biomass Correlati on coefficient Probability

Total phosphorus Periphyton 0.01 0.97 H OMOSASSA

Ammonium Macroalgae -0.34 0.07 Nitrate Macroalgae 0.60 < 0.05 Soluble reactive phosphorus Macroalgae 0.44 < 0.05 Total nitrogen Macroalgae 0.34 0.07 Total phosphorus Macroalgae -0.55 < 0.05 Ammonium All vegetation -0.37 < 0.05 Nitrate All vegetation 0.64 < 0.05 Soluble reactive phosphorus All vegetation 0.49 < 0.05 Total nitrogen All vegetation 0.25 0.19 Total phosphorus All vegetation -0.66 < 0.05 Ammonium Macrophytes -0.31 0.09 Nitrate Macrophytes 0.43 < 0.05 Soluble reactive phosphorus Macrophytes 0.32 0.09 Total nitrogen Macrophytes 0.02 0.92 Total phosphorus Macrophytes -0.62 < 0.05 Ammonium Periphyton 0.44 < 0.05 Nitrate Periphyton 0.12 0.53 Soluble reactive phosphorus Periphyton -0.25 0.20 Total nitrogen Periphyton 0.57 < 0.05 Total phosphorus Periphyton -0.37 < 0.05 W EEKI W ACHEE

Ammonium Macroalgae -0.47 < 0.05 Nitrate Macroalgae 0.20 0.30 Soluble reactive phosphorus Macroalgae 0.54 < 0.05 Total nitrogen Macroalgae -0.12 0.53 Total phosphorus Macroalgae -0.36 < 0.05 Ammonium All vegetation -0.48 < 0.05 Nitrate All vegetation -0.02 0.90 Soluble reactive phosphorus All vegetation 0.29 0.12 Total nitrogen All vegetation -0.23 0.21 Total phosphorus All vegetation -0.24 0.20 Ammonium Macrophytes -0.33 < 0.05 Nitrate Macrophytes -0.22 0.25 Soluble reactive phosphorus Macrophytes 0.04 0.82 Total nitrogen Macrophytes -0.30 0.10 Total phosphorus Macrophytes -0.13 0.51 Ammonium Periphyton 0.43 < 0.05 Nitrate Periphyton -0.46 < 0.05 Soluble reactive phosphorus Periphyton -0.32 0.12 Total nitrogen Periphyton -0.32 0.13 Total phosphorus Periphyton -0.04 0.85

Kurz et al . (2004) reported no statistica lly significant, positive rela tionships between nutrients and coverage or biomass of submersed aquatic ve getation in the main stem of the Ichetucknee River. There was a weak, but statistically significant, negative correlation between terrestrial canopy cover and submersed aquatic vegetation cover suggesting that , in some cases, light may limit the distribution of macrophytes. In a compleme ntary analysis, these i nvestigators related a suite of chemical and physical parameters to periphyton abundance on macrophytes in several small spring runs adjoining the main river. Pe riphyton abundance was not correlated with either nitrate or total nitrogen concen trations, but it was significantly co rrelated with soluble reactive192 phosphorus concentrations s uggesting phosphorus limitation.

The issue of nutrient limitation of algal produc tion can be viewed at several levels of organization, from individual algal cells or fila ments to complex mat communities. For example, the nitrate half saturation constants for the growth of the common filamentous green alga Cladophora range from 100–200 μg l -1 (Lohman and Priscu 1992), and recent research on growth saturating concentr ations of nitrate for Lyngbya wollei suggests values between 100–200 μg l -1 (Stevenson et al . 2007). Saturation levels for the growth of natural populations of benthic algae in Sycamore Creek, Arizona (Gri mm and Fisher 1986) and Saline Creek, Missouri (Lohman et al . 1991) were 55 and 100 μg l -1 , respectively. Thus, the limited literature listing levels of nitrate that saturate th e growth rates of natural, benthic, algal mats suggests that nitrate concentrations currently observed in the wa ter column of many springs are not limiting. However, this does not preclude the possibility of nitrogen limitation in large, well-developed algal mat communities, where growth may be imp acted by diffusion rates and nutrient gradients within the mat. In fact, Stevenson et al . (2007) showed that water column concentrations limiting growth may be considerably higher for mats than for exposed, individual algal cells or filaments. Half saturation constants for phosphorus exhibit similar variability. Constants have been reported to vary from 9 to 45 μg l -1 for the green alga Spirogyra (Borchardt et al . 1994) and from 15 to 240 μg l -1 for Cladophora (Lohman and Priscu 1992). Lower values of 0.6 to 6 μg l -1 were reported for mixed periphytic diatoms (B othwell 1985). Half sa turation constants for natural, benthic, algal mats ranged from 7 to 50 μg l -1 (Wuhrmann and Eichenberger 1975; Welch et al . 1988; Horner et al . 1990).

In contrast to the widespread and dramatic incr ease in nitrogen levels in many Florida springs, phosphorus concentrations are reporte d to have remained similar to historic levels (Stevenson

et al . 2007; Munch et al . 2006). Current phosphorus concentrations vary considerably among springs, from low levels in system s like the Weeki Wachee River (Frazer et al . 2001a and 2006b) to high concentrations in systems like Silver Springs and the Ichetucknee River and springs (Munch et al . 2006; Kurz et al . 2003 and 2004). Some springs exhib it concentrations that fall below the reported saturation levels for growth of benthic algae. The possibility that phosphorus may be a limiting factor for algal growth in some springs has been further highlighted by recent observations of high N/P ratios in plant and algal samples from several Florida springs (Stevenson et al . 2007). However, some springs with very low phosphorus concentrations in the water column have extensive benthic algal mats (Frazer et al . 2001a and 2006b). We hypothesize that the mobilization of phosphorus from sediment s may provide sufficient nutrients to sustain this biomass. This hypothesis is supported by observations of increasing phos phorus levels in the pore water of some benthi c algal mats (Stevenson et al . 2007). The absence of sufficient information on nutrient dynamics within algal ma t communities in spri ng ecosystems hampers our ability to predict the outcome s of changes in nutrient loads.

Community Responses to Increased Nutrients

Plant and algal communities appear to exhib it qualitatively similar responses to increased nutrient loads. The primary responses and the mechanisms by which algae replace submersed vascular plants have been reviewed in depth by Duarte (1995), and we reiterate the key points from his significant contribution.193 Interactions between supplies of nutrients and light availability, two resources that most often limit the production of plants and algae, are at th e root of observed change s in floral assemblages (Duarte 1995). Because both plants and algae require the same basic resources (nutrients and light), they are likely to compete (intra-specifically and inter-specifically) for the resource that is in the shortest supply (Liebig’s Law of the Mi nimum). At low nutrient concentrations in the water column and under favorable light conditi ons, slow-growing, rooted vascular plants are superior competitors owing to their relativel y low nutrient requirements, their capacity to efficiently recycle nutrients internally and their ability to access sedimentary nutrient stores. In contrast, algae become more abundant when nutr ient limitation is released, in large part, by capturing light more efficiently.

In the eutrophication progression scheme laid out by Duarte (1995), faster growing macroalgae and microflora, i.e., periphyton, can ultimately ove rgrow long-lived and relatively slow-growing plants and algae, e.g., rooted vascular plants and macroalgae with thick fronds. This change exacerbates light limitation and reduces scope for growth for the vascular plants and “thick” algae. Ultimately, these long-li ved forms are outcompeted and extirpated from the system. In extreme cases of nutrient over-enrichment, the fast-growing, benthic algae will be replaced by planktonic microalgae via the same mechanism. Th e short residence times for water in Florida’s springs and most spring runs make this final stage in the progression unlikely.

A critical point in Duarte’s (1995) review, however , is that a change in the structure of the vegetative community is not a co ntinuous, gradual process that r uns parallel to the rate of increased nutrient delivery. Rather, changes occu r abruptly. Sudden shifts in the community, in many cases, are not tightly coupled to increased nutrient loading, and once a shift has occurred, the system can become decoupled from existing c oncentrations of nutrients in the water column. Thus, indirect effects of nutri ent inputs and complex feedback mechanisms must be involved. For example, fast-growing algae tend to put less energy into creating structural tissues so they tend to be subject to higher grazing rates. In creased algal respirati on during blooms may create hypoxic or anoxic conditions that kill grazers, whic h decreases grazing pressure on the fastgrowing algae and creates a positive feedback in the eutrophication prog ression. Thus, there are likely to be nonlinear responses and thresholds that need to be fully understood in order to predict the longer-term consequences of nutrien t over-enrichment for vegetative communities in Florida’s various spring systems.

One potentially key shift in the floral communiti es in Florida’s springs is from palatable species to unpalatable species, especially Lyngbya wollei . Recent evidence indicates that this cyanobacterium may have the capacity to produce neurotoxins that affect ecosystem structure and function, as well as human health (Carmichael et al . 1997; Onodera et al . 1997; Teneva

et al . 2003; Berry et al . 2004). For example, it has been hypothe sized that heavy infestations of Lyngbya on the skin of manatees in springs may be related to recent obs ervations of health problems (Bledsoe et al . 2006). From a human health persp ective, it is unclear whether Lyngbya toxins are associated with recent reports of contact dermititus by recreational users of springs.

A final point relates to the recovery of s ubmerged aquatic vegetation communities following nutrient remediation and reduction. Recovery is likel y to be a very slow process, in large part, because it involves the replacement of fast-gro wing species with relatively slow-growing species. Duarte (1995) employed a simulation model that predicted recovery times of centuries194 for systems originally occupied by slow-growing seagrasses. It is worth noting here that many of the historically dominant vascular pl ants in Florida’s spring systems (e.g., Vallisneria americana and Sagittaria kurziana ) share morphologies and life hist ories with these seagrasses. Expectations regarding responses to nutrient reduction should be tempered by these considerations.

EFFECTS OF NUTRIENTS ON FAUNA

Florida’s 700+ springs represent important aquatic ha bitats for fauna that are subject to stress by increased nutrients. Key elements of spring sy stems resemble key elements of many lotic or flowing water systems (Figure 2; Vannote et al . 1980; USEPA 2006). Therefore, effects of nutrients can be elucidated from science c onducted in these springs and by extracting key principles drawn from beyond Florida and from other aquatic systems.

Figure 2 – Key components of lotic ecosystems (USEPA 2006).195 Components of Spring Ecosystems

In general, phytoplankton and zooplankton represent minor components of lotic systems until they are relatively large, deep and slow flowing; therefore, th e focus falls on benthic organisms and trophic links. The base of tr ophic webs in flowing systems includes allocthonous material, or external inputs such as leaves from terrestri al vegetation, and autocthonous material that is produced in the system. Autocthono us material becomes more important as shading by riparian vegetation decreases an d in-stream primary producers photos ynthesize more effectively. Key primary producers in flowing systems include m acroalgae, vascular macrophytes and periphyton, with these algal species growing on almost al l biotic and abiotic surfaces. Primary producers supply food to primary consumers dominated num erically by a variety of invertebrates, especially snails and larvae of aquatic inse cts. Primary and secondary consumers include invertebrates and fish, with piscivorous fish representing the major tertiary consumers.

Past and recent surveys of faunal assemblages in Florida’s springs documented species that fill the expected ecological niches. For example, i nvertebrates from five functional feeding groups were found in the Wekiva and Ichetucknee Rivers (Warren et al . 2000; Steigerwalt 2005). These groups comprise grazers on attached algae or microbiota, shredders that consume coarse particles of living or dead plant material, gatherers that colle ct fine particles of organic detritus, filterers of suspended particulate organic matter, and pr edators feeding on live animal prey. Similarly, fish in the Wekiva River spanned the expected trophic roles, including de tritivores, herbivores that often feed on algae, planktivores, invertiv ores and piscivores of various sizes (Warren

et al . 2000).

Karst habitats in Florida support cave fauna or stygobiota and obligate fauna in springs that are not found in other types of aquatic systems (W alsh 2001; Shelton 2005). Nearly one third of stygobites have recorded distributions limited to si ngle cave systems, and 67% of these taxa have been found in 10 systems or less. Therefore, a hi gh percentage of these sp ecies are listed as of special concern, rare, threatened or endangered. Obligate fauna of springs include hydrobiid and pleurocerid snails with highly re stricted ranges leading to about eight species being considered endangered, threatened or of special concern (Walsh 2001). For example, the upper St Johns River basin holds ten species of rare and endemic snails, with Aphaostracon monas , Floridobia alexander

, Floridobia petrifons and Floridobia wekiwae found only at a si ngle spring and Aphaostracon pycnum found at two springs (Shelton 2005).

In aquatic systems, nutrients exert effects on animals primarily through changes in primary producers or direct toxicity. Changes in prim ary producers flow through ecosystems to affect animals primarily by altering habitats or trophic interactions. This section focuses on these indirect effects and direct toxicity.196 Relationships Between Nutrients and Invertebrates in Florida Springs

Nutrients can affect the type and amount of flor a found in spring systems. In turn, changes in vascular plants, attached macroa lgae and mats of filamentous al gae can alter the food and refuge available to invertebrates in Fl orida springs. However, few studies have assessed invertebrate assemblages in different types of ve getation to obtain a synoptic view. A survey of the Homossassa and Weeki Wach ee Rivers found differences in the number of species and relative densities of certain insect s along the length of the waterways (Sloan 1956). Habitats were classified according to the areal coverage of do minant macrophytes. Our reanalysis of the data showed that numbers of species did not correlate with variations in coverage from 10% to 70%, and the relative density of mayflies was positively correlated with increasing cover of macrophytes in the Homosassa River but not in the W eeki Wachee River. In general, the insect assemblages appeared to vary primarily according to concentrations of dissolved oxygen near spring boi ls and salinity near the estuarine endpoints of the rivers.

In Sulphur Springs Run, quantitativ e samples taken in bare sediment and algal mats with a petite Ponar dredge showed no significant differenc es in numbers of individuals and no obvious differences in assemblage composition (Allen et al . 2001). An introduced snail, Tarebia granifera , dominated all samples, and evenness and diversity indices indicated assemblages dominated by few species.

A more in-depth survey of four key habitats in the Wekiva River yielded similar an d statistically equal numbers of invertebrates, i.e., 20,000–34,000 individuals m –2 (Warren et al . 2000). Sampling methods were designed specifically for each habitat; however, the comparative efficiencies of the different me thods were not assessed. Theref ore, the data support general comparisons of faunal assemblages among habita ts, but detailed quantitative comparisons may prove unreliable.

The relative abundance of different functional t ypes varied among habitats (Figure 3; Warren

et al . 2000). Gatherers were found in all habitats, and they dominated ba re sediments. The relative importance of grazers increased in floating mats composed primarily of Hydrocotyle umbellate (water pennywort) and beds of Vallisneria americana (tape grass). Predators were found in all habitats in similar relative abundan ces. Filterers were most common in beds of Vallisneria americana and essentially absent from floating mats, which was attributed to the increased flow measured at sites with Vallisneria americana . Shredders, which are considered key components of most shallow freshwater system s, represented less than 1% of the organisms sampled. Overall, this survey indicated that a ltering the areal coverage and species compositions of habitats might lead to change s in invertebrate assemblages.197 = grazers, = gatherers, = filterers, = predators Figure 3 – Relative abundance of functional groups of aquatic invertebrates in May/June and October samples taken at multiple sites in the Wekiva River (Warren et al. 2000).

Invertebrate assemblages in lo tic systems are used as indicat ors of ecological health (e.g., USEPA 2006). For example, Flor ida has two biotic indices, the Stream Condition Index (SCI) that relies on laboratory processing of sample s and the more rapid Biological Reconnaissance (BioRecon) that uses so rting in the field (Fore et al . 2006).

Indices combine various metrics derived from standardized sampling of invertebrate assemblages, e.g., number of taxa, abundances of specific taxa, and re lative abundances or percentage dominance of select ed taxa. The concept that inve rtebrate taxa exhibit some consistency in their response to di sturbances underpins biotic indi ces. In Florida and elsewhere, biotic indices were developed empirically using samples of invertebrates collected from locales chosen to represent different leve ls of disturbance. The levels of disturbance used to classify locales combined multiple factors, such as habitat quality, one or more water quality parameters, and evaluation of land use. Overall, these indices provide indications of differences or changes in assemblages; however, the causes of such diffe rences or changes typi cally are derived from correlations with auxiliary data on water quality, habitat quality, or other putative forcing factors.198 Interpreting correlations between biotic indi ces and forcing factor s may be less than straightforward. For example, over a 14-year period, invertebrate ab undance was positively correlated with nitrate concentrations in the waters of the Suwannee River basin (Hornsby

et al . 2004). In contrast, number of taxa, diversity, and evenness we re negatively correlated with nitrate concentrations (Hornsby et al . 2004). These results were inte rpreted as indicating that higher nitrate concentrations led to an increase in the absolute number of invertebrates and a decrease in the biodiversity of the invertebrate assemblage. However, 17 of the 40 taxa that were common in these samples have been classified as indicative of clean water in the Florida Department of Environmental Protection’s Flor ida Index, a metric for surface water quality based on invertebrate assemblages.

In Florida, assessments have been conducted in various streams. Analysis of assessments conducted after recalibration of the Stream Cond ition Index (SCI) showed that SCI values were not significantly co rrelated with nitrate–nitrite concentra tions or habitat assessment scores (Table 2; Figure 4). In nine springs with very poor to poor SCI va lues, nitrate–nitrite concentrations ranged from 0.02 to 1.90 mg L –1 . Concentrations in four springs with fair SCI values ranged from 0.06 to 1.60 mg L –1 , which is a nearly identical range. Correlations with nitrate–nitrite con centrations and habitat assessment scor es accounted for less than 1% and 16% of the variation in the data, respectively. In co ntrast, SCI values were significantly correlated with dissolved oxygen concentrations, with a pos itive correlation that ac counted for 36% of the variation in the data (Table 2; Figure 4). Such re sults illustrate the complexities associated with interpreting biotic indices, and they provide lit tle insight into the form and magnitude of ecological responses to nutrient concentrations.

Table 2. Results from sampling in springs after recalibration of the Stream Condition Index (Agharahimi 2006a and b; Deck 2005a and b; Deck 2006a, b and c; Patronis 2006a, b, c, d, e and f). VP = very poor; P = poor; F = fair; < DL = below detection limit; < PQL = below practical quantitation limit; O = optimal; SO = sub-optimal; LO = low optimal; I–F = impairedfair; I = impaired; F = fair.

Spring Sampling NO
x NH 3 TP DO Overall
date Stream Condition Index (mg L –1 ) Habitat Assessment rank (DD-MM-YY) Score Rank Score Rank Rock (upper) 25-05-06 11 VP 1.40 < DL 0.11 4.2 134 O I–F Alexander 07-06-06 17 VP 0.04 < DL < PQL 3.7 119 SO I Volusia Blue 26-04-05 17 VP 0.57 < DL 0.08 2.5 105 SO I Wakulla 11-04-06 20 P 0.37 < DL < PQL 6.6 129 O I Manatee 21-06-06 21 P 1.90 < DL < PQL 2.6 98 SO I–F Ponce de Leon 22-06-06 32 P 0.25 < DL < PQL 7.0 114 SO I Silver Glen 08-06-06 32 P 0.02 < DL < PQL 6.8 118 SO I Wekiwa 26-04-05 35 P 1.10 < DL 0.13 3.6 117 SO I–F Blue Hole 22-06-06 36 P 0.42 0.02 < PQL 6.9 111 LO I–F Ichetucknee 18-04-06 40 P 0.57 < DL < PQL 4.5 145 O F Silver River 07-06-06 46 F 1.30 < DL < PQL 6.3 129 O F Rainbow 25-05-05 47 F 1.60 < DL < PQL — 140 O F Rock (lower) 25-08-06 47 F 1.40 < DL 0.10 5.4 126 O I–F Juniper 08-06-06 51 F 0.06 < DL < PQL 8.7 129 O F199 a

Correlation coefficient = 0.06 p = 0.83

0.0 0.5 1.0 1.5 2.0

0 10 20 30 40 50 60
Stream Condition IndexNitrate-Nitrite (mg L-1)

b

Correlation coefficient = 0.40 p = 0.15
0 40 80 120 160

0 10 20 30 40 50 60
Stream Condition IndexHabitat assessment score

c

Correlation coefficient = 0.60 p = 0.03

0 2 4 6 8 10

0 10 20 30 40 50 60
Stream Condition IndexDissolved oxygen (mg L-1)

Figure 4 – Correlations between Stream Conditi on Indices (SCI) and (a) nitrate–nitrite concentrations, (b) habitat asse ssment scores and (c) concentr ations of dissolved oxygen (Agharahimi 2006a and b; Deck 2005a and b; Deck 2006a, b and c; Patronis 2006a, b, c, d, e and f). black squares, triangles and circles = very poor SCI values ; gray squares, triangles and circles = poor SCI values; white squares, triangles and circles = fa ir SCI values; lines provided to illustrate the correlations200 Relationships Between Nutrients and Fish in Florida Springs

By affecting the flora of springs a nd other aquatic systems, nutrients also can affect fish that rely on this flora for food and shelter, including fi sh species that provide considerable economic value. Florida’s lotic systems contain 289 species , with 113 species native to North America and 16 species endemic to Florida and adjacen t portions of Alabama and Georgia (Bass 1991). Unfortunately, most studies, such as a broad survey of fish in streams throughout Florida, lack detailed information about assemblages in specific habitats (Bass and Cox 1985).

For example, the fish community in Silver Springs has been exam ined in two studies separated by approximately 50 years (Odum 1957b; Munch et al . 2006). The assemblages were reported to be similar, with the primary difference being re duced numbers of mullet and catfish ascribed to disrupted migratory patterns fo llowing construction of a dam. However, neither study focused on differences in fish assemblages among habitats.

In contrast, two habitat classes were included in a survey of 16 Florida streams (Canfield and Hoyer 1988a and b). These streams were sampled w ith blocknets and electrofishing gear as part of a study assessing the effects of nutrients from treated sewage effluent on the Wekiva River. Habitat classes were limited to open and closed forest canopy, and sampling was biased towards fish with total lengths of over 100 mm. Regardle ss, higher standing stocks and biomasses were found in reaches with an open forest canopy, wh ich coincided with r eaches where macrophytes were more abundant.

Detailed data on fish assemblages in specific ha bitats comes from a second study of the Wekiva River (Warren et al . 2000). Throw traps and a combination of blocknets and electrofishing were used to sample fish in four specific habitats: bare sediment, Hydrocotyle umbellate

, Nuphar advena subspecies lutea (spatterdock) and Vallisneria americana

. Throw traps and blocknets collected forty species of fish from various reaches, with more species collected from vege tated habitats (Figure 5). Hydrocotyle umbellate yielded the highest mean density of fish, followed by Vallisneria americana
, Nuphar advena and bare sediment (Figure 6). In contrast, samples from Vallisneria americana tended to yield the greatest biomass because these samples contained more sunfish in the family Centrarchidae (Figures 7–11). Samples from Hydrocotyle umbellate were dominated by small, livebearing Poeciliidae, and samples from Nuphar advena and bare sediment contained la rge numbers of small minnows in the family Cyprinidae and killifishes in the fa mily Fundulidae (Figures 8–11). Mullet also were found in samples taken from within blocknets set over bare sediment in the Lower Wekiva River (Figures 10 and 11). These differences in fish assemblages were attributed predominantly to interactions among characteristics of the habitats and the biology and ecology of fish species. Similar to comparisons for Silver Springs, broa d comparisons between the two studies of the Wekiva River indicated that the fish assemb lage had changed litt le between 1985–1987 and 1997 (Canfield and Hoyer 1988a; Warren et al . 2000). Thus, we do not have data linking changes in fish assemblages to changes in nutrient c oncentrations in springs at this time.201 a

b

Figure 5 – Species richness in (a) throw trap and (b ) blocknet samples (Warren et al. 2000).202 a

b

Figure 6 – Densities of fish in (a) throw trap and (b) blocknet samples (Warren et al. 2000).203 a

b

Figure 7 – Biomass of fish in (a) throw trap and (b) blocknet samples (Warren et al. 2000).204 a b

c d

Figure 8 – Assemblage composition by abundance in throw trap samples (Warren et al. 2000).205 a b

c d

Figure 9 – Assemblage composition by biomass in throw trap samples (Warren et al. 2000).206 Figure 10 – Assemblage composition by numbers for blocknet samples (Warren et al. 2000).

Figure 11 – Assemblage composition by biomass fo r blocknet samples (Warren et al. 2000).207 Summary of Relationships Between Nutrients and Fauna in Florida Springs

In summary, surveys of faunal assemblages in Fl orida springs have suggest ed the possibility of indirect influences from nutrien ts mediated through differences or changes in habitats. However, results of repeated surveys indicate faunal assemb lages have not changed even if nitrate levels have risen. Unfortunately, these st udies were not designed to identi fy the form and magnitude of interactions among nutrient inpu ts, primary producers and fauna.

Studies conducted in streams outside Florida a nd in lakes can provide additional insights into how nutrients affect habitat use and trophic relati onships. Such insights can guide future work in Florida’s spring ecosystems.

Nutrient Effects Operating Through Ecosystem Processes

Interactions among animals and habitats vary not only among species but also among life history stages or ages within species. Such interactions may result in nonlinear effects on key parameters such as growth rates (Mittelbach and Osenberg 1993; Diehl and Kornij ό w 1998). For example, fish may rely heavily on macrophytes for refuge fr om predation in the earl y portion of their life histories even if it reduces their growth rate (F igure 12). A nonlinear change in growth rate may be associated with changing habi tat use after achieving a size ref uge (S) from piscivores, which emphasizes the importance of having detailed information on the form and magnitude of responses to changes in habitats (Figure 11). Such information is not available for most species of fish, including most species found in Florida springs.Growth rate

Body size S

Figure 12 – Diagram depicting a change in growth rate experienced by a fish when it switches from a suboptimal habitat used as a refuge from predators to a more op timal habitat once it has achieved a size refuge (adapted from Diehl and Kornij ό w 1998). Solid line represents growth rate exhibited by a fish that switches from one h abitat to another once it achieves a size refuge from piscivores at body size S. Dotted line repres ents the growth rate of a fish able to use all available habitats and food resources due to th e absence of piscivores. The difference between the lines represents the cost or benefit associated with the use of each habitat.208 Up to some level, increased nutrient inputs can increase the pr oductivity of macrophytes, which provides more food for herbivores and may lead to increases in s econdary and tertiary consumers. The role of aquatic macrophytes as a food source for the most numerous grazers remains a controversial issue (Sheldon 1987; Brönmark 1990; Sheldon 1990; Newman 1991). Living macrophytes (mosses, flowering plants a nd macroalgae) have been regarded as a little used food source due to high C:N ratios (e.g., 20–1,340), larg e quantities of cellulose and lignin, decreased digestibility of their proteins a nd presence of secondary chemicals that retard grazing (Brönmark 1990). Larger grazers , such as manatees, turtles and Pomacea paludosa , the Florida apple snail, do graze macrophytes. The effect of this grazing can be significant (Hauxwell et al . 2004; Jacoby et al . 2007). In addition, some insects, crayfish, birds and mammals have been reported to redu ce macrophyte biomass through consumption or damage during grazing on periphyton (Cum mins and Klug 1979; Mann 1988; Lodge 1991; Lodge et al . 1998; Newman 1991).

Regardless of their contribu tion as a source of food, macr ophytes do provide a significant substratum for periphyton, which is grazed heavily by the most abundant animals. As discussed above, periphyton can outcompete macrophytes for resources, such as light and nutrients. Excessive growth of periphyton ha s the potential to lead to loss of macrophytes , with subsequent effects on refuge and trophic interactions.

Grazers can mediate the growth a nd effects of periphyton. Experiment s using elevated tiles as a mean of excluding the herbivor ous larvae of the caddisfly, Helicopsyche borealis , demonstrated effects of grazing on biomass and turnover of microalgae (Lamberti and Resh 1983). In grazed treatments, biomass, m easured as chlorophyll a cm –2 , was 5–500 times lower and turnover, measured as oxygen production per unit chlorophyll a , was approximately 5 times higher (Lamberti and Resh 1983). Meta-analyses of expe riments comparing grazers at ambient densities to treatments without grazers indicated that grazers significantly reduced periphyton biomass in 70% of experiments and measurably alte red the taxonomy or physiognomic structure of assemblages in 81% of experiments (Feminella and Hawkins 1995). Typically, less than 70% of experiments showed effects on productivity, and th e effects varied with the parameter measured (Feminella and Hawkins 1995). Significant effects were equally likely in laboratory and field experiments and across durations of less than and greater than 4 weeks, although factors such as the total amount of periphyton accrued, the type of grazer manipulated, and the total biomass of grazers influenced the magnitude of the effects recorded.

Grazers represent multiple levels of organismal complexity, and they interact with periphyton and each other in complex ways. For example, the extracellular pol ysaccharide sheath produced by Lyngbya wollei deterred grazing by the amphipod Hyallela azteca (Camacho and Thacker 2006). In streamside experimental trou ghs, herbivorous protozoa were shown to graze on benthic diatoms, with estimated consum ption rates varying from 4,000 diatoms cm -2 h -1 in systems with few or no macrograzers to 150 diatoms cm -2 h -1 in systems with large numbers of macrograzers, supposedly due to competition or predation (McCormick 1991).

Overall, the effects of nutrients in aquatic syst ems appear to be mediated by a combination of bottom-up and top-down effects. For example, larvae of the chironomid Orthocladius rivulorum build tubes and harvest the diatoms growi ng on them. Addition of phosphorus and phosphorus plus nitrogen to a stream led to a greater biomass of periphyton and larger larvae (Hershey209 et al . 1988). Nitrogen and phosphorus appeared to be co-limiting factors as indicated by greater accumulation of periphyton downstream of combined additions (Hershey et al . 1988). The system involved a positive feedback loop, with larvae in fertilized reaches growing larger, building larger tubes that supported more perip hyton, and, in turn, having more food to harvest (Hershey et al . 1988). In addition to interactions among nutrients, periphyton may respond to other limiting factors and grazi ng. Periphyton in streamside, fl ow-through channels exhibited statistically significant increases above controls only when nitrogen and phosphorus were added, light was increased, and grazing was eliminat ed (Rosemond 1993). In addition, manipulating single factors led to different responses. For example, incr eased light elicited both a physiological response resulti ng in decreased chlorophyll a concentrations in cells and a change in the periphyton assemblage when alg ae adapted to low light conditions were lost (Rosemond 1993).

All the bottom-up influences of nutrients and li ght and top-down influences of grazing can be mediated through actual or potential trophic inte ractions. Manipulations of pools in a stream demonstrated that grazing minnows controlled pe riphyton growth unless predatory fish were present, in which case predation and avoidance of predation led to increased standing crops of periphyton in deeper secti ons of the stream (Power et al . 1985).

In some cases, spring systems may exhibit ve ry complex trophic cascades, which involve reciprocal interactions among pred ators and prey that result in ecological changes across trophic links. Manipulations of lakes demonstrated that such effects remain stable across a range of conditions (Carpenter et al . 1987; Carpenter et al . 2001). For example, a reference lake was characterized by variable productivity of phytoplankt on related to the effects of climatic factors and the natural dynamics of food webs. Manipulating numbers of zooplanktivores and piscivores in other lakes yielded consistent changes in zooplankton assemblages and their grazing, which led to altered phytoplankton assemblages and pr oductivity. For example, gelatinous, colonial green algae became more abundant in a lake with reduced feeding on zooplankton. Additions of significant quantities of nitrogen and phosphorus di d not disrupt the trophic cascade. Piscivorous fishes reduced numbers of zoopl anktivorous fishes to an exte nt that allowed zooplankton numbers to increase. The re sulting increased gr azing by zooplankton controlled the phytoplankton assemblage.

Over time, the interactions among bottom-up a nd top-down effects can change. For example, nutrient additions to a river led to increased primary production in years 1 and 2, with increased production of insects and fish lagg ing until years 3 and 4 (Peterson et al . 1993). Competition among insects moderated the flow of increased primary production through the trophic web (Peterson et al . 1993).

Given the complexity of potential interactions, it is not surprising that an attempt to correlate standing crop of fish with phos phorus levels in streams led to the conclusion that increases in fish biomass with increased phosphorus were grea ter at lower initial pho sphorus concentrations (Hoyer and Canfield 1991). Other responses of lotic systems to increased nutrients also are likely to be nonlinear, with the potential for thresholds and feedback loops.210 In fact, nonlinear responses, such as thresholds and feedback l oops, characterize one of the most cogent descriptions of eu trophication (Duarte 1995). In the eutrophication progression, interactions among nutrients, plan ts and algae, and grazers comb ine to produce changes that are not linearly correlated with nutrien t concentrations. With large and chronic increases in nutrients, an aquatic system is highly likely to follow a qualitative progression from long-lived, rooted, vascular plants that provide a key structural habitat to sy stems dominated by fast-growing macroalgae and microalgae and potentially to sy stems dominated by phyt oplankton if residence times are long enough. Nonlinearities make the quantitative responses of any given system difficult to predict, without detailed information.

In addition to effects on trophic interactions among primary produ cers, grazers and consumers, nutrient enrichment also may alter interactio ns between parasites a nd hosts. Additions of nutrients to mesocosms led to increased algal production, higher densities of snails, greater production of trematode cercariae, and ultimately, increased infection of frogs yielding deformities that decreased their survival (Johnson et al . 2007). The prevalence of such interactions and their consequences in aquatic systems remains unknown.

Nutrient Effects Operating Through Toxicity

Ammonia, nitrate and other chemicals can have toxic effects at a variety of levels from sublethal effects on the physiol ogy, development, reproduction, behavior or morphology of individual organisms through to lethal effects that operate at the level of populations, assemblages and ecosystems (Mattson et al . 2006; Mattson et al . 2007). In general, toxicological experiments address endpoints related to indivi duals. For example, acute toxicity often is addressed by determining a concentration that kills 50% of the test organisms over a defined time interval, or an LC 50 . Florida law contains a definition of a 96-hour LC 50 , which is a concentration that kills 50% of the test organisms over a 4-da y period (Florida Administrative Code Chapter 62–302.200). Sublethal endpoints can al so be used to yield concentrations that cause a defined effect in a defined time inte rval. Terms for such endpoints include least observable effect level (LOEL), wh ich is the lowest concentration at which a change of interest is observed, or no observable effect level (NOEL), which is the concentration at which a change of interest ceases to be obser ved. In addition, experiments can ex amine chronic toxicity or the effect of longer exposure s to lower concentrations of toxi cant; however, the time needed to conduct such experiments has led to various approach es to extrapolate results of acute tests. The results of laboratory experiments documenting eff ects on individuals typically are extrapolated to predict ecotoxicological effects, or effects at the population, assemblage or ecosystem levels. Effects at higher levels of or ganization may arise from density-mediated changes resulting from the death of organisms or trait-mediated chan ges resulting from subl ethal influences, with extrapolations from laboratory results often relying on untested effects on reproductive success or trophic links. Trait-mediated effects represent a primary focus because they typically occur at lower levels of contamination.

Predictions of ecotoxico logical effects may employ uncertainty factors to account for limitations imposed by testing relatively few individuals of relatively few sp ecies for relatively short periods at concentrations that cause an observable effect (Mattson et al . 2006). For example, an intraspecies uncertainty factor can be used to account for variation in sensitivity among individuals within a species. Th e USEPA recommended an intras pecies uncertainty factor211 of 10 applied to any NOEL especially for endangered species (Mattson et al . 2006). An interspecies uncertainty factor ranging from 1 to 1,000 may be applied to account for uncertainties associated with extrapolating toxi cokinetics (processes determining uptake rates, distribution among tissues, detoxification and elimination) and toxicodynamics (processes determining effects such as enzyme inhibition, enzyme facilitation an d binding to DNA) from one species to another (Mattson et al . 2006). The magnitude of the uncertainty factor can be derived from allometric scaling of body sizes, know n variation in acute and chronic sensitivity, and best professional judgment. Another uncertainty factor is used to translate the results of short-term studies to protection during long-term exposures. This adjustment accounts for the facts that contaminants often reach equili brium levels only after long-term exposure; physiological compensations to toxins may fail af ter longer periods of e xposure; latent, adverse effects may appear after exposure ceases; and longer exposures may a ffect life history stages that are more sensitive than those tested. Larger adjustments are needed when available data come from studies much shorter than the life expectancy of the target organism or much shorter than the time needed to achieve equilibrium concentrati ons in the tissues of the target organism. An uncertainty factor also can be used to predict NOELs from obser ved LOELs, with severe effects calling for a larger factor and unbounded LOELs from steep dose-response curves requiring smaller uncertainty factors than unbounded LOELs from curves with shallow slopes. Whenever they are applied, uncertainty factors lower th e concentration considered acceptable in the environment.

The toxicity of aqueous solutions of ammonia, nitrite and nitrate has received attention primarily due to harmful effects in closed circulating syst ems, such as aquaria or fish hatcheries, and through drinking water for livestock that has been contaminated by urine, manure or fertilizer (Armstrong 1979; Shimura et al . 2002; Mattson et al . 2006). Ammonia tends to be more toxic than nitrite, which is typically mo re toxic than nitrate (Armstrong 1979).

Toxicity of Ammonia

The USEPA has invested significant effort in id entifying water quality criteria for ammonia (e.g., USEPA 1999). Their results provide important guidance for establishi ng criteria related to concentrations of ammonia considered safe in aquatic environments. Some key findings from their analysis of data relate d to ammonia toxicity are:

• toxicity of ammonia can depend on the ioni c composition, pH and temperature of water due to poorly understood mech anisms, with pH dependence strongly suggesting joint toxicity of unionized amm onia and ammonium ions

• ionic composition generates small effects on ammonia speciation in freshwater, and its other effects on toxicity are poorly establishe d; therefore, the ionic composition of water was not considered in the derivations of criteria

• temperature can strongly affect the relative amounts of unionized ammonia and ammonium ion, but its effect on the toxicity of ammonia is not strongly indicative of joint toxicity so it was addressed stri ctly by an empirical approach212 • pH effects were described by equations that incl ude basic features of joint toxicity due to unionized ammonia and ammonium ion, along with an empirical component that addressed the incomplete knowledge of these effects

Detailed and updated analysis of acu te toxicity data yielded criter ia expected to protect aquatic life, except for unusually sensi tive species. Criterion maximum c oncentrations (CMCs) drawn from acute toxicity tests address short-term exposure, and analysis of chronic toxic ity data yields criterion continuous concentra tions (CCCs) that address long- term exposure. The ultimate efficacy of the proposed criteria will be known only after monitoring. For example, chronic exposures to ammonia showed that a mayfly species in New Zeala nd would be protected minimally by the proposed criteria (Hickey et al . 1999). In Florida springs, the effects on obligate residents in springs and stygobiota represen t special concerns. For example, a survey of stygobitic crayfish linked their pr esence or absence to ammonia c oncentrations in source water (Spanjer and Cipollini 2006). Regardless of the asso ciated uncertainties, the proposed criteria are based on detailed interpretation of the best avai lable information. Given temperature (T) and pH of a waterbody, the proposed criteria are: 1. The one-hour average concentration of total ammonia nitrogen (in mg–N L -1 ) does not exceed, more than once every three years on the average, the CMC (acute criterion) calculated as:

0.275 39.0 CMC = 1 + 10 7.204 – pH + 1 + 10 pH – 7.204 where salmonid fish are present

0.411 58.4 CMC = 1 + 10 7.204 – pH + 1 + 10 pH – 7.204 where salmonid fish are not present

2A. The thirty-day average concentratio n of total ammonia nitrogen (in mg–N L -1 ) does not exceed, more than once every three years on the average, the CCC (chronic criterion) calculated as:

0.577 2.487 CCC = 

# ( 1 + 10 7.688– pH + 1 + 10 pH – 7.688 

# )

· Minimum of (2.85, 1.45 

### · 10 0.028 · (25 – T) )
where early life history stages of fish are present

0.577 2.487 CCC = 

# ( 1 + 10 7.688– pH + 1 + 10 pH – 7.688 

# )

· 1.45 

### · 10 0.028 · (25 – Maximum of (T, 7))

where early life history stages of fish are not present

2B. In addition, the highest, four-day averag e concentration of total ammonia nitrogen (in mg–N L -1 ) within the 30-day period should not exceed 2.5 times the CCC.213 Toxicity of Nitrate

Nitrate (NO 3 – ) and nitrite (NO 2 – ) coexist as the two major forms of oxidized nitrogen in aquatic systems. Nitrifying bacteria produce nitrite from organic forms of nitrogen and from ammonia as part of the nitrogen cycle (Lewis and Morris 1986). In ae robic environm ents, nitrite concentrations tend to be low because nitrite is c onverted rapidly to nitrate, a more stable form (Guillette and Edwards 2005). Analyses of nitrat e in aquatic systems typically measure both nitrate and nitrite (Edwards et al . 2004). The solubility of nitrate means that any nitrate not taken up by plants or converted to nitr ogen gas by denitrification is tr ansported in surface water and groundwater.

Like all chemicals, the toxicity of nitrate and nitrite is mediated by its uptake and metabolism. Nitrate and nitrite enter animals e ither through permeable membranes, such as the gills of fish or tadpoles and the skin of adult amphibians, or through the gut wall after ingestion, which is the primary means for birds and mammals. It appears that the chloride transport mechanism of gill membranes actively transports nitrate (Jensen 2003; Lewis a nd Morris 1986). Nitrate in body fluids can be converted to nitrit e or it can accumulate because of hepatic detoxification of nitrite (Edwards et al . 2004).

Several detrimental effects have been ascribed to nitrate. Concerns rela ted to drinking water, including groundwater, have cente red on nitrate interf erence with hemoglobin in the bloodstream resulting in a reduced ability to carry oxygen that leads to tissue hy poxia and cyanosis. In vertebrate animals, nitrate oxidizes the iron in hemoglobin molecules disrupting the reversible binding of oxygen and causing methemoglobinemia (Lewis and Morris 1986) . In human infants, the resulting cyanosis is one cause of blue baby syndrome. Nitrate also can act as an endocrine disruptor, interfere with steroid hormone synthe sis, affect sperm motility and viability, affect fecundity, or be toxic to embryos (Edwards et al . 2004; Edwards 2005; Guillette and Edwards 2005; Edwards et al . 2006a; Edwards and Guillette 2007). In fact, nitrate levels in some freshwater systems may reach and be sustained at levels that lead to declines in amphibians (Rouse et al . 1999).

Nitrate toxicity represents a concern in Florida springs because nitrate le vels have risen during the past decades (Mattson et al . 2006; Mattson et al . 2007). Estimates of nitrate toxicity are available for a variety of organisms (Table 3; Camargo et al . 2005; Corrao et al . 2006; Mattson

et al . 2007 and references therein). The bulk of the tests have been conducted on fishes, amphibians and aquatic invertebrates, includ ing crustaceans, insects, and mollusks.214 Table 3. Nitrate concentrations yielding 50% mo rtality in 96-hour tests (96-hour LC 50 ) or other changes deemed relevant, e.g., altered grow th, time to metamorphosis or reproductive characteristics (Camargo et al . 2005; Corrao et al . 2006; Mattson et al . 2007).

Taxonomy Life stage 96-hour LC 50 (mg L –1 NO 3 –N) Other threshold (mg L –1 NO 3 –N) Order: Ephemeroptera

Baetis vernus Nymph 0.73 Potamanthus luteus Nymph 0.70 Baetis fuscatus Nymph 0.74 Cloeon bifidum Nymph 0.72 Ephemerella lenoki Nymph 0.74 Heptagenia sulphurea Nymph 0.70 Order: Trichoptera

Cheumatopsyche pettiti Early instar larvae 113.5 2.4 Last instar larvae 165.5 3.5 Hydropsyche exocellata Last instar larvae 269.5 11.9 Hydropsyche occidentalis Early instar larvae 97.3 1.4 Last instar larvae 109 2.2 Order: Cladocera

Ceriodaphnia dubia Neonates 7.1–56.5 Daphnia magna Neonates 358 Daphnia sp. Adults 34.8–87.1 Order: Amphipoda

Eulimnogammarus toletanus Adults 85.0 4.4 Eulimnogammarus echinosetosus Adults 62.5 2.8 Order: Decapoda

Cherax quadricarinatus Juveniles 1,000 Astacus astacus Juveniles 14 Penaeus monodon (marine) Protozoea 0.23 Macrobrachium rosenbergii Juveniles 160 Juveniles 175 Class: Gastropoda

Pomacea paludosa Adults & juveniles > 500 > 500 Potamopyrgus antipodarum Adults 1,042 195 Family: Salmonidae

Oncorhynchus mykiss Eggs 1.1 Fry 1.1 - 4.5 Fry 2.3 Fingerlings 1,355 14 Oncorhynchus kisutch Eggs 4.5 Fry 4.5 Oncorhynchus tshawytscha Eggs 4.5 Fry 2.3 & 4.5 Fingerlings 1,310 Oncorhynchus clarki Eggs 2.3 & 4.5 Fry 4.5 & 7.6 Salvelinus namaycush Swim-up fry 1,121.4 Embryo 100 Alevin 2,342.5 400 Embryo to swim-up fry 1.6 Coregonus clupeaformis Swim-up fry 1,902.7 Embryo 25 Alevin 2,185.7 25 Embryo to swim-up fry 6.25215 Taxonomy Life stage 96-hour LC 50 (mg L –1 NO 3 –N) Other threshold (mg L –1 NO 3 –N) Family: Moronidae

Morone sp. hybrid Unknown (juveniles?) 200 Family: Centrarchidae

Lepomis macrochirus Fingerlings 1,975 761 Micropterus salmoides Fingerlings? 140 Micropterus treculi Fingerlings 1,261 Family: Ictaluridae

Ictalurus punctatus 1,355–1,423 90 Family: Cyprinidae

Pimephales promelas Larvae 1,010–1,607 358 Catla catla Fry 1,484 Family: Poeciliidae

Gambusia affinis Embryos 4 Adults 4 Poecilia reticulatus Fry 191 Family: Adrianichthyidae

Oryzias latipes Spawning 30 Juveniles 50 Family: Bufonidae

Bufo bufo Tadpoles 384.8 22.6 Bufo americanus Tadpoles 13.6–39.3 Fertilized eggs 9.0 Bufo terrestris Tadpoles 30.0 Tadpoles 4.9 Xenopus laevis Embryos 24.8 Tadpoles 1,655.8 65.6 Family: Hylidae

Hyla versicolor 20 Litoria caerulea Tadpoles 22.7 Pseudacris triseriata Tadpoles 17 10.0 Pseudacris regilla Embryos 643 56.7 Tadpoles 1,749.8 30.1 Family: Ranidae

Rana pipiens Tadpoles 22.96 10.0 & 30.0 Rana clamitans Tadpoles 32.4 Rana sylvatica Fertilized eggs 9.0 Rana pretiosa Newly-hatched larvae 16.5 Rana temporaria Larvae 5 Rana aurora Embryos 29.0 Family: Ambystomatidae

Ambystoma maculatum Fertilized eggs 9.0 Ambystoma jeffersonianum Fertilized eggs 9.0 Ambystoma gracile Newly-hatched larvae 23.4 Family: Salamandridae

Triturus vulgaris Larvae 17.4 Order: Crocodilia

Alligator mississippiensis Juveniles 0.01–0.04216 Nitrate concentrations from 0.23 mg L –1 NO 3 –N to over 1,000 mg L –1 NO 3 –N produced effects on invertebrates. Lethal concentrations in 96-hour tests ranged from 62.5 mg L –1 for an adult amphipod to over 1,000 mg L –1 for an adult snail. The lowest concentration causing an effect was 0.23 mg L –1 , which was reported for a larval stage of a marine shrimp. The lowest levels recorded or calculated to be of import for freshwater species were 0.70 mg L –1 NO 3 –N for nymphs of mayfly species when they were subjecte d to a mixture of nitrat e, nitrite and ammonia. A value of 2.2 mg L –1 was calculated as a safe concentratio n for long-term expos ure of caddisfly larvae to nitrate alone. In general, early life history stages appear to be most sensitive, and freshwater invertebrates appear to be more se nsitive than marine invertebrates. Thus, nitrate toxicity may have its most severe ecotoxic ological effect on inve rtebrate growth and development rates, due to great er effects on larval phases.

Toxicity values for the effects of nitrate on fishes range from 1.1 mg L –1 NO 3 –N for least observable and no observable eff ects on eggs to over 1,000 mg L –1 NO 3 –N for effects on early life history stages. Channel catfish ( Ictalurus punctatus ), largemouth bass ( Micropterus salmoides ), bluegill ( Lepomis macrochirus ), and mosquitofish ( Gambusia holbrooki ) all exhibited responses to nitrate in laboratory e xperiments or abnormalities correlated with nitrate concentrations in freshwater habitats. Impact s on eggs, embryos or fr y may represent special concerns. Responses may vary with environmental conditions and exposure. For example, increased salinity may buffer the toxic effects of nitrate and nitrite, an d longer-term exposures may produce increased tolerance (Knepp and Arkin 1973; Russo et al . 1981).

Amphibians have been promoted as sentinel specie s or species that will provide an early warning of the ecological effects of toxicants. They suit th is role because their skin increases uptake rates for contaminants as a byproduct of being adapte d for respiration and hyd ration and because they have complex life histories spanning aquatic and terrestrial habitats. Negative effects of nitrate on amphibians include decreased activity, re duced feeding, disequilibrium, deformities, mortality, and altered timing of metamorphosis. Evid ence points to nitrate acting as an endocrine disruptor. Overall, sensitivity to nitrate va ries among amphibian species, with 13.6 mg L –1 to over 1,000 mg L –1 being reported to cause mortality in embryos over 96 hours and 4.9 mg L –1 to 65 mg L –1 reported to cause other eff ects in tadpoles. Factors such as water chemistry and the type of nitrogen compound introducing the nitrat e may influence toxicity. For example, tadpoles of Bufo terrestris , the southern toad, raised at nitrate levels of 30 mg L –1 took less time than controls to reach metamorphosis in water prep ared by reverse osmosis and about 6 days longer than controls in unfiltered spring water (Edwards et al . 2006b). Furthermore, tests using ammonium nitrate may yield results that conf ound ammonia and nitrate toxicities. In addition, chronic, sublethal effects may not appear within the 96-hour time limit used in many assays. The role of nitrate toxicity in the decline of amphibian popul ations is confounded by habitat destruction and the impacts of in troduced species; however, nitrat e toxicity should factor into amphibian conservation.

Toxicity tests using birds and most mammals are lacking, and one study a ddressed nitr ate effects on the reptile, Alligator mississippiensis . Preliminary findings indicate that nitrate concentrations between 0.01–0.04 mg L –1 NO 3 –N appear to affect alligators (Guillette and Edwards 2005). Plasma testosterone levels in both male and fema le juvenile alligators were negatively correlated with nitrate levels, and estradiol levels in juven ile females were positively correlated with nitrate levels. Toxicity to birds and mammals probably o ccurs through ingestion of contaminated water.217 Herds of cattle were unaffected by exposur e to water contaminated with 0–44 mg L –1 NO 3 –N; were at risk if continually exposed to 133–220 mg L –1 ; and exhibited losses at 221–660 mg L –1

, more losses at 661–880 mg L –1 , and heavy losses above 880 mg L –1 (http://www.vetmed.ucdavis.edu/vetext/INF-B E_cca/INF-BE_cca01/INF-BE_cca0111.html). In addition, the human health standard for nitrate in drinking water is 10 mg L –1 (http://www.dep.state.fl.us/water /drinkingwater/standard.htm).

Average nitrate levels in 92% of 130 Florida springs were reported to be below 2 mg L -1 ; therefore, they were typically lower than conc entrations reported to cause effects on animals (Munch et al. 2006). However, nitrate concentrations ex ceed levels of concern in some springs, and they may reach levels of concern for short periods in other springs, especially after the application of safety factors. Given the uncerta inty surrounding estimates of toxicity, nitrate concentrations in Florida’s springs do represent a cause for concern.

Summary of Nutrient Effects on Fauna in Springs

In summary, a synthesis of information indicates that nutrients can affect faunal assemblages in aquatic systems. Effects are generated in three primary ways: 1) toxicit y, 2) changes to trophic webs, and 3) changes in habitats.

Toxicity from elevated ammonia concentrations and changes to respiratory functioning from elevated nitrate concentrations are well-established facts. In addition, evidence of other detrimental effects from elevated nitrate c oncentrations is accumulating. Increases in concentrations of nitrate may initially lead to changes in traits of i ndividual organisms. These sublethal effects on individuals could translate to detrimental effects on populations, assemblages or ecosystems, although we lack definitive data on these outcomes. Concentrations of ammonia and nitrate in most Florida springs have not r eached toxic levels, but concentrations are of concern in some springs. The ecotoxicological effects of nutrients should factor into management of nutrients in Florida’s springs.

Increases in concentrations of nutrients also can change trophic webs and habitats, with outcomes determined by complex interactions among bottom-up and top-down processes. At this time, we lack the data needed to rigorously evaluate the role of nutrients in creating perceived and observed changes in the habitats or trophic webs of Florida springs.

Overall, management of nutrients in spring ecos ystems would benefit from diagnostic studies of bottom-up and top-down interactions. These interac tions, with their associated complexity and nonlinearity, ultimately determine the level of management needed to protect springs ecosystems. Furthermore, the complex and nonlinea r nature of key interactions make it highly likely that management applied at real world, spat iotemporal scales will generate unforeseen and unforeseeable responses or emergent properties, which create a need to manage adaptively.218 RESEARCH TO IMPROVE MANAGEMEN T OF NUTRIENTS IN SPRINGS

Florida’s springs and spring runs are enriched in nitrate due to broad-scale contamination of groundwater supplies. As a resu lt, the potential for eutrophication becomes a major concern. The eutrophication progression scheme suggests that in creased nutrient delivery to aquatic systems favors production of fast-growing algae that ulti mately outcompete and di splace native vascular plants (Duarte 1995).

In some of Florida’s spring systems, both the a pparent proliferation of nuisance algae and the apparent decline of native macr ophytes are consistent with such a progression. However, we have few quantitative data that definitively link these changes to an increase in nitrate delivery, and the eutrophication progression scheme predicts that nonlinear responses in systems may have “decoupled” them from contemporary concentrations of nutrients in the water column. Currently, patterns in the distribution and abundance of plan ts and algae do correlate with other physical and chemical parameters. However, we know little a bout their individual or combined effects or the form and strength of poten tial feedback mechanisms, incl uding those linked to nutrient concentrations. Overall, numerous processes and their interactions probabl y dictate the structure and function of submersed aquatic vegetation, and changes in these vegetated communities generate effects on faunal assemblages.

In general, nutrients can affect the faunal assemblages in aquatic systems in three primary ways: 1) toxicity, 2) changes to troph ic webs, and 3) changes in habi tats. Toxicity from elevated concentrations of ammonia and nitrate is not a majo r concern in most Florida springs at this time. In addition, sublethal effects of nitrate on individuals may not impact populations, assemblages or ecosystems. Regardless, the ecotoxicological effects of nutrients should factor into management of nutrients. Changes in trophic we bs and habitats, along with flow-on effects on faunal assemblages, are determined by comple x interactions among bottom-up and top-down processes. At this time, we lack data to support rigorous evaluation of th e role of nutrients in such changes.

Overall, management of nutrients in spring ecosystems would benefit from compilation of inventories, collection of baseli ne data, and completion of diagnostic studies focused on bottomup and top-down interactions. Th e interactions, complexities and nonlinearities uncovered by these studies should influence the management of spring systems. Furthermore, the likelihood that unpredictable, emergent properties will ar ise when management is implemented at real world, spatiotemporal scales creates a need to manage adaptively.

Designing, implementing, evaluating and applyi ng research that improves management of Florida’s springs and spring-fed systems require s input from multiple stakeholders, including policy makers, managers, the public, and scientis ts from many disciplines. Initial targets for discussion include:

• inventories and baselines designed to:

− measure nutrients, oxygen, carbon, light, flow , substrates, grazing, algae, plants, invertebrates, fish, and other key driver s and valued components of spring systems chosen with the best availa ble information and, eventually, with guidance from the results of diagnostic studies219 − target systems where effectiveness of management can be assessed (e.g., systems with total maximum daily loads, minimum flows and levels, pollution load reduction goals, surface water improvement plans or basin management action plans)

− sample synoptically or at spatial and tem poral scales that su pport rigorous analyses of interactions (note that these inventor ies and baselines form the platform for adaptive management)

− balance breadth with depth (e.g., opti mize taxonomic detail and replication)

• diagnostic studies that determine assimilation of nutrients by flora as key elements in nutrient budgets for spring systems, includi ng contributions by epiphytes, macroalgae, vascular plants and microphytobentho s or microalgae found in sediments

• diagnostic studies that determine how nutrien ts and other drivers affect overgrowth, shading and other relationships among periphyt on, macroalgae and vascular plants in an effort to determine the need for management actions, including aquatic plant management and restoration of native species

• diagnostic studies designed to:

− focus on species of algae and plants that are hypothesized or known to play major roles in nutrient assimilation

− extend beyond nitrogen and phosphorus to in clude other potentially limiting factors and interactions among limiting factors that can significantly influence the outcomes of management (e.g., micronutrients, iron, oxygen, carbon, light, flow and substrate type)

− explore relationships and consequences across multiple levels of biological and ecological complexity, includi ng uptake and other physiolo gical responses, growth and reproduction of individual organisms, and ecological changes in populations or assemblages to promote the success of management actions at the system level (note that some of the key, large-scale relatio nships form the core of an adaptive management approach)

− elucidate the direction, magnitude and form of limitations (e.g., nonlinear and nonreversible) to identify suitable targ ets and goals for management actions

• diagnostic studies to identify and elucidate changes in comp osition and function of faunal assemblages related to changes in habitats in an effort to identify habita ts to be protected or restored by management actions

• diagnostic studies to determine palatability of flora to grazers, rates of grazing, and the form and magnitude of predation by primary c onsumers in an effort to identify flora, grazers and trophic links to be protected or restored by management actions

• diagnostic studies to identify and elucidate th e toxic effects of ammonia and nitrate beyond physiological and individual levels of or ganization (e.g., population and assemblage effects) in an effort to determine thresh old levels, safety fact ors and the need for management actions220 REFERENCES

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# 5

Ecosystem Implications of Invasive Aquatic Plants and Aquatic Plant Control in Florida Springs

Jason M. Evans School of Forest Resources and Conservation University of Florida

Summary Nonnative and nuisance plants such as water hy acinth, water lettuce, and hydrilla are a primary management concern in many Florida springs. In fiscal year 2005 – 2006, the Department of Environmental Protection’s Bureau of Invasi ve Plant Management spent approximately $173,000 to control these plants in springs, primar ily through the use of chemical herbicides. This chapter reviews a broad range of literatu re to outline what is known and unknown about 1) the history of these plants and their control in Florida; 2) th e growth potential of the nonnative plants in springs as a function of elevated ni trate-nitrogen concentrati ons; 3) the social and ecological consequences of aqua tic plant overgrowth; 4) the eco logical risks associated with current aquatic plant control methods; and 5) the potential benefits of a lternative aquatic plant management approaches in some springs systems.

The Literature: What is Known

• Major problems with nonnative plants in Flor ida began with the introduction of water hyacinth, a floating aquatic plant, into the St. Johns River in the late 19 th century. Water hyacinth was documented in several springs ecosystem along the St. Johns River by the mid 1890s. Chemical control programs have maintained water hyacinth populations at low levels throughout Florida since the mid-1970s.

• Historical sightings by William Bartram indi cate that water lettuce, a floating aquatic plant, has been present in a number of Fl orida springs since at least 1765. Scientists disagree as to whether water lettuce was presen t in Florida before Eu ropean colonization, or was introduced by early Spanish settlers. Chemical control programs have maintained water lettuce at low levels thr oughout Florida since the mid-1970s.

• Hydrilla, a submersed aquatic plant, became established in severa l areas of Florida, including the Kings Bay/Crys tal River springs complex, by 1960. Sustained control of hydrilla has proven more difficult than the floating plants in Florida. Most aquatic plant management costs in Florida springs ecosy stems over 2005 – 2006 were associated with chemical control of hydrilla.

• There are clear relationships between nitrogen enrichment and increased growth of water hyacinth, water lettuce, and hydrilla in non -flowing aquatic systems. Such overgrowth232 can have severe adverse effects on native plant communities, navigation, fisheries, and recreational desirability.

• Water hyacinth and water lettuce emit allelopa thic compounds capable of suppressing a number of algal taxa. Water hyacinth, water lettuce, and hydrilla all can be effective competitors with nuisance algae due to nutrien t uptake, shading, faunal habitat, direct filtration, and allelopathic mechanisms.

• Ecosystem surveys indicate that water hya cinth, water lettuce, and hydrilla provide attractive habitat for crayfish, apple snails , amphipods, fish, manatees, and other springs fauna at moderate levels of coverage.

• Copper and diquat herbicides may have signi ficant toxicological effects on algal and faunal community dynamics at levels used for aquatic plant control.

• Lyngbya wollei , a filamentous cyanobacterium of great concern in many Florida springs, is notable for its relative resistance to herbicidal compounds as compared to other common algal and cyanobacteria taxa.

• Depression of dissolved oxygen due to decaying biomass is a primary concern to animals following aquatic plant control.

• Biotypes of hydrilla that are resistant to fl uridone, a systemic he rbicide commonly used for hydrilla control in Florida lakes, have been documented in recent years. There is also increased concern about the potential evolution of hydrilla strains that are resistant to Aquathol, the contact herbicide most commonl y used to control hydrilla in springs.

• Established biological control organisms are known to adversely affect water hyacinth, water lettuce, and hydrilla in some spring s systems, although biocontrol organisms generally do not maintain a quatic plant populations at low levels achieved by chemical control. A promising biol ogical control for hydrilla, Cricotopus lebetis , has been documented in Kings Bay/Crystal River.

The Literature: What is Not Known, and Recommendations for Future Research

• The few studies available for springs and ot her flowing waters ha ve not definitively determined a concentration of nitrate-nitrogen in springs that would be limiting to any of the nonnative plants. Nutrient assays in flowing water mesocosms and/or in situ field studies are needed to develop nitrate-nitrog en limitation values for nonnative plants in springs conditions.

• Observations from several springs suggest a “boom-bust” successional sequence in which nonnative plants first out-compete native plant communities, and then suffer catastrophic population crashes associated wi th aquatic plant control or natural disturbances. It is hypothesized, but not known, that succession of springs into algal-dominated ecosystem states may be promoted by the nutrient pulse s and ecological opening s associated with the rapid loss of aquatic plant populations.233 • The effects of allelopathic compounds em itted by nonnative plants on algal dynamics in springs ecosystem are not presently known.

• Ecosystem effects of long-term aquatic plant control have not been well-studied in the specific context of Florida springs. Variable s such as dissolved nutrients, dissolved oxygen, biomass deposition, and floral and faunal community structure should be monitored before and after major aquatic plan t control operations in Florida springs.

• Water hyacinth and water lettuce are currentl y being managed for algal-suppression, nutrient recovery, and biomass utilization in a number of tropical countries, including places in which they are considered nonnativ e. While it is not know n if such methods would be helpful in springs ecosystems, car eful experimentation with floating plant treatment methods may be worthwhile in hi ghly degraded spring s ecosystems where these plants are established.

• Long-term effects of biocontrol organisms, including the possibility of improved efficacy over time, are not well-studied in nonnative plant populations in springs ecosystems. Increased research into biocontrol organi sms and experimental release in springs ecosystems, particularly those affected by hydrilla, should be a priority for adaptive ecosystem management.234 INTRODUCTION

One of the primary management concerns in many of Florida’s freshwater spring systems is the growth of nonnative invasive pl ants. Although there are more than a dozen nonnative aquatic plants established in Fl orida’s springs ecosystems, the vast majority of hi storic and ongoing management expense is associated with three species: water hyacinth ( Eichhornia crassipes ), hydrilla ( Hydrilla verticillata ), and water lettuce ( Pistia stratiotes ) (Bureau of Invasive Plant Management 2007). Many researchers and managers fe ar that rising nitrate- nitrogen levels in Florida springs may further promot e growth of these, and perhap s other, nonnative plants to the overall detriment of native plan ts and animals (e.g., Florida Springs Task Force 2000; Loper et al. 2005).

Aquatic plant control programs that primarily targ et hydrilla, water hyacint h, and/or water lettuce are actively employed in several major springs sy stems, including Rainbow River, Silver River, Wekiva River, Weeki Wachee River, Wakulla Ri ver, and Kings Bay/Crys tal River (Bureau of Invasive Plant Management 2007). Selective applic ation of chemical herbicides is the primary operational method for aquatic pl ant control, although manual a nd mechanical harvest methods are also used in some springs systems. Approximately $173,000 was spent by the Florida Department of Environmental Protection (DEP) fo r aquatic plant control in springs ecosystems in 2005 – 2006. Over 80% of this amount was allo cated for herbicidal suppression of hydrilla in Rainbow River, Weeki Wachee River, Wakulla Ri ver, and Silver River (Bureau of Invasive Plant Management 2007).

This review chapter begins by generally desc ribing the introduction, eco logical effects, and historical control of the major invas ive aquatic plant species in Florida. Available scientific literature is then used to explore four sets of questions raised by the presence of invasive plants and their ongoing management in Florida springs:

1. To what extent is increased nitrate-nitrog en a factor in the spread of invasive plants, and at what level is invasive plant growth limited by nitrate-nitrogen in springs?

2. What are the potential implications of invasi ve plant proliferation on ecosystem structure and func tion in springs?

3. What are the potential implications of aquatic plant control on ecosystem structure and function, particularly in term s of shifting springs toward more algaldominated ecosystem?

4. For springs in which increased algal dominance is the primary conservation concern, are there opportuniti es for experimenting with alternative a quatic plant management strategies that attempt to utilize the known functions of extant invasive plants in support of l ong-term ecosystem recovery goals?235 Major Invasive Aquatic Plants: HIstory and Control in Florida

Water Hyacinth

Water hyacinth is a floating aqua tic plant native to South America’s Amazon River drainage. Introduced into many countries throughout the subtropics and tropics in the late 19 th and early 20 th centuries, water hyacinth is now commonly regard ed as one of the world’s most problematic invasive plant species (Holm et al. 1977). A number of researcher s have found that, under ideal conditions, water hyacinth is mo re productive than any other known vascular plant (Gopal 1987). This extremely high productive potential permits water hyacinth to quickly overgrow and outcompete other plant species (Schmitz et al. 1993), particularly in ecosystems affected by elevated nutrient levels and other forms of anthropogenic disturbance (Gopal 1987).

The ecological and economic effects of water hyacinth overgrowth are often quite dramatic and severe. Along with the competitive displacement of native plant communities, one of the most common ecological consequences of water hyacinth overgrowth is severe depression of dissolved oxygen in underlying wa ters due to the syne rgistic effects of rapid deposition of organic matter from senescing leaves, restric tion of phytoplankton produc tion due to shading, and interference with atmospheric oxygen ex change (Penfound and Earle 1948; Joyce 1985). Development of anoxic conditions under water hyacint h can directly lead to exclusion of fish and other oxygen-dependent species, thereby radically changing the structural composition of faunal communities (Schmitz et al. 1993). In addition, human economies are often adversely affected by restricted navigation, loss of fisheries access, and siltation of drainage systems that can be associated with the formation of substantial water hyacinth co ver in a water body (Gopal 1987).

It is generally accepted that water hyacinth was first introduced into Florida in 1885 by a Palatka-area farmer who obtained a specime n from an exhibit at the 1884-1885 World’s Industrial and Cotton Centennial Exposition held in New Orleans (Buker 1982). The plant was apparently propagated in a small garden pond, with excess plants periodically thrown into the nearby St. Johns River (Buker 1982). The subsequent effects of this introduction are legendary. By the early 1890s, the discarded plants had multip lied to such an extent that extremely large floating mats of water hyacinth became establis hed at many locations along the St. Johns River and its tributaries (see Figure 1). Boat naviga tion and logging operation s along the river flood plain were both severely curtailed, and massive accumulations of the plant were even reported to structurally undermine the columns of several railroad bridges (Webber 1897).

Aided in large part by farmers who valued the prolific plant for cattl e forage (Buker 1982) and an organic mulch source for citrus trees pl anted in sandy soils (Ma ltby 1963), water hyacinth continued to spread rapidly into many of the state’s other draina ge basins throughout the first half of the 20 th century. Water hyacinth was reported in several springs of the St. Johns River basin, including Volusia Blue Spring, Green Cove Springs, Silver River, and the Wekiva River, as early as the 1890s (Webber 1897), and introductions into new springs systems continued throughout the 20 th century. Reports from the Panhandle’ s spring-fed St. Marks River suggest that water hyacinth was first intr oduced into that system as la te as the 1970s (Bartodziej and Leslie 1998).236 The history and evolution of effo rts to bring water hyacinth under control in Florida are almost as legendary as the plant’s rapid spread. The United States Rivers and Harbors Act of 1899 contained specific provisions aut horizing the United States Army Corps of Engineers to destroy water hyacinth in navigable waters, and the Florid a Legislature at nearly the same time enacted a law that prohibited the intentional transfer of water hyacinth into new waterways (Buker 1982). Sodium arsenite and other inorganic herbicid e sprays were briefly used by the Corps of Engineers for hyacinth control ar ound the turn of the century, bu t reports of se vere cattle mortality associated with these sprays led both the Florida Legislature and U.S. Congress to prohibit use of these compounds in Flor ida by 1905 (Buckman and Company 1930).

As a result of this prohibition, most hyacinth cont rol efforts from the turn of the century through the 1940s involved three techniques: 1) construction of physical barriers to prevent movement of plants into new areas; 2) manual labor to disl odge plants and facilitate downstream discharge into the Atlantic Ocean; and 3) mechanical shredding and/or harvest of plants using large machines (Buckman and Company 1930; Buker 1982 see Figure 2). While such methods were successful in terms of removing major plant bloc kages, they apparently were insufficient for abating the proliferation of wate r hyacinth. Thus, the spread and areal extent of water hyacinth coverage continued to incr ease in Florida th roughout the first half of the 20 th century, peaking at approximately 51,000 acres in the early 1960s (Schardt 1997; see Figure 3).

Figure 1 – Water hyacinth overgrowth in St. Jo hns River. Illustration from March 19, 1898 edition of Harper’s Weekly, as adapted by Un iversity of Florida Ce nter for Aquatic and Invasive Plants. http://aquat1.ifas.ufl.edu/guide/stboatbg2.gif237 By all accounts, the inven tion of the organic phenoxy herb icide 2,4-D in 1942 marked an important turning point in water hyacinth control. Unlike the inor ganic herbicidal sprays used previously, 2,4-D effectively cont rolled water hyacinth at dosages that posed little direct toxic risk to cows, fish, and other animals (Joyce 1982), likely because its mode of herbicidal action specifically targets plant horm onal production. A number of 2,4-D- based spraying programs that targeted heavily concentrated populations of wa ter hyacinth in Florida commenced in the late 1940s and continued to expand throughout the 1950s and 1960s (Schardt 1997). While these early 2,4-D programs were successful in rapidly suppressing water hyacinths, many observers expressed concerns about mass nutrient releases, heavy organic detritus loads, and severe algal blooms that tended to follow larg e-scale aquatic plant c ontrol operations (e.g., Clugston 1963; Tilghman 1963). Over the late 1960s and early 1970s, aquatic plant control researchers and practitioners addressed these concerns by developing a management philosophy that came to be known as maintenance control (S chardt 1997). Adopted in to state law in 1974 as the official guiding principle fo r aquatic plant management, main tenance control is defined by the operational goal of maintaini ng water hyacinths (and other i nvasive nonnative aquatic plants)

Figure 2 – Water hyacinth shredder in the Withlac oochee River, circa 1940. Photo by United States Army Corps of Engineers, as adapted by the University of Flor ida Center for Aquatic and Invasi v e Plants. http://plants.ifas.ufl.edu/guide/mechconold7js.jp g238 at the “lowest feasible level” (Florida Statut e 369.22). Inherent in this objective is a pragmatic acknowledgment that eradication of the invasive aquatic plants is impractical, but that control at population numbers that do not interf ere with beneficial uses of water bodies can be maintained. Aside from the obvious objective of effectively suppressing nonnati ve plants, some research suggests that maintenance of plants at small popul ations may reduce overall detrital loading and herbicide use as compared to less frequent treatment of large plant populations (Joyce 1985; Schardt 1997). Since the 1970s the maintenance c ontrol program has greatly reduced the overall population of water hyacinth throughout the state (see Figure 3), generally through sustained use of 2,4-D, diquat dibromide (a contact herbicide), and glyphosate (a system ic herbicide) (Bureau of Invasive Plant Management 2007). Potential im plications of herbicidal control methods used for water hyacinths and other invasive plants in springs ecosystems are discussed in more detail in later sections of this chapter.

Figure 3 – Effects of maintenance control operati ons on water hyacinth populations in Florida.Graph adapted from Florida Departmen t of Environmental Protection, Bureau of Invasive Plant Management. http://www.dep.state.fl.us/lands/inva spec/images/Graph%20wh%20acres.jpg239 Water Lettuce

Like water hyacinth, water lettuce is a floating aquatic plant that can commonly reach nuisance levels in Florida and other areas of the tropics and subtropics (Holm et al. 1977). The basic economic and ecological concerns about water le ttuce are quite similar to those of water hyacinth, namely that its prolific floating growth habit can serve as an impediment to navigation, block drainage canals, out-compete native plants, and radically change aquatic communities through increased siltation and depression of dissolved oxygen concentrations (Schmitz et al. 1993).

Unlike water hyacinth, there is significant uncertain ty as to the origins of water lettuce and its introduction to Florida. Most of this uncertainty derives from William Bartram’s frequent reports of water lettuce along the Suwannee River and St. Johns River during his Florida travels in the 18 th century (Stuckey and Les 1984). Some researchers speculate th at water lettuce observed by Bartram possibly stemmed from an introduction of the plant during the Spanish colonial period (e.g., Stuckey and Les 1984), while others argue th at water lettuce’s Florida population may have preceded European contact by millennia (e.g., Stoddard 1989). Bartram’s writings make specific note of large water lettuce populations in or n ear the Suwannee River’s Manatee Springs and several springs along the St. Johns River, indica ting that these springs have had water lettuce populations for well over two centu ries. Other historical account s in springs include Webber’s (1897) sighting of large water lettuce populati ons in the Wekiva River during the late 19 th century, and Carr’s (1994) reports of water lettuce as a common component of springs ecosystems throughout north Flor ida during the 1940s. Field obs ervations by Odum (1957) suggest that water lettuce growth in Silver Springs during the mid 20 th century followed distinct

Figure 4 – Water lettuce in Florida, 1765. Drawing by William Bartram, as adapted by Swarthmore College. http://www.swarthmore.edu/Humanities/ kjohnso1/pictures/stratiotes.jpg240 seasonal patterns and that larg e population accumulations were often suppressed by insect herbivory.

It is commonly suggested that water hyacinth’ s similar habitat requirements and superior competitive abilities led to a larg e-scale displacement of water lettuce in Florida during the late 19 th and early 20 th centuries (Schmitz et al. 1993; Carr 1994). Howeve r, large populations of water lettuce quickly emerged in Florida and other states after the commencement of water hyacinth control in the 1950s, likel y due to water lettuce’s relativ e resistance to 2,4-D (Eggler 1953; United States Army Corps of Engin eers 1973). Researchers during the mid 1960s discovered that water lettuce was controlled effectively by diquat dibromide (Weldon and Blackburn 1967), and since that tim e water lettuce and mixed stands of water lettuce and water hyacinth in Florida have been almost exclusively treated with this herbicide as part of maintenance control programs (Mossler and Langeland 2006). One exception is the spring-fed Ichetucknee River, where a control program ba sed solely on hand removal of plants has drastically reduced water lettuce pop ulations in much of the river over the past several years.

Hydrilla

Hydrilla is a submersed aquatic plant native to Southeast Asia and Africa. Commonly imported into Florida by the aquarium trade during the mid 20 th century, naturalized populations of hydrilla were identified in several Flor ida waterways by the early 1960s (Schmitz et al. 1993). Over subsequent decades the plant has rapidl y spread into many aquatic systems throughout Florida and the southeast United States. Of th e $29 million requested by Florida’s Bureau of Invasive Plant Management (2007) for aquatic plant control in public waters for 2007 – 2008, well over half ($16 million) is marked for ongoing hydrilla suppression. The Bureau of Invasive Plant Management (2007) also reported that approximately $142,000 wa s spent for hydrilla control in spring systems such as Wakulla Ri ver, Weeki Wachee River, Silver River, and Rainbow River in fisc al year 2005 – 2006.

Several features make hydrilla an extremely effe ctive and problematic invasive species. Perhaps most critical to its spread is the viability of even very small plant fragments, which are often inadvertently brought into new water bodies by boat trailers and propellers (Schmitz et al. 1993). Once established in a water body, hydrilla has the ab ility to grow from sediments up to the water surface (often referred to as “topping out”) and form a canopy that can effectively shade out native submersed species, severely restrict navigation, and interfer e with flood control structures (Schmitz et al. 1997; Jones and Beardall 2005). In a ddition, hydrilla’s root system is characterized by many large tubers that often resp rout after leafy growth is suppressed through herbicides or other disturbance (Schardt 1997). Some research suggests that hyd rilla may also have lower light requirements for photosynthesis relative to na tive submersed plants, thereby allowing it to colonize deeper ar eas of water bodies that previous ly would have been free of vascular plant growth (Van et al. 1976). Like the floating species discussed above, large-scale hydrilla coverage can result in the depression of dissolved oxygen levels due to reduced atmospheric diffusion and suppression of phyt oplankton production in underlying waters (Schmitz et al. 1993). Dense hydrilla also has been known to adversely affect sports fisheries by providing increased habitat cover th at restricts prey availability fo r species such as large mouth bass (Bureau of Invasive Plant Management 2007).241 Although the deleterious effects of hydrilla overgrowth are quite dramatic, some research indicates that hydrilla can bene fit certain fish and wildlife populations. For example, it is commonly reported that moderate hydrilla cove rage provides superior game fish habitat (Schmitz et al. 1993), which leads many fishermen to advo cate for management strategies that maintain significant hydrilla coverage in pop ular fishing lakes (Jon es and Beardall 2005). Hydrilla is also known to provide highly attrac tive habitat for manatees (Campbell and Irvine 1977), certain macroinvertebrates (Schramm and Ji rka 1989), and diverse assemblages of native water fowl (e.g., Johnson and Montalbano 1984; Es ler 1990). Thus, one of the most complex challenges currently facing Florida’s aquatic plan t managers is development of techniques that better maintain the positive habitat values of hydrilla in areas where it is permanently established, while still preventi ng severe overgrowth and spread of the plant into new areas (Jones and Beardall 2005).

Modern hydrilla control methods in Florida rely heavily upon fl uridone, a slow-acting systemic herbicide, and endothall, a fast-acting contac t herbicide (Jones and Beardall 2005). Because effective treatment by fluridone requires long c ontact times that are prohibitively difficult and expensive to obtain in flowing sy stems, a dipotassium salt formulation of endothall (trade name Aquathol) is typically used for hydrilla cont rol in springs (DEP 2005). However, serious concerns about the long-term sust ainability of current chemical control methods are raised by recent findings of independently evolved fluri done-resistant hydrilla biot ypes in Florida lakes frequently treated with fluridone (Michel et al. 2004). Although no endothall -resistant strains have been identified, many researchers argue th at future development of resistance among hydrilla populations repeatedly tr eated with endothall, such as those in springs, is possible or perhaps even inevitable (J ones and Beardall 2005).

Figure 5 – Hydrilla in Wakulla Springs, 1998. Photo by Vic Ramey, University of Florida Center for Aquatic and Invasive Plants http://aquat1.ifas. ufl.edu/wakhyd.jpg242 Diquat and copper (including chel ated copper and copper sulfate formulations) are the major herbicides aside from fluridon e and endothall that are curre ntly listed for hydrilla control (Langeland 1996; Bureau of Invasive Plan t Management 2007). Copper and diquat were commonly used for hydrilla cont rol in the spring-fed Kings Bay/Crystal River complex during the 1970s and early 1980s (Haller et al. 1983), but apparently with uneven operational success (Dick 1989). Recent research indicates that copper and diquat may be most effective against hydrilla when applied together and/or in combination with endothall (e.g., Pennington et al. 2001), and such a strategy may be increasingly empl oyed in Florida due to resistance concerns (Jones and Beardall 2005). Howeve r, increased use of copper for hydrilla control in some Florida springs is problematized by regulatory language that places restrictio ns on copper herbicide usage in areas frequented by manatees (State of Florida Division of Administrative Hearings 1993). These restrictions on copper herbicides were a dopted by the Florida DEP in response to the documentation of elevated copper levels in the tissues of Kings Bay/Crystal River manatees (O’Shea 1984) and in the sediments of water bodies, including Kings Ba y/Crystal River, in which copper was commonly used as an herb icide/algaecide throughout the 1970 and 1980s (Leslie 1992). Toxicological concerns associated with chemical control of hydrilla are discussed at greater length in subsequent sections of this chapter.

NITRATE-NITROGEN AND NONNATI VE PLANTS IN SPRINGS

The increased nitrate-nitrogen contamination obser ved in springs throughout Florida is of great concern to ecosystem managers an d the general public. Most direct research into the ecosystem effects of nitrate-nitrogen in springs has focused on filamentous algae such as Lyngbya wollei and Vaucheria sp. (Cowell and Dawes 2004; Stevenson et al. 2004). However, significant concerns have also been expressed about th e potential for nitrate- nitrogen to favor the proliferation of water hyacinth, water lettuce, hydrilla, and other inva sive nonnative plants (Florida Springs Task Force 2000; Loper et al. 2005).

Although there is a fairly large body of literature that describes nutrient uptake of these aquatic plants for the purposes of wastewater treatmen t and other forms of environmental remediation (see Ho and Tsang 1998; Gu 2006), much less direct productivity research has been conducted at nutrient concentrations or condition s relevant to those in Florida springs. As discussed in more detail below, the available litera ture strongly suggests that nitroge n limitation of th ese plants in most Florida springs would only occur, if at all, at the upper bounds of background aquifer nitrate-nitrogen concentrations.

Water Hyacinth and Nitrogen Limitation

A topic explored throughout the vast literature about water hyacinth ecology and control is the growth and uptake responses of water hyacinth to increased loading of nutrients. A common observation is that water hyacinth problems tend to be most serious in waters that suffer from nutrient enrichment, and, thus, nutrient mitiga tion is often recommended as a strategy for reducing plant growth (Gopal 1987). Limitation by either phosphorus or nitrogen is most common, although low levels of calcium, potassi um, and iron have also been found to limit water hyacinth growth (Gopal 1987). While nitrogen limitation to wa ter hyacinths is frequently indicated when N/P ratios are less than 7 (Wilson et al. 2005), experimental trials indicate that, at243 very high nutrient levels, dissolved nitrogen typically will be de pleted by water hyacinth stands at higher rates than phosphorus due to both luxury uptake by plants and denitrification (Reddy and Tucker 1983).

The natural concentrations of phosphorus, calci um, phosphorus and iron in typical Floridan aquifer water (Scott et al. 2004) are of sufficient quantity to make it unlikely that water hyacinth historically would have been li mited by these nutrients in most Florida spring ecosystems. While long-term enrichment trends observed in spring s provide a compelling rationale for considering nitrate-nitrogen the parameter of most concern for water hyacinth growth in Florida springs, the few direct studies of water hyacinths in springs have not definitively determined a concentration of nitrate-nitrogen that might limit biomass production. For example, Bartodziej and Leslie (1998), in a long-term study on the spring fed St. Marks River, re port a water hyacinth biomass doubling time of 10 days at nutrient levels of 0.28 mg/l TN 1 and 0.06 mg/l TP. Because this biomass doubling time is essentially the same as those recorded in Japanese experiments (Sato and Kondo 1981, cited in Bartodziej and Leslie 1998) where water hyacinth was grown under conditions of extremely high nutrient enrichment (28 mg/l TN and 7.7 mg/l TP), Bartodziej and Leslie (1998) conclude that th e observed nutrient levels in the St. Marks River were not a limiting factor for water hyacinth growth. Similarly, Odum’s (1957) measures of water hyacinth productivity in Silver River, which at the time of study in the mid 1950s had a nitrate-nitrogen level of 0.4 mg/l, indicate that nutrie nts were not a major limiting factor.

Two recently developed water hyacinth models pres ent nitrogen concentration values that could be expected to limit water hyacinth gr owth in some aquatic systems. Wilson et al. (2005) develop a synthetic model indicating that water hyacinth growth responds to nitrogen concentrations according to a logarithmic function, with the most dramatic increases in specific growth rate occurring between 0.1 and 1 mg/L of TN. While these n itrogen concentrations clearly fall within a range of di rect relevance to Florida springs, direct inference of these results into springs conditions is confounded by the model’ s stated omission of flow velocity effects on nutrient availabili ty and growth response. A water hyacinth growth model developed by Mahujchariyawong and Ikeda (2001) for Thailand’ s Tha-Chin River suggests that maximum growth rate of the water hyacinth requires 0.16 mg/l TN and 0.02 mg/l TP. Although this model does have the notable advantage of accounting fo r nutrient availability and replacement as affected by flow velocity, inferential caution is warranted due to important morphological, chemical, and climatic differences between th e Tha-Chin River and typical Florida springs systems.

Perhaps the most intriguing clue about potential n itrate-nitrogen thresholds for water hyacinth in Florida springs comes from Webber’s (1897) obse rvation of water hyacinth plants showing a stunted growth form in the Silver River during the late 19 th century. A data record reported by the USGS from 1907 indicates a nitrate-nitrogen concentration of 0.03 mg /l in Silver Springs (Munch et al. 2006). Taken together with the scient ific literature discussed above, it is reasonable to hypothesize that Webber’s (1897) observations may be i ndicative of nitratenitrogen serving as a physiological constraint to water hyacinth grow th in the Silver River at the

1 All nitrogen concentrations reported by Bartodz iej and Leslie (1998) are in terms of TN. Relative contribution of nitr ate-nitrogen to the TN measurement is not given.244 background water quality levels of the late 19 th and early 20 th century. Confirmation of this suggested relationship, however, would require detailed assays of water hyacinth growth response to variable nitrate-nitr ogen concentrations in flowing mesocosms, and/or comparative field studies of water hyacinth growth characteris tics in springs systems, including those with background levels of nitrate-nitrogen.

Water Lettuce and Nitrogen Limitation

The similarity of growth and uptake responses observed for wa ter hyacinth and water lettuce under experimental nutrient trea tments (e.g., Aoi and Hayashi 1996) may make it reasonable to assume that the potential nutrient limitation thres holds for water lettuce growth in Florida springs are similar to those of water hyacinth. For example, it is well-established that water lettuce, like water hyacinth, will remove much higher levels of dissolved nitrogen than are necessary to sustain its maximum growth (A oi and Hayashi 1996). Schmitz et al. (1993) hypothesize, while also noting the lack of specific studies testing this hypothesis, th at increased levels of nitrogen may favor water lettuce to out-compete nativ e submersed plants in freshwater systems throughout Florida.

Unfortunately, there is still very little research suggestive of the nitrate-nitrogen concentrations that might limit water lettuce production in Fl orida springs. Productivity measurements given by Odum (1957) indicate that growth of water le ttuce in Silver Springs, much like water hyacinth, was not limited by nitrate-nitrogen levels of 0.4 mg /l. Instead, Odum’s (1957) work suggests that extreme accumulation of water lettuce in Silver Springs was largely prevented by downstream export, seasonal senescence, and insect attack. More anecdotally, Webber’s (1897, 13) reports of “large quantities of the water lettuce” in the Wekiva River during the late 19 th century could suggest that background nutrient le vels in that particular spri ng fed river were not a serious limiting factor for water lettuce growth. As sugge sted above for water hyacinth, detailed growth assays and/or comparative field studies would be needed to make more definitive determinations about the effects of nitrate-nitrogen enrichme nt on water lettuce population dynamics in Florida springs.

Hydrilla and Nitrogen Limitation

As a rooted submersed plant, determination of dissolved nutrient limitation for hydrilla is inherently more difficult than for the floating plants due to the confounding influences of sediment nutrient availability and uptake. With that said, a variet y of evidence is suggestive of important relationships between hy drilla growth and increased ni trogen enrichment in Florida’s aquatic systems. For example, tank experiments cl early indicate that hydrill a’s growth and ability to out-compete native tape grass ( Vallisneria americana ) under high light condi tions increases as a function of increased sediment nitrogen (Smart et al. 1994; Van et al. 1999). Similarly, Barko and Smart (1986) identified positive correlations between hydrilla shoot density and sediment nitrogen levels in field studies of North American lakes. More recent field surveys presented by Gu (2006) indicate a marginally si gnificant relationship between dissolved tota l nitrogen and the extent of hydrilla occurrence in Florida lakes, although stronger correlativ e relationships were identified with other water quality parameters such as alkalinity, pH, and total dissolved phosphorus.245 It must be cautioned, however, that straightforw ard inference from these studies into Florida springs conditions is confounded by at least two factors: 1) whether or not there is a significant positive relationship between dissolved nitrate- nitrogen and sediment nitrogen in springs ecosystems; and 2) what effects stream flow may have on disso lved nutrient availability for hydrilla and other submersed plants (Canfiel d and Hoyer 1988). Follo wing this second point, Terrell and Canfield (1996) found that dissolved nitr ogen availability for a quatic plants in Kings Bay/Crystal River, which had a mean TN concen tration of approximately 0.26 mg/l at the time of study (neither nitrate-nitroge n contribution to TN nor sedi ment nitrogen were reported), greatly exceeded the biological demands necessa ry for maximum growth and accumulation of hydrilla. Based on this finding, Te rrell and Canfield (1996) conc luded that nutrient reduction likely was not an effective strategy for reducing gr owth of hydrilla in Ki ngs Bay/Crystal River. Detailed studies that take into account dissolved nutrient availability, sediment nutrient stores, flow rates, native plant competition, and hydrilla productivity would be ne eded to determine if nitrate-nitrogen reduction is a f easible strategy for hydrilla suppre ssion in the context of other Florida springs.

NONNATIVE PLANTS AND AL GAL COMMUNITY DYNAMICS

While overgrowth of nonnative invasive plants is a serious management i ssue, most conservation and management concern in Florida springs over r ecent years has been associated with increased cover of filamentous algae and cyanobacteria such as Vaucheria sp. and Lyngbya wollei (Stevenson et al. 2004). Given the strong trophic and stru ctural associations between native aquatic plants, epiphytic algae, and animal communities histor ically found in springs ecosystems (Odum 1957), it stands to reason th at the obvious structural cha nges in plant community caused by invasive plant growth potential ly could trigger structural ch anges in the algal and animal communities.

This section utilizes scientific literature about major invasive pl ants to explore two mechanisms by which these plants might cont ribute to a successional trajectory in wh ich undesirable algae become dominant in springs ecosystems. The first proposed mechanism, referred to as “boom bust,” focuses on the implications of the increased productiv ity, biomass accumulation, and vulnerability to catastrophic lo sses and rapid community reorgani zation that are associated with invasive plant communities. The second propos ed mechanism focuses on the allelopathic properties of the invasive plants, with particul ar attention to the potential implications of algicidal allelopathy.

Boom-Bust Hypothesis

Perhaps the most basic functiona l trait shared by water hyacinth, water le ttuce, hydrilla and other invasive aquatic plant species is higher productiv ity relative to competing native plants (Schardt 1997). Thus, the introduction of a successful inva sive aquatic plant is almost axiomatically associated with a boom cycle in which more biomass is produced and accumulated in the ecosystem. Another feature of water hyacinth, water lettuce, and hydrilla in Florida is that these initial booms are often followed by rapid population crashes, or busts. These busts can be caused by aquatic plant control or a variet y of stochastic natura l events such as saltwater storm surges246 (Terrell and Canfield 1996), high water/floods (Bar todziej and Leslie 1998; Bureau of Invasive Plant Management 2007), and killing frosts (Joyce 1985; Schardt 1997).

Figure 6 gives a simplified diagram of mechanis ms by which a boom bust cycle associated with invasive plants potentially could lead to in creased nuisance algae in springs systems. The diagram shows that one major ecosystem effect often associated with invasive plants is competitive displacement of nativ e plant species (e.g., Schmitz et al. 1993; Bartodziej and Leslie 1998), some of which may have previously suppre ssed the growth of problematic algae such as Lyngbya wollei (see Doyle and Smart 1998). While there is significant evidence to suggest that hydrilla, water hyacinth, and wate r lettuce are likely to suppress algal production through shade, nutrient uptake, and other feedbacks (e.g., Cohen 1993; Cowell and Botts 1994; Kim et al. 2001), it is plausible that the sequence of compet itive displacement of native plants followed by a bust cycle characterized by de struction of dominant invasi ve plant populations may favor succession by opportunistic filamentous algae. Fila mentous algae have been observed to fill in such ecological voids after aqua tic plant control of hydrilla in Kings Bay/Crystal River (Cowell and Botts 1994) and Wakulla River (Loper et al. 2005). Similarly, Evans (2007) observed an apparent increase of filamentous algal bioma ss in the Ichetucknee River following manual removal of water lettuce. Large algal blooms following control of water hyacinths, while not specifically documented in Florida springs system s, have been documented in other ecosystem contexts (see Clugston 1963; Bicudo et al. 2007).

A mechanistic feedback that may further favor proliferation of nuisan ce algae after a boom bust invasive plant cycle is implied by the work of Stevenson et al. (2007), who f ound that nutrient enrichment of sediments was significantly associ ated with presence of nuisance algae in springs. While there has to date been no detailed study of nutrient accumulation in springs sediments as a function of aquatic plant control or nonnative pl ants, aquatic plant control activities were suggested as a possible source of enriched sediments documen ted by Wetland Solutions, Inc. (2006) in the Wekiva River and Rock Springs Run system. More detailed study of such Higher productivity

Native Plants Invasive Plants Nuisance Algae Disturbance event

Faunal changes, sediment accumulation

Figure 6 – Conceptual diagram of boom-bust hypothesis in Florida springs Successional change over time247 relationships appears warranted in springs, part icularly because invasive plant growth and control has been linked with nutri ent enrichment of sediments in other ecosystem contexts. For example, Brenner et al. (1999) suggested that the introduction and subsequent chemical treatment of water hyacinth and hydrilla likely was a major contri butor to nutrient enrichment and bulk density increases documented in the sediments of Lake Hell ‘n’ Blazes from approximately 1900 to 1995. Similarly, Grimshaw ( 2002) identified chemical treatment of water hyacinth and water lettuce as a significant source of nutrient-rich orga nic sediments in the Kissimmee River and Lake Okeechobee. In a highl y detailed study of water hyacinth control (based on a “trituration” technique of shredding the aquatic plants and loading the biomass into the water column) in Mexico’s Valsequillo reservoir, Mangas-Ramirez and Elias-Gutierrez (2004) recorded increases of ammonia from 3 mg/l to 60 mg/l in sediment interfaces, a 320% increase in dissolved nitrate-n itrogen, declines in dissolved oxygen to below 0.4 mg/l, increased rates of organic sedimentation, disappearance of most fish, and large blooms of cyanophytes in the reservoir following weed trituration.

Although these studies are suggestive of the impacts that invasive aquatic plants and subsequent control efforts may have in terms of sediment enri chment in springs, it is important to note that a variety of factors such as flow velocity, external inputs of or ganic matter, upwelling of nutrient rich groundwater through karst streambeds, and the densities and accumulation of plant biomass in specific springs would all be expected to ex ert significant influen ce over sediment nutrient levels. Detailed measurements of sediment nut rients, dissolved nutri ents, and plant/algae community characteristics in discrete samples ta ken directly before, directly after, and in sequential intervals following aquatic plant contro l would provide invaluab le information about the overall ecosystem effects of current management techniques on springs ecosystems.

Allelopathy Hypothesis

Algal community composition in springs and ot her streams are clearly shaped by ecological factors such as nutrient availability, flow ve locity, light intensity, grazer abundance, and substrate quality (e.g., Stevenson et al. 2007). In addition to th ese more general ecosystem variables, algal communities are also shaped by complex competitive and/or synergistic interactions between algae and vascular plants at the species level. Emission of allelopathic compounds, or compounds that directly suppress the growth of other primary producers, is increasingly regarded as one of the most important, if not well understood, competitive factors in aquatic systems (Gross 2003).

Several scientific studies indi cate that water hyacinth, wate r lettuce, and hydrilla all emit allelopathic compounds that restri ct growth of algal and/or plan t competitors. In the case of water hyacinth, Jin et al. (2003) conducted lab experiments indicating that various compounds extracted from water hyacinth r oots had algicidal properties on Chlorella sp., Scendesmus obliquus , and undifferentiated phyto plankton that were comparable in activity to c opper sulfate – a commonly used commercial algicide. Similarly, Aliotta et al. (1991) isolated several allelopathic chemicals from water lettuce that sh owed inhibitory effects on seventeen of nineteen algal cultures, with Lyngbya kuetzingii and Chlorella saccharophila showing no inhibition. While specific algicidal compounds have not been identified in hydrilla to date, Kulshresthna and Gopal (1983) did find that hydr illa negatively affected the grow th of vascular aquatic plant248 Ceratophyllm sp. through allelopathic mechanisms. However, Glomski et al. (2002) later argued that allelopathy is only a very minor factor in hy drilla’s overall competitive success against other aquatic plants.

The potential implications of al gicidal allelopathy in the floating plants are interesting to consider in the context of Florida springs system s. On the one hand, allelopathic emissions from floating plants may be a mechanis m by which fringe mats of floati ng plants in spring runs serve as a constraint for algal growth and accumulatio n in habitat areas beyond those covered by the plants, including beds of native submersed plants lo cated in the main stream flow. If this is the case, then the algicidal propertie s of floating plants may potenti ally serve as a buffer against algal overgrowth in nutrient-enric hed springs. On the other hand, it is plausible that persistent presence of allelopathic compounds with algicidal properties might select for relatively resistant algal taxa, potentially including Lyngbya sp. (e.g., Aliotta et al. 1991).

Better understanding of the algicida l activity of floating plant allelopathic emissions on springsspecific algae species would requi re isolation of algicidal co mpounds and detailed bioassays similar to those conducted by Aliotta et al. (1991) and Jin et al. (2003). In addition, measurements of the concentration ranges for su ch compounds in the wate r of springs would be required to understand the extent to which allelopa thic mechanisms may be an important driver of algal community selection in springs communities.

AQUATIC PLANT CONTROL AND ALGAL SUCCESSION

Control of nonnative invasive plants is one of the most common management actions performed in Florida’s aquatic ecosystems, including many of the state’s springs. As discussed above, chemical control using herbicides registered by th e EPA for use in aquatic systems is the primary tool used by aquatic plant managers. The most obvious ecological concern a ssociated with use of aquatic herbicides is non-target damage to nativ e plants, algae, and animal communities, either through direct toxicology of herb icides or the ecological conseque nces of rapid plant senescence. Although aquatic plant managers and herbicide app licators take great precautions to avoid major non-target impacts, the complexities of ecosystem s make some non-target damage unavoidable.

This section reviews literature associated with two of the most plausible and commonly suggested mechanisms by which chemical cont rol of nonnative invasive plants can disrupt aquatic ecosystems: 1) preferential selection of herbicide resistant algae species; and 2) suppression of key algal-grazing fauna through di rect toxicology and/or habitat destruction. Potential implications of these mechanisms for Florida springs, partic ularly in terms of succession towards filamentous algae/cyanobacteria communities, are also discussed.

Herbicide Selection Hypothesis

Development of herbicide resistance among algae has been repeatedly de monstrated at both the cellular level of individual species (i.e., a specific algal species evolves resistance to an herbicide) and the ecological le vel of community assemblages (i .e., more herbicide-resistant species become dominant) (e.g., Solberg and Higgins 1993; Boswell et al. 2002; Garcia-Villada et al. 2004; Cooke et al. 2005; Lopez-Rodas 2007). Thus, a ma jor concern associated with249 repeated use of herbicides in aquatic systems is the potential selection for herbicide-resistant algal biotypes. In particular, recen t research suggests that some herbicide-resistant algal species often may be less palatable and/or nutritious to algal-grazing fauna , thus causing a cascade effect on faunal community structure and overall ecosystem organization (e.g., Weiner et al. 2007). This section discusses potential algal-selection concerns associated with the major herbicides used in Florida springs: copper, diquat, en dothall, glyphosate, and 2,4-D. When relevant information is available, attenti on is particularly given to potentia l selection concerns associated with Lyngbya wollei

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Copper

Copper compounds are most often us ed in aquatic systems as al gaecides, but are listed an approved method for control of hydrilla and other s ubmersed plants species in Florida (Bureua of Invasive Plant Management 2007). Largely due to concerns about sediment contamination and bioaccumulation risks in manatees (O’Sh ea 1984; Facemire 1991; Leslie 1992), copper herbicides have not been wide ly used in Florida since th e early 1990s. However, copper herbicides were used extensivel y for hydrilla control in the Ki ngs Bay/Crystal River ecosystem during the 1970s and 1980s (Haller et al. 1983), and currently are bei ng researched as a potential option for filamentous algae contro l in Kings Bay/Crystal River.

Development of copper resistant algal communities is a well-documented risk associated with long-term usage of copper herbicides. Recent re search, for example, indicates that repeat exposure to copper sulphate can facilitate evolution of copper-resistant strains in both the cyanobacterium Microcystis aeuruginosa (Garcia-Villada et al. 2004) and the green algae Chlamydomonas reinhardtii (Boswell et al. 2002). Cooke et al. (2005) report that repeated treatment with copper sulfate resulted in increas ingly copper resistant algal communities in a number of lakes and reservoirs throughout the northeaste rn United States. With such selection dynamics in mind, it is notable that Lyngbya sp. and Oscillatoria sp., which are problematic taxa in some Florida springs (Stevenson et al. 2004), are known for their ability to develop copper resistant biotypes (Spencer and Lembi 2005). Evans et al. (2007) argue that such selection dynamics may provide a key clue for better unde rstanding of algal community dynamics in Florida spring systems, such as Kings Bay/Crysta l River, where copper herbicides have been used historically.

Diquat

Scientific literature indicates that concerns asso ciated with selection of herbicide-resistant algae communities may also be associated with large-scale use of diquat dibromide in Florida springs. Diquat is most commonly used for treatment of mixed stands of water lettuce and water hyacinth, but has also been used for control of hydrilla and other submersed species in some springs systems (Haller et al. 1983). Although residues of di quat do accumulate in aquatic sediments, most recent research suggests that diquat forms complex bonds with sediments that render it biologically unavai lable (Emmett 2002).

Bioassay research indicates that there is a wide range of tolera nces among algal ta xa to diquat exposure. Peterson et al. (1997) found that 50% inhibitio n of biomass production occurred250 among two green algae taxa ( Scendesmus quadricauda and Selenastrum capricornutum ) at concentrations of approximately 0.6 mg/l of di quat, while 50% growth inhibition for diatoms ( Nitzschia sp. and Cyclotella meneghiana ) and cyanobacteria ( Oscillatoria sp. and Pseudoanabaena sp.; Microcystis sp.) occurred at concentrati ons approximately one order of magnitude lower (0.074 mg/l and 0.079 mg/l of diquat, respectively). Phlips et al. (1992), however, found a wider range of diquat se nsitivity among the taxonomic groups, with Euglena gracilis

, Chlorella vulgaris , and Skeletonema costatum showing high levels of tolerance to diquat (50% inhibition predicted at over 2.94 mg/l). Lyngbya wollei showed a higher level of tolerance to diquat (50% inhibi tion predicted at a range of 0.081 mg/l to 0.205 mg/l of diquat) in the Phlips et al. (1992) bioassays as compared to the cyanobacteria tested by Peterson et al. (1997). Aquatic plant control observ ations indicate that diquat often proves ineffective for longterm control of Lyngbya wollei (Bayne 2005), which may be suggestive of an evolutionary resistance capability with regards to diquat in this cyanobacterium.

A potential implication of the differential tolera nces exhibited by algal- taxa is that repeat treatment with diquat may exert a chronic selective pressure that favors community-level dominance by diquat-resistant sp ecies and/or biotypes over time. Much like with the copper herbicides, diquat treatment hist ory and sediment residues may be an important variable to consider in studies of algal community dynamics in Florida springs where this herbicide has been commonly used for aquatic plant control.

Endothall

The dipotassium salt formulation of endothall, whic h is generally referred to by the trade name Aquathol, is the primary herbicide used for cont rol of hydrilla in flowing water systems. While Aquathol is not typically recomm ended for algae control purposes, an amine endothall salt (trade name Hydrothol) is registered fo r control of filamentous algae and submersed plants in Florida. However, Hydrothol is not commonl y used in Florida’s public waters due to its relatively high toxicity to fish and other aqua tic fauna. Field observations sugge st that Hydrothol applications often prove ineffective for Lyngbya wollei control (DuBose et al. 1997), and, as noted above for diquat, such observations may be indicative of an evolutionary resistance to endothall-based compounds among Lyngbya sp. Specific bioassays testing th e algaecidal action of Aquathol would be needed to determine if repeated use of this compound has significant implications for algal community structure in Florida springs.

Glyphosate

An aquatic-registered form of glyphosate (trade na me “Rodeo”) is sometimes used for control of water hyacinths in Florida (Bur eau of Invasive Plant Manage ment 2007). While generally not used as an algaecide in aqua tic environments, bioassays i ndicate that glyphosate does have algaecidal action. Wong (2000) reporte d significant growth inhibition of Scendedesmus quadricauda at 2 mg/l of glyphosate and complete inhi bition at 20 mg/l. In very recent bioassay work with Microcystis auroginosa , Lopez-Rodas et al. (2007) reported cons iderable inhibition in the cyanobacterium from 10 mg/l to 60 mg/l of glyphosate solution, and observed the evolution of glyphosate resistant M. auroginosa biotypes at concentrati ons of 120 mg/l. While the glyphosate concentrations of concern to Lopez-Rodas et al. (2007) are at least an order of251 magnitude over realistic field conc entrations associated with aquatic plant control (Langeland 2006), it is plausible that the lower end of concen trations showing growth inhibition in the Wong (2000) experiment may be reached during aquatic plant control operations. Bioassays testing glyphosate response curves of alga l species in Florida springs would be needed to better determine algal-selection risks associ ated with this herbicide.

2,4-D

Algal toxicity bioassays with 2,4-D have found th at this herbicide is generally non-toxic and even stimulatory to most algae and cyanobacteria at concentrations used for aquatic plant control (Okay and Gaines 1996; Wong 2000). The likely reason for the stimulat ory effect is that 2,4-D’s mode of herbicidal action is mimicry of the plan t hormone auxin, rather th an direct targeting of photosynthesis mechanisms. Toxic effects on al gae are reported at over 200 mg/l of 2,4-D by both Okay and Gaines (1996) and Wong (2000), but such concentrations are well outside of those associated with aquatic pl ant control. Interestingly, Okay and Gaines (1996) report that the amine form of 2,4-D is preferentially consumed as a nitrogen source over nitrate-nitrogen by the phytoplankton Phaedactylum tricornutum and Dunaliella tertiolecta when 2,4-D is found at concentrations up to100 mg/l, mean ing that changes in algal comm unity could be promoted by 2,4-D through an enrichment mechan ism, rather than a toxicity mechanism. Like with the other herbicides, specific bioassays testing growth response to 2,4- D among algal sp ecies found in Florida springs would be needed to better un derstand the community level effects of this herbicide.

Attractor – Catastrophe Hypothesis

A final mechanism by which inva sive plants and their subse quent management may profoundly affect ecosystem structure in sp rings is through a seque ntial cycle that can be deemed “attractorcatastrophe,” which is somewhat similar to the boom-bust cycle discussed above. The basic thrust of this hypothesis is that nonnative invasive plants can beco me preferred habitat, or serve as an “attractor,” for key faunal species, but that this habitat ultimately becomes a persistent sink for faunal populations due to the bust cycle “cata strophes” that a ffect nonnative plant communities. Although any habitat disturbance that destroys non-native plant communities is potentially a catastrophic sink for associated faun al communities, primary focus is given here to the non-target toxic effects and/or general hab itat disturbance that may be associated with aquatic plant control activities. The major rationale for this focu s is that aquatic plant control activities are unique in the sense that they are directly controlled by management agencies, and thus may be adjusted for the purposes of adaptive learning.

Habitat Values of Water Hyacinth

Bartodziej and Leslie’s (1998) long-term study of ecological communities a ssociated with water hyacinth in the St. Marks River cl early demonstrates the high habita t values that can be provided by this nonnative plant in Florida springs. Contrary to other ecosystem c ontexts in which large mats of water hyacinth have been clearly show n to depress dissolved oxygen, coverage by water hyacinth in the flowing waters of the St. Mark s River did not show si gnificant oxygen profile differences as compared to strap-leaf sag ( Sagitraria kurziana ) or other native plant252 communities. While native strap-leaf sag in the St. Marks River was reduced by water hyacinth expansion, it was not displaced entirely as the water hyacinth community was rarely observed by Bartodziej and Leslie (1998) to reach more than 25% coverage along the river width (see Figure 7) due to consistent flushing of the floating plants by river current.

Long-term faunal surveys indicated that the co mmunity associated with water hyacinths was generally more diverse in terms of invertebrate taxa a nd included signif icantly higher abundances of spring run crayfish ( Procambarus paludosus ), amphipods ( Hyalella azteca ), grass shrimp ( Palaemonetes paludosus ), and several fish species as compared to strap leaf sag communities (see Figure 8). Apple snails ( Pomacea paludosa ) were also found to have much higher population densities in water hyacinth compared to strap leaf sag. Conversely, rasp elimia snail ( Elimia floridensis ) populations in strap leaf sag were much higher than those found in water hyacinth communities (see Figure 8). Surveys of bird popul ations among water hyacinth in the St. Marks River indicated that water hyacint h was commonly used for forage habitat by little blue herons ( Egretta caerulea ) and tri-colored herons ( Egretta tricolor ), and was also observed to increase available nesting ha bitat for the common moorhen ( Gallinula chloropus ) (Bartodziej and Weymouth 1995).

Figure 7 – Water hyacinth fringe in River Rise Spri ng #3, St. Marks River. Photo by Northwest Florida Water Management District (2006). http://www.nwfwmd.state.fl.us/rmd/sprin gs/Wakulla_StMarks/photos/rise3_1.jpg253 Figure 8 – Ecosystem survey results from St. Marks River (Bartodziej and Leslie 1998) showing higher numbers of spring run crayfish (Procambarus paludosus) , grass shrimp (Palaemonetes paludosus) , and total fishes in water hyacinth (treat ment) as compared to strap leaf sag communities. Rasp elimia snail (Elimia floridensis) populations are considerably higher in strap leaf sag.

Habitat Values of Water Lettuce and Hydrilla

Detailed habitat assessments in a springs ecosystem are not available for either water lettuce or hydrilla. However, as su ggested by Carr (1994), it may be reas onable to assume that the similar structure of water lettuce and water hyacinth is likely to provi de habitat for a similar faunal species assemblage. Schmitz et al. (1993) note that water lettuce supports high concentrations of Hyalella azteca , and Evans (2007) anecdotally reports la rge numbers of spri ng run crayfish and apple snails in water lettuce ha rvested from the Ichetucknee River in 2000 – 2001. Similarly, Corrao et al. (2006) report high concentrations of appl e snails on hydrilla and apple snail eggs on water lettuce at Wacissa Springs. In other ecos ystem contexts, hydrilla is known to support large macroinvertebrate populations (Schramm and Jirk a 1989) and attract dive rse assemblages of water fowl (Johnson and Montalbano 1984; Esler 1990). Dramatically increased manatee utilization of Kings Bay/Crystal Ri ver beginning in the 1960s (Kochman et al. 1985) coincides254 with the introduction and rapid spread of hydrill a, a preferred manatee forage (Campbell and Irvine 1977), in that system.

Non-Target Concerns

As noted by Bartodziej (1992) for water hyacinth-associated amphipod populations and Corrao et al. (2006) for apple snails, the effects of habita t loss associated with aquatic plant control on faunal communities in springs is largely unknown. Corrao et al. (2006) suggested that direct removal of apple snails during harvest of hydrilla at Wakulla Sp rings and herbicide suppression of plants containing snail eggs at Wacissa Springs may have had significant negative effects on apple snail populations, but noted th at more detailed studies woul d be necessary to test these observations. Similarly, Evans (200 7) observed that harvest of wa ter lettuce from Ichetucknee Springs resulted in direct mortal ity of several faunal species, in cluding apple snails and spring run crayfish, but also noted that more detailed studies would be necessary to quantify the overall faunal impacts of plant harvest.

Use of copper herbicides has ra ised significant non-target concer ns in Florida springs. Most notably, copper herbicide use in Kings Bay/Crys tal River was discontinued in the late 1980s due to the finding of severe accumulation of coppe r in aquatic sediments (Facemire 1991; Leslie 1992) and the tissues of manatees (O’Shea 1984). More systematic concerns about long-term copper herbicide usage are suggested by Cooke et al. (2005), who note that dissolved copper can severely suppress zooplankton gr azers and that sediment cont amination can have long-term effects on the structure and function of be nthic communities. Increased levels of Lyngbya wollei often were observed in Kings Ba y/Crystal River subsequent to hyd rilla herbicide treatments (see Cowell and Botts 1994), which led Evans et al. (2007) to hypothesize that toxicological suppression of grazer communities a ssociated with copper herbicide usage may have contributed to subsequent algal overgrowth.

Risk assessments on four grazer species perfor med by Mastin and Rodgers (2000) indicate 50% lethal concentration dosages 2 (LC50) at 48 hours for copper herbicides between 0.011 mg/l – 0.029 mg/l for Daphnia magna ; between 0.158mg/l to 0.433 mg/l for Hyalella azteca ; between 0.374 and 0.114 mg/l for Chironomus tetanus ; and between 0.019 and 0.48 mg/l for Pimephales promelas . Copper concentrations reported by Haller et al. (1983) in the water of Kings Bay following a copper herbicide trea tment ranged from 0.002 mg/l to 0.011 mg/l, which is generally lower than the above LC50s (with the exception of Daphnia magna ) reported by Mastin and Rodgers (2000). Dick (1989), how ever, reported that poor contro l of hydrilla using standard dosage rates sometimes led aquatic plant managers to apply significant hi gher rates of copper in Kings Bay/Crystal River than those monitored by Haller et al. (1983). Copper’s toxicity to organisms in dissolved form generally decrease s as a function of incr eased hardness due to carbonate-induced precipitation of copper compounds (Mastin a nd Rodgers 2000). Thus, it can be argued that non-target risks in Florida sp rings ecosystems from di ssolved copper may be mitigated due to the relatively high hardness values in most springs as compared to the water hardness values (48 – 100 mg/l) reported in the Mastin and Rodgers (2000) experiment. Clearly,

2 A 50% lethal concentration dos e (LC) is the concentration of a chemical that results in 50% mortality of a test organism within 24 to 96 hours.255 any future use of copper compounds in Flor ida springs ecosystems should be implemented cautiously and carefully monitored to mini mize the secondary effects on fauna.

Risk assessments and toxicological data for diquat indicate that use of this herbicide may also raise serious non-target concerns for springs fauna. In a risk a ssessment for diquat conducted for the State of Washington Department of Ecology, Emmett (2002) reports that diquat is “very highly toxic” to H. azteca , with a LC50 of 0.048 mg/l, and “highly toxic” to the apple snail ( Pomacea paludosa ) with a LC50 of 0.34 mg/l. Nicholson and Clerman (1974) noted the high sensitivity of H. azteca to diquat toxicity in laboratory st udies, and suggested that a combination of direct toxicity and the dest ruction of aquatic plants thr ough herbicide usage were likely mechanisms for drastic H. azteca declines observed in New York ’s Chautauqua Lake during the 1960s and 1970s. Given that H. azteca is an important grazer of Lyngbya wollei (Camacho and Thacker 2006), it is reasonable to hypothe size that toxicological suppression of H. azteca potentially may be a factor in the increased dominance of L. wollei in Florida springs where diquat has been widely used.

Available literature suggests that very littl e non-target toxicological effects on fauna are associated with fluridone (e.g., Hamelink et al. 1986; Haag and Buckingham 1991), 2,4-D (Johnson and Finley 1980), or gl yphosate (Langeland 2006) at con centrations used for aquatic plant control. However, Wang et al. (1994) does document bioaccum ulation of 2,4-D after water hyacinth treatment in some fish specie s, and a breakdown product of 2,4-D (2,4- dicholorophenol) has been shown to have relativ ely high toxicity to ea rthworms (Roberts and Dorough 1984). Relyea (2005) has raised concerns about the effects of the Roundup formulation of glyphosate on amphibians, but La ngeland (2006) notes that the toxicological effects reported by Relyea (2005) are associated wi th an adjuvant in Roundup that is not used in the aquaticapproved formulations of glyphosate (e.g., Rodeo).

Aside from habitat loss and toxi cology, the other major concern as sociated with aquatic plant control is prolonged dissolved oxygen depression as plant biomass decays. Wetland Solutions, Inc. (2006) measured an increase in community respiration of 120% and a reduction of net primary production by 150% in a segment of Rock Springs Run following herbicide treatment of floating and emergent plants in December 2005. It was also noted that the trend of increased respiration continued for approx imately two weeks following the herbicide treatment, thereby resulting in “a detectable but te mporary impairment of the aqua tic community in Rock Springs Run” (Wetland Solutions, Inc. 2006).

Some of the most detailed monitoring of aquati c plant control in a spri ngs ecosystem has taken place over recent years in the Wakulla River. A report by DEP (2002) indicates that the stream condition index and habitat assessment scores at Wakulla River improved after an Aquathol treatment of hydrilla (Figure 9). However, significant non-target c oncerns were raised after state park biologists observed a signifi cant die-off of spring run cr ayfish following at least one subsequent Aquathol treatment (DEP 2006). In response to these concerns, DEP (2006) conducted a management experiment in which cr ayfish were monitored following aquatic plant control. Twenty chambers containing crayfish were deployed immediately prior to herbicide application, and nineteen were retrieved after four days. The remaining chamber was initially missed, but retrieved with a live crayfish after nine days. Another twenty chambers were256 deployed nine days after the herbicide applicati on and again retrieved f our days later. Data sondes showed that dissolved oxygen dropped fr om as high as 4.3 mg/l to approximately 2.0 mg/l after the herbicide treatment, and Aquathol exposure was meas ured at concentrations that ranged from 0.41 mg/l to 2.06 mg/l – levels known to be well under toxicological thresholds for crayfish. Despite these low oxygen levels, only one crayfish death among the forty replicates was observed during the duration of the experime nt. Although these results did not support the hypothesis that previously observed crayfish mort ality was caused by aquatic plant control, the possibility of drastically reduced crayfish f itness as a result of pr olonged dissolved oxygen depression that extended beyond the temporal scope of the experiments could not be ruled out (DEP 2006).

Figure 9 – Habitat and water quality assessments at Wa kulla River before (4/4/2002) and after (5/6/2002) Aquathol treatment of hydrilla (DEP 2002)

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RECOMMENDATIONS FOR FURTHER RESEARCH

This chapter opened by asking four sets of qu estions related to nonnative plants in Florida springs. The first asked to what extent increased n itrate-nitrogen might be a factor in the spread of the invasive plants water hyacinth, water lettuce, and hydril la, and at what level might invasive plant growth in springs be limited by nitrate-nitrogen. Availa ble literature does not provide a clear answer for determ ining a concentration of nitrate- nitrogen that would be limiting to nonnative springs, but observations suggests that nitrogen limitation is most likely alleviated at the very low end, and perhaps even close to background, concentrations in Florida springs. Detailed nutrient assay experiments in flowin g mesocosms and, in the case of hydrilla, with realistic sediment nutrient concentrations w ould be necessary to determine more accurate nutrient limitation parameters for nonna tive plants in spring systems.

The second question asked about the potential implications of i nvasive plant proliferation on ecosystem structure and function in springs. Available literature indicates that biomass accumulation, suppression of native plants, and al gicidal emissions associated with nonnative257 plants may produce feedback mechanisms that t ogether make springs systems more prone to a successional trajectory of overgrowth by nuisance alga l taxa. At the same time, scientific studies and observations indicate that the major invasive plants often provide highly attractive habitat for key springs ecosystem fauna and may support some feedbacks that could be expected to reduce algal overgrowth. Detailed habitat assessments, monitoring of sediment accumulation, and multiyear monitoring of vegetational succession patt erns in nonnative plant communities would be needed to better understand the re lationships between nonnative plan ts and algal overgrowth.

The third question asked about the potential imp lications of aquatic pl ant control on ecosystem structure and function. What emerge d from a review of literature was a wide range of concerns about possible algal selection, habitat loss, fa unal toxicology, and disso lved oxygen suppression that may be associated with a quatic plant control. With the notable exceptions of circa 1980s work conducted by several agencies in Kings Ba y/Crystal River due to concerns over copper herbicide contamination (Haller et al. 1983; O’Shea et al. 1984, Facemire 1991) and more recent work in Wakulla Springs associated with con cerns over crayfish mortality after Aquathol applications (DEP 2005), ther e has been very little monito ring of the overall ecosystem consequences of aquatic plant control in springs ecosystems. To better understand the overall ecosystem effects of aquatic plant control in springs communities, deta iled monitoring programs that measure key variables such as dissolved nutrients, dissolved oxygen, biomass deposition, and floral and faunal community structure should be monitored as a matter of course before and after major aquatic plant control operations.

As for the fourth question: are there opportunities for alternative management of invasive plants that may assist in the restorati on of desired ecosystem functions in highly degraded springs and spring run ecosystems? Given the notorious histor y of these plants recounted above, the mere suggestion that there may be anything “good” or even redeemable about water hyacinth, water lettuce, or hydrilla in the c ontext of springs ecosystems is undoubtedly controversial from the outset. However, discussions amon g ecological restoration practiti oners indicate an increasing willingness to consider the proposition that, given the diverse realities of modern environmental change (e.g., climate change, nutrient enrich ment, toxic contamination, and global species mixing), more adaptive approaches to nonna tive species control s hould be considered (D’Antonio and Meyerson 2002; Ewel and Putz 2004; Gobster 2005). From an adaptive management perspective, the overriding questi ons might be: Are nonnative species management efforts directly leading to an even worse set of problems (e.g., successi on into an even more undesirable ecosystem state), and, if so, how might future efforts be adjusted in the face of these emergent problems?

Assuming that the most prominent ecological c oncerns facing many springs ecosystems is the proliferation of nuisance algae such as Lyngbya wollei , an adaptive management question implied from several lines of evidence presente d above is the extent to which aquatic plant control may be a contributing factor in this proliferation. A compleme ntary question is how alternative methods of aquatic pl ant management could potentially help mitigate nuisance algae and/or restore habitat for desirable species.

But before exploring such alternatives in more detail, an initial boundary must be set that distinguishes between “prevention” and “contro l” when managing nonnative species. Clearly,258 none of the nonnative plants should be intentionally introduced into springs where they have not been present historically, and precautionary steps to prevent their introduction and/or spread into new systems should remain a high management priority. In addition, early detection and eradication efforts aimed at preventing ne wly discovered nonnative pl ant populations from becoming irreversibly established are also clea rly justified. Instead, th e opportunity for adaptive reflection is within those springs systems where nonnative plant species are extant and considered to be permanently established.

Alternative Management of Floating Plants

Establishment of selective biocontrol organisms often is regarded as th e most sustainable longterm method for controlling water hyacinth, wa ter lettuce, and other nonnative plants. A successful biocontrol organism should greatly reduce the potential for invasive overgrowth, which in theory should lessen, or even render unnecessary, the use of chemical herbicides and/or harvester machines over time (Center 1996; Dray et al. 2001). In practice, however, aquatic plant control methods in Florida have rarely been adju sted at a large scale in response to biocontrol introductions (Haller 1996). For sp rings in which curren t aquatic plan t control methods may be a significant disturbance factor th at selects for algal overgrowth, increased use of floating plant biocontrol organisms may be a promising alternat ive aquatic plant management strategy.

Interestingly, observations by Odum(1957) suggest that attack by an unknown insect (or insects) historically was an important control mechanism for water lettuce in Silver Springs. New research studies would be needed to determine whether or not such natural control mechanisms for water lettuce are currently at work in springs systems where water lettuce is established, or if the water lettuce moth ( Spodoptera pectinicornis ), an introduced biocontrol organism (Dray et al. 2001), is having signifi cant control effects.

Center et al. (2005) found that herbivory by water hyacinth weevils ( Neochetina eichhorniae and N. bruchi ) greatly reduced the competitiveness of wa ter hyacinth in relation to water lettuce. While water hyacinth’s superior competitive abil ities led to complete displacement of water lettuce, weevil herbivory allowed for competitive parity in which water lettuce could persist and, in some cases, become dominant. Bartodziej and Leslie (1998) evaluated the effects of water hyacinth weevil ( Neochetina eichhorniae ) on water hyacinth populations in the St. Marks River, and calculated that baseline weev il populations reduced the rate of water hyacinth expansion by approximately 10% over the course of a grow ing season. Bartodziej and Leslie (1998) also reported that augmentation of weevils successfully reduced population to an even greater extent, but estimated a prohibitively high augmentation cost of $40,000 per acre, or approximately 100 times the cost of chemical control. No field st udies or augmentation estimates for water hyacinth moth ( Sameodes albiguttalis ), another effective biological c ontrol agent for water hyacinth in Florida (Center 1984), were given by Bartodziej and Leslie (1998). Revised studies of weevil and moth herbivory on water hyacinth population dynamics in Florida springs, particularly if paired with detailed study of regular aquatic plant treatment si tes and provisional “no aquatic plant management” control sites, could potentially provide useful information for adjusting the intensity of aquatic plant treatment programs.259 A more controversial alternative management strategy for floating plants is experimental utilization of these plants in support of eco system recovery in highly degraded springs ecosystems. Both of these plants are well known for their luxury uptake of nitrogen (Aoi and Hayashi 1996; Ho and Tsang 1998), and also are known for their ability to suppress algal production through direct shading, nutrient co mpetition, allelopathy, and grazer habitat feedbacks (Joyce 1985; Aliotta et al. 1991; Cohen 1993; Ho and Tsang 1998; Kim et al. 2001; Jin et al. 2003). Mahujchariyawong and Ikeda (2001) de tail an ambitious use of water hyacinth for nutrient remediation in Thailand’s Tha-Chin River, and St. Johns River Water Management District recently has commissioned a proj ect to better understand the nitrogen reduction implications of allowing for greater colonizati on of water hyacinth and water lettuce in Lake George (SJRWMD 2006). Detailed ecosystem surveys by Bartodziej an d Leslie (1998) in the St. Marks River clearly indicate the high habitat value of water hyaci nth to a diverse variety of native springs fauna, and observations suggest that similar habitat values are likely associated with water lettuce. In springs with large numbe rs of resident and wintering manatees, it is reasonable to suspect that an increased availa bility of water hyacinth, which is known as a nutritious and sometimes pref erred manatee forage source (Lomolino 1977), might reduce herbivory pressure that can adversely affect desirable native subm ersed plant populations (Hauxwell et al. 2004). Studies could include small scal e pilot projects that study the ecosystem effects of variable water hyacinth and water lettuce coverage levels in springs, pa rticularly in those now affected by severe algal overgrowth. Comparisons among springs ecosystems with no aquatic control of floating plan ts (e.g., St. Marks River) and wi despread control of floating macrophytes (e.g., Wekiva River) may also be valuable.

Alternative Management of Hydrilla

Alternative management of hydrilla poses more apparent concerns than water hyacinth and water lettuce, as the disruptions associ ated with hydrilla in some spri ngs systems (e.g., Wakulla River) may actually be more severe than nuisance algae. In addition, the submersed growth habit makes hydrilla inherently harder to contain than the floating plants for the purposes of phytoremediation. However, hydrilla’s apparent abil ity to evolve resistance to fluridone (Michel et al. 2004) and concerns that it may do the same for endothall (Jones and Beardall 2004) may force alternative management strategies.

The most serious problem associat ed with hydrilla in Florida spri ngs is not that it provides poor wildlife habitat; rather, the major concern is its ability to form extensive topped out canopies that competitively exclude native plant species and radica lly diminish the recreational desirability of springs systems. Thus, the suggestion by Cuda et al. (2002) that herbivory by Cricotopus lebetis

, an aquatic midge of unknown origin discove red in Kings Bay/Crystal River in 1992, may effectively prevent hydrilla from topping out is potentially of gr eat importance for other springs ecosystems affected by hydrilla. Similar to the fi eld observations of stunted hydrilla in Kings Bay given by Cuda et al. (2002), Van et al. (1998) found that hydrilla’s competitiveness with Vallisneria americana was severely reduced in experiment al treatments containing biocontrol organisms Hydrillia pakistanae and Bagous hydrillae . Increased experimentation with release of such biocontrol organisms clearly should be a prio rity for Florida springs ecosystems affected by well-established hydr illa populations.260 Surprisingly, recent research in the Potomac Rive r has identified significant positive correlations between hydrilla colonizati on and long-term recovery of native plants, including Vallisneria
americana (Rybicki and Landwehr 2007). Cowell and Botts (1994) found that hydrilla coverage in Kings Bay had a significant nega tive relationship to coverage by Lyngbya wollei , perhaps indicating that hydrilla – particularly if overgrowth is contro lled by constant herbivory by insects and, in some cases, manatees – could, similar to the Potomac River (Rybicki and Landwehr 2007), have valuable functional benefits in te rms of guiding community succession away from filamentous algae and toward increased submersed plants in some springs systems. In springs ecosystems now affected by large hydrilla populations, management experiments might be set up to comparatively monitor the successional conse quences of different c ontrol strategies (e.g., biocontrol, herbicides, manual re moval, native plant restoration, and various combinations of these) over time.

CONCLUSION

There is little question that nonnative plant inva sions have had profound effects on the aquatic environments of Florida, including many springs , over the past several decades. While it is reasonable to suspect that nitrate-nitrogen en richment of springs may have contributed to overgrowth by nonnative aquatic plan ts, current studies provide insufficient information for making a definitive determination of nitrate-nitrog en concentrations that would effectively limit water hyacinth, water lettuce, or hydrilla growth in springs. Scientific literature suggests that both nonnative plant overgrowth and aquatic pl ant control techniques to suppress such overgrowth have the potential to se rve as severe disturbances th at could promote succession to algal dominated states in springs ecosystems. Because it is extremely unlikely that nonnative plant species can be entirely eliminated from sy stems in which they are established, the chapter argues that it may be beneficial to consider alternative and adaptive management strategies for nonnative plants, particularly in the context of sp rings and spring runs where algal overgrowth is now the primary management concern. Priorities for more adaptive aquatic plant management include: 1) intensive monitoring of ecosystem a nd successional impacts associated with current aquatic plant control methods; 2) increased re search and experimentation with biocontrol organisms; and 3) careful expe rimentation with ecosystem rec overy methods that utilize the nutrient uptake, algal suppression, and hab itat values of floating plants.261 REFERENCES

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# 6

Effects of Nutrients on Spring Ecosystems

Robert L. Knight, 1, 2 and Sky K. Notestein 1

1 Wetland Solutions, Inc.
2 Department of Environmental Engineer ing Sciences, University of Florida

Summary The relationship between nutrients and spring ecosystem structur e and function primarily focuses on the state-wide increase in spring nitrate co ncentrations derived from anthropogenic sources and the concurrent observed vi sual decline of these ecosyste ms. However, the apparent correlation between increased nitrate loading a nd declining aesthetic appearance of spring ecosystems has only anecdotally provided evidence for a causative relationship. Organism-level studies, focused on single species of algae, va scular plants, aquatic macroinvertebrates, and primary and secondary consumers, have produced mixed results that do not currently present a clear or consistent link between nutrient availability and populat ion structure in springs. Firm conclusions linking rising spring nitrogen concen trations and changes to individual components of biotic communities remain illusive, in some m easure due to the immense complexity of spring ecosystems and the large number of external forc ing functions (energy a nd matter inputs) that shape their structure and functions.

General systems theory predicts that any significa nt change in an important external factor ( e.g. , nitrate nitrogen delivery in ground water feeding a spring) should result in a significant change in the target system (in this case th e spring ecosystem), as long as the system of interest is highly adapted and operating at its maximum potential. If this type of cause and effect relationship holds for nutrients and springs, then measurements of whole ecosystem res ponses over a gradient of nitrate nitrogen concentrations should be an especially productive approach to applied research. Springs data presented and reviewed in this chapter reveal a significant correlation between rising nitrate concen trations and declining ecos ystem primary production and photosynthetic efficiency. While this possible cau se and effect relationship is based on detailed data from only a limited number of springs and their associated spring runs, so far this relationship has proven consistent.

While an inverse correlation between nitrate concentration and primary productivity appears counter-intuitive based on typical single plant species responses to increasing nutrients, it can be explained based on the subsidy- stress hypothesis that has been supported by a variety of ecosystem and organism-level studies. While low nitrate levels may be optimal for ecosystem productivity in pristine springs, higher nitrate le vels appear to have the opposite effect of reducing overall primary productivity. It can be theorized that spring eco system responses to nutrient increases integrate the observed complexity of their hundreds of important plants and animals. In their natural, moderate to low nut rient state, springs may be adapted for optimal272 efficiency of light utilization and maximum eco system metabolism (Maximum Power Theory). Rising nutrient concentrations may result in the competitive advantage of “weedy” algae and plant species that can capture li ght and spatial resour ces through higher rates of net productivity but lower gross productivity. C ontinuing elevated nutrient conc entrations in concert with differential grazing pressure, reduc tions in flow, or loss of othe r top-down control mechanisms may be linked to observed plant community shifts from adapted submersed aquatic plants with sparse but highly productive periphyton commun ities to systems dominated by benthic and attached filamentous algae. Data are also pres ented which link elevated nitrate concentrations with alterations in the reproductive functioning of amphibians and fish. These complex interactions between increased nutrients and algae, decreased dissolved oxygen due to lowered primary productivity, and altered grazer populations illustrate the importance of employing an ecosystem perspective when examining the multiple factors potentially affecting springs.

It is the conclusion of this re view that an ecosystem approach is essential for future research efforts to provide a greater understanding of th e relative interactions between the myriad physical, chemical, and biological fl uxes present in springs and thei r normal responses to rising nutrient levels. In summary these recommended a pplied research activiti es might include the following:

• Detailed baseline investigations of the “norma l” or “existing” ecosystem-level structure and response of a broad sample of first and second magnitude sp rings to key forcing functions, including sunlight, flow, and nutrient levels. These studies need to be fairly long-term (multiple years of repeated measures) and conducted over a representative sample of springs with a range of forc ing functions including discharge rates, groundwater chemistry, nutrient concentr ations, and recreational intensities;

• Multivariate statistical anal ysis of these baseline data to look for significant and repeatable patterns between independent an d dependent variables such as nutrient concentrations and photosynthetic efficiency;

• Design of in situ complex (multi-species) mesocosm studies to allow replication of spring plant community responses to a range of nut rient conditions under realistic and relevant spring environmental conditions;

• Whole-spring manipulation studies (if neces sary) to test the effects of possible management techniques such as controlling le vels of springshed nutrient loading, human recreational activities, alternative control methods for i nvasive species, and estimation of optimal consumer carrying capacities ( e.g. , manatee density).

There exists a wide-spread recognition of th e environmental and economic importance of artesian springs in Florida. A generous commit ment of public dollars and a focused and logical research agenda will be critical to pulling thes e precious but threatened natural resources back from their current declining path.273 ECOSYSTEM STRUCTURE

Primary Producers

The paradigm of eutrophication is that nutri ent enrichment (nitrogen and or phosphorus primarily) will result in an increase in biom ass in primary producers often with negative ecological and economic consequences (C loern 2001, Nixon 1995, Rabalais 2002). Although this view assumes that nutrients are the primary limiting factor for algae and vascular plants, and not some other factor such as light or grazing, observations from a variety of aquatic systems broadly support this assertion. For marine and es tuarine systems, nitrogen is typically identified as the nutrient limiting primary production. Th e effects of nitrogen on marine and estuarine systems, the pathways for nitrogen transport be tween land and aquatic ha bitats, and the positive correlation between nitrogen and primary production and often secondary production ( i.e. fishery yields) have been widely reviewed ( e.g. Hecky and Kilham 1988, Howarth 1988, Rabalais 2002). However in freshwater systems, phosphorus is the nutrient more co mmonly implicated as limiting for primary production and has been high ly correlated with phytoplankton and fisheries biomass, especially in lakes (Bachmann et al. 1996, Canfield 1983, Hoyer and Canfield 1991).

Among streams and rivers, a consistent limiting nut rient for primary production does not appear to be the case ( e.g. Tank and Dodds 2003) likely because of distinct regional differences and because the abundance and rates of primary pr oducers represent an integration of physical, chemical, and biological conditions . Examples of the variability of stream primary producers include a cross-system analysis of temperate streams in which Dodds et al. (1997) established that both total nitrogen and total phosphorus in th e water column were significantly related to benthic algal biomass, while Francoeur’s (2 001) meta-analysis of 237 nutrient enrichment studies in temperate streams f ound that 16% indicated a nitrogen (N) response, 18% indicated a phosphorus (P) response, 23% requi red N and P be added together for a response, and 43% had no response to N or P.

It is important to remember that spring runs and other lotic (flowing water) ecosystems can transform inorganic nutrients into organic materi al, thus preventing or reducing the downstream export of inorganic nutrients. However, thes e lotic ecosystems can become saturated with respect to relatively conservati ve inorganic nutrients such as phosphorus and recalcitrant forms of organic nitrogen and in some instances may not be able to fu rther assimilate inorganic forms through uptake, microbial utilization, and denitrification. It is unc lear at what concentration or loading rate saturation can occur, but likely is dependent on whether it is an acute or chronic nutrient load and on the magnitude of other lim iting chemical constituents or forcing functions such as light availability. Once saturation occurs, nitrogen e xport to downstream ecosystems will increase proportionally to nitrogen loading (Bernot and Dodds 2005). Whether saturation of nitrogen or phosphorus has occurred in Florid a spring ecosystems remains uncertain in most cases and is obviously a functio n of individual spring biogeoch emistry. In at least one case within portions of the Wekiva River system, wh ich is fed by Wekiwa and Rock Springs (WSI 2007a) there is evidence that phosphorus saturati on may have occurred (as evidenced by zero or negative assimilation rates).274 It is important to remember that thresholds exist for the capacity of all aquatic ecosystems to productively assimilate increase d nutrient loads and that if these thresholds are exceeded, harmful ecological consequences may occur. Our us e of this concept has or igin in the paper by E. P. Odum et al . (1979), titled: Perturbation Theory and the Subsidy-Stress Gradient , in which the model of “too much of a good thing” is exam ined in the context of ecosystem response to perturbation (Figure 1). An ex ample of a subsidy-stress gradient is the productivity of hardwood forested swamps in response to flooding, in wh ich forested swamp produc tivity was observed to increase under seasonal flooding conditions bu t to decline under permanent flooding (Conner and Day 1992). With regard to spring ecosyst ems, it is likely that a variety of physical, biological, and chemical inputs may operate indivi dually or simultaneously within subsidy-stress gradients. For instance, stream velocity can re place nutrients and remove waste products at low velocities, but is capable of sc ouring at high velocities, and gr azing of primary producers may promote algal productivity at moderate levels but a reduction in primary production at higher levels. Knight (1980) examined and described several examples of the subsidy-stress gradient in research at Silver Springs, including the effect s of herbivorous snails, carnivorous fish, and a trace metal (elemental cadmium) on prim ary productivity in stream mesocosms.

It is hypothesized in this chapter that nutrient s also function along a s ubsidy-stress gradient in aquatic ecosystems. At extremely low concentr ations overall system productivity of plant and animal life is low; as nutrient concentrations increase, so too does primary and secondary productivity, but at some point, productivity may decrease. Within aquatic ecosystems, nutrient subsidy-stress gradients are like ly to have unique performance curves for primary producers by functional group, i.e. algae vs. vascular plant, or ev en by species with in these groups. Considering this, when monitoring ecosystems for impairment or the effects of nutrient gradients, whole ecosystem measures may pr oduce diametrically opposite results than singlespecies studies and will be more informative for documenting overall ecosystem impairment than measures of individu al functional groups.

Algae

Pseudo-phytoplankton – The development of plankton in lo tic ecosystems is largely preventeddue to inadequate water resident times (Hynes 1970). A more accurate term is pseudophytoplankton, which is a mix of sloughed material from periphytic primary producers. Once entrained in the water column, this material will e ither be consumed, settle out or if it remains in suspension continue to grow through cellular division. The factors which affect pseudophytoplankton include the abundance of peri phytic primary producers, light, nutrients, temperature, and velocity. In flowing water system s, stream discharge or velocity will influence their rate of delivery downstream . Given that phytoplankton popul ations can double in as little as one to two days, the relationship between stream flow and phytoplankton abundance is likely to be inversely correlated (Allan 1995). Howeve r, should declines in spring discharge rates occur to the point that water resi dence time is greater than about three days, the development of true phytoplankton communities could occur in the spring run, as was suggested by Frazer et al. (2002) in their examination of limiting nutrients in several sp ring ecosystems using traditional phytoplankton bioassay methods.275 Figure 1 – A hypothetical example of two types of inputs and their resulting ecosystem perturbations due to increasing i nput levels. Nutrients and nitrate in particular could be viewed as an example of a usable input (top curve) , which have a subsidy effect on ecosystem productivity to a point beyond which stress is incurred (from E. P. Odum et al. 1979).

A limited number of studies are available whic h have examined pseudo-phytoplankton collected from spring ecosystems and their response to manipul ations of ambient nutr ient levels. Frazer et al. (2002) conducted nutrient add ition bioassay of pseudo-phytoplankt on collected from several locations of the Chassahowitzka, Crystal, Homosassa, and Weeki Wachee Rivers. These samples were allowed to grow under near-opt imal light and temperature conditions under a variety of nutrient treatments. These treatme nts included a control (without any additional nutrient additions), nitrogen (N) addition (0.4 mg NO 3 -N/L), phosphorus (P) addition (0.04 mg PO 4 /L), silica (Si) addition (0.4 mg Si/L), N+P addition (0.4 mgNO 3 -N/L + 0.04 mgPO 4 /L), and N+P+Si (0.4 mg NO 3 -N/L + 0.04 mg PO 4 /L + 0.4 mg Si/L). For those pseudo-phytoplankton samples collected at the upper sp ring run, bioassays with additi onal nutrients did not produce an increase in growth compared to controls with am bient nutrient levels in 61% of the experiments, suggesting that the majority of locations we re not nutrient limited. In the balance of experiments, 19% were phosphorus limited, 10% we re nitrogen limited and 10% were co-limited276 by nitrogen and phosphorus. Given that all four of these spring ecosystems are supplied by waters with elevated nitrate concentrations relativ e to historic values, it is not surprising that nutrients did not appear to be limiting in majority of the experiments. This observation was supported by the average (n=8) ambient concentr ations of nitrate where pseudo-phytoplankton were collected: Chassahowitzka was 0.39 mg/L , Crystal was 0.03 mg/L, Homosassa was 0.42 mg/L, and Weeki Wachee was 0.53 mg NO 3 -N/L. Crystal River, the spring ecosystem with the lowest ambient levels of nitrate, had a pseudo-phytoplankton community which was most responsive to nitrate additions as it exhibited ni trogen limitation in 75% of the experiments. The data from the Frazer et al. (2002) study suggest that when sp ring run nitrate concentrations exceed approximately 0.4 mg/L, nitrogen limitation is unlikely.

A second springs’ study utilizin g a pseudo-phytoplankton bioassay fr om material collected from Kings Bay, which contains multiple spring vent s, was completed by Saindon (2005). This researcher compared pseudo-phytop lankton growth rates for experi mentally manipulated nitrate and soluble reactive phosphorus concentrations at two sites, which had ambient nitrate concentrations of 0.083 and 0.039 mg N/L and ambient soluble reactive phosphorus concentrations of 0.011 and 0.006 mg P/L. Sai ndon (2005) observed signi ficant differences between sites in the relative growth rates of pseudo-phytoplankton, and that maximum biomass was influenced by both nutrient ra tios and nutrient concentrations, but that growth rates “did not determine the maximum achieved biomass, just how long it took to reach that biomass”. An important conclusion from Saindon’s (2005) wo rk was that maximum pseudo-phytoplankton biomass could be maintained at very low ambi ent nutrient concentrations when the N:P (by weight) ratios were either very low (N:P < 3) or very high (N:P > 27) and that when the N:P (by weight) ratios were between 3 and 27, maxi mum pseudo-phytoplankton biomass appeared to increase with increasing nutrient availability. By way of comp arison, historical N:P ratios for spring waters were reported by Odum (1957a) and Odum et al. (1953) and were generally below 7:1 (by weight) but not always, su ggesting that historically most springs had the potential for nitrogen limitation. By way of comparison, N:P ratios can be reported by weight or by atomic ( i.e. molar) ratio, the commonly repor ted Redfield ratio for N:P of 16:1 is expressed in atomic units and represents an N:P ratio of approximately 7:1 by weight (Redfield et al. 1963, Duarte 1992).

Filamentous and Benthic Macroalgae - These categorical terms include macroalgae commonly viewed as nuisance species and associated with declines in aesthetics of many springs such as Lyngbya or Vaucheria . It is important to note that Lyngbya or Vaucheria are indigenous genera of macroalgae and that Florida spring ecosystems ha ve (or had) a natural di versity of macroalgae (Whitford 1956). The mere presence of these speci es is not an indicator of decline, but an overabundance or monoculture of macroalgae especially in conjunction with a loss of native submersed vascular species may be cause for concern.

A comprehensive synthesis report on macroalg ae inhabiting Florida springs in relation to nutrients was recently completed by Stevenson et al. (2007). The macroalgae study included sampling from 29 different Florida springs and multiple laboratory approaches to elaborate on prior related work (Stevenson et al. 2004). Results for the Stevenson et al. (2007) report suggest that macroalgal taxa may not respond to nutrients uniformly, as the abundance of Vaucheria spp. in spring ecosystems was positiv ely correlated to nitrogen and phosphorus concentr ations with a277 clear threshold increase at 0.59 mg total nitrogen/L ; however, the abundance of Lyngbya wollei was not correlated to either nitrogen or phosphorus water concentrations in surveyed springs, although it was related to sediment phosphorus c oncentrations and indices of human activities within 1,000 m of the sampling site. Several possible explanations fo r the lack of better correlations between ambient nutr ient concentrations and macroalgae abunda nce are that other factors limit and influence the standing crop. Principally, inst antaneous field measures are incapable of incorporating prior environmental c onditions which have caused the current state. Other important forcing functions affecting macroa lgae were not measured as well, for instance in spring runs with strong flow and heavy recr eational use, macrophytes can easily be dislodged and swept downstream preventing the accumulation of substantial coverage or biomass. Another alternative is that some other factor, such as light, is primarily limiting, and hence nutrient supply can only be secondarily limiting. The phenomenon in which light supply appears to be the primarily limiting factor was observed by Ca nfield and Hoyer (1988) as well as Frazer et al. (2006) in their studies on the abundance of macrophytes in spring runs. The role that grazing plays on determining the standing crop of filamentous algae, particularly Lyngbya sp. is largely unknown as well.

Cowell and Dawes (2004) determined that increased populations of Lyngbya in the Rainbow River (nitrate > 1.0 mg/L) may be due to increased nitrate concentrations. Lyngbya wollei growth was stimulated by nitrate additions in the laboratory although no biomass increase was observed above about 1 mg/L of n itrate nitrogen. They concluded that based on their lab studies nitrate nitrogen concentrations would need to be reduced to below 0.3 mg/L in order to significantly reduce Lyngbya biomass. This research confirms that increased nitrate levels can increase benthic algal biomass, at least in the short term.

Stevenson et al. (2007) also conducted in situ bioassays of Vaucheria and Lyngbya in three different microcosm designs each with a ma trix of nitrogen and phosphorus treatments maintained under relatively opt imal growth conditions. From these experiments it appears nutrient concentrations affect th e growth rate and ultimately the abundance of macroalgae as growth occurred in most nutrient treatments, even at very low nutrient con centrations, yet fastest growth rates occurred in high nut rient concentrations and were a pproximately two or more times greater than under low nut rient concentrations. Based on the bioassay experiments, it was preliminarily estimated that to init iate a reduction in the abundance of Vaucheria in Florida springs total nitrogen and total phosphorus concentrations would have to be less than 0.59 and 0.026 mg/L, respectively; while Lyngbya reductions would require total nitrogen and total phosphorus concentrations lower than 0.25 and 0.033 mg/L, respectively (Stevenson et al. 2007). These conclusions suggest that the response of macroalgae to nutrients is complex, that both nutrient ratio and concentration ar e important, and that growth rate (or grazing rate) can have a large impact on ultimate biomass ( i.e. standing crop).

Periphyton and Epiphytes – These are terms for another ca tegory of primary producers which are commonly thought to respond to nutrient enri chment of spring water through an increase in their coverage, biomass, and/or occurrence of w eedy species. It is possible that these weedy species may escape top-down controls by grazers, reduce the light energy available to their vascular plant hosts, and ultimately have the potential to modify the spring ecosystem plant community. One study to directly examine nutrien t effects on periphyton abundance in a spring278 fed system was completed by Notestein et al. (2003) as an in situ bioassay in the Chassahowitzka River, a spring run in coastal Citrus County, Flor ida. This study added nitrate, phosphate, and a combination of nitrate and phosphate to the upstr eam end of flow-through tubes containing glass microscope slides in the spring run. This expe rimental design allowed the amount of periphyton colonizing glass microscope slides to be comp ared between nutrient treatments and with a control treatment over a one month period. Increased epiphyte abund ance (as measured by chlorophyll per unit area) was observed foll owing the addition of phosphorus, as well as phosphorus in combination with nitrogen, when co mpared to the control (no nutrient addition treatment). The results suggested that phosphor us was the primary nutrient limiting periphyton abundance in the upper spring fed portion of this system, however nitr ate additions supported additional periphyton growth over the control treat ment and therefore nitrogen appeared to be secondarily limiting (Notestein et al. 2003). The Chassahowitzka spring system had elevated nitrate concentrations, approxima tely 0.45 mg/L at the time of the study, so the finding that phosphorus (approximately 0.016 mg/L) was limiting is consistent with the elevated N:P nutrient ratios (28:1 by weight) observed during the study period. The observation that additional nitrate could stimulate periphyton abundance, even though nitrogen was not the primary limiting nutrient, suggests that a nutrient need not be absolu tely limiting for it to have a biological effect in a spring ecosystem.

In stream systems outside of Florida, there ha ve been numerous publica tions on the relations between periphyton abunda nce and nutrients ( e.g. Dodds et al. 1997, Francoeur 2001). Interestingly, there do not appear to be consis tent trends or factors which control periphyton abundance in stream systems, which may be the cas e with Florida’s spring systems as well. As the following interstate examples show, increas ed nutrient concentr ations can influence periphyton abundance, but the responses to nutrien t additions are frequently variable. Nitrogen alone may stimulate periphyton abundance (Crawford 1979, Mosisch et al . 1999) but often only when accompanied by increased light availability from removing riparian shading or adding artificial light (Busch 1978, Gregory 1980, Hill and Harvey 1990). The addition of phosphorus, both on its own and in combination with nitr ogen, has been shown to increase periphyton abundance (Elwood et al . 1981, Peterson et al. 1993, Bothwell 1985). Ot her essential elements including carbon supplied by organic sources such as sucrose (Warren et al . 1964) or carbon supplied via inorganic forms such as carbon dioxide and bicarbonate (Dickman 1973 and Crawford 1979) have also been shown to increas e periphyton abundance. Another factor that interacts with periphyton a bundance is grazing (Steinman et al. 1991, Hill et al. 1992). In a recent study, grazers were shown to control periphyton abundance at a variety of ambient nutrient levels, while in the absence of grazers there was a significant positive correlation between periphyton and nutrients (nitrogen and phosphorus simultaneously added, Chen et al. 2007). These examples show that periphyton abundance in streams can be influenced by nutrients and other ecological contro ls originating from the bottom up ( i.e. light availability) or top down ( i.e. grazers controls) and suggest that the factors controlling pe riphyton abundance in Florida’s spring systems needs furthe r research, likely at the level of in situ complex mesocosms that can colonize with realis tic populations and diversity of algae and invertebrates.279 Vascular Plants

Studies of two Florida springs systems, th e Chassahowitzka (Notestein 2001) and the Ichetucknee (Kurz et al. 2004), involved measuring the abundance of submersed vascular macrophytes and the concentration of water co lumn nutrients. Chassahowitzka had ambient water column nitrate concentra tions of approximately 0.45 mg/L and phosphate concentrations of 0.016 mg/L; while Ichetucknee had approxima tely 0.55 mg/L nitrate and 0.04 mg/L phosphate in the water column. These measurements are el evated compared to historic concentrations. Neither study showed a correlation between surface water nutrient concentr ations and the growth rate and/or abundance of submer sed aquatic macrophytes, which appe ars to be consistent with research from other non-spring fed stream syst ems. For example, Peltier and Welch (1969) experimented with Potamogeton pectinatus , and observed that nitrate concentrations above 0.44 mg/L and phosphate concentrations above 0.03 mg/L did not significantly increase growth, nor was there a correlation betw een the field growth of P. pectinatu s or Najas sp . when inorganic nitrogen concentrations were 0.2 to 0.3 mg/L and phosphate concentrations were 0.03 to 0.08 mg/L. Mulligan and Baranowski (1969), studied several macrophytes including Myriophyllum spicatum under greenhouse conditions, and found that optimum growth could be achieved with 0.1 mg/L nitrogen and 0.02 mg/L phosphorus, values wh ich are equal to or greater than historic concentrations for many Florida springs. In a study of 17 Florida streams, many of them spring fed, Canfield and Hoyer (1988) concluded that at concentrations above 0.3 mg/L nitrogen and 0.03 mg/L phosphorus, the potential for nutrient limitation of submersed aquatic vegetation was unlikely. Therefore, in many of the spring ecosy stems with current aver age surface water nitrate concentrations exceeding these values, particularly for nitrate, the point of nutrient limitation may have already been passed. Because rooted aquatic macrophytes are known to derive much of their nutrient supply from the interstitial pore water of the substrate (Bristowe and Whitcombe 1971, Carignan and Kalff 1980, Chambers et al . 1989) and that these pore-water nutrient concentrations are often much higher than the overlying waters (Stevenson et al. 2007), it is likely that vascular macrophyte nutrient demand can be met from sediments and helps explain the lack of correlation between water column nut rient concentrations and submersed vascular macrophytes.

Much concern over the apparent reduction from the historical abundance of submersed vascular plants in spring ecosyst ems exists, and is warranted due to the important habitat and the food web contributions these primary pr oducers make (see Chapters 1 a nd 4). As noted in the section on periphyton above, the potential causes for declines in submer sed vascular plant abundance could include resource competition from periphyton or filamentous and benthic algae, overgrazing by consumers, physical disturbance from recreation, and/or diseas e; it is likely that causative mechanisms responsible for declines in submersed vascular plant abundance are system specific and may not be universal. An additional explanation of reduced submersed vascular plant abundance, which could be rela ted to elevated nitrate concentrations, was provided by Boedeltje et al . (2005) who reported that high water column concentrations of nitrates (> 7 mg/L, a value higher than that obser ved in 98% of Florida springs) can significantly reduce the growth of amm onia-preferring rooted submersed species such as Potamogeton alpinus. Boedeltje et al. (2005) hypothesized that nitrate dominated nitrogen assimilation may lead to strong metabolic disturbances affecting th e levels of organic anions within those species adapted to ammonia uptake by the roots (such as P. alpinus ). This may explain the observed280 reduction in growth under high water column nitrate con centrations, although further experiments are needed to assess potentia l negative effects on other species.

Primary Consumers

Invertebrates

Nitrate pollution in Florida springs has been im plicated as a possible reason for the apparent decline in populations of the Florida apple snail (Pomacea paludosa ). To explore this hypothesis, Corrao et al. (2006) correlated snail density with nitrate conc entration measurements in six Florida springs and conducte d laboratory studies to examine short-term acute impacts of nitrate on adult and juvenile apples snails. They observed no correlation between apple snail densities and ambient nitr ate concentrations. LC 50 levels (the concen tration at which 50% mortality occurred) could not be determined de spite nitrate concentrations >500 mg/L (Corrao et al. 2006). Although juvenile apple snail growth was affected in two trials with nitrate concentrations of 504 and 622 mg/L, respectively, the EC 50 (the effective concentration of a compound where 50% of its maximal effect is observed) for these a pple snails generally exhibited little to no response to these nitrate concentrations which are orders of magnitude greater than those found in Florida springs. The authors suggested that perhaps other factors, such as habitat structural ch anges due to non-indigenous plan t invasions, might explain the observed declines in apple snail abundance (Corrao et al. 2006).

Studies on the effects of nitrat e on invertebrates outside of Florida vary considerably, with changes in survival of caddisfly ( Hydropsyche occidentalis ) demonstrated at concentrations starting around 2.2 mg/L (Camargo and Ward 1995), while juvenile marine shrimp ( Penaeus monodon ) tolerate nitrate concentrations > 2,000 mg/L (Tsai and Chen 2002), and no mortalities were observed in juvenile Australian crayfish ( Cherax quadricarinatus ) exposed to nitrate concentrations up to 1,000 mg/L (Meade and Watts 1995) suggesting that th e impact of nitrate must be assessed on a case by case basis.

Vertebrates

Studies examining the effects of nitrate on the reproduction and developmen t of two vertebrate species found in Florida spring ecosystems, ha ve been made for th e Southern toad ( Bufo
terrestris ) by Edwards et al. (2006) and the Eastern mosquitofish ( Gambusia holbrooki ) by Edwards and Guillette (2007). These researchers repor ted that nitrate concentrations in the range of 4 to 30 mg/L had the ability to alter endocrine physiology of these amphibians and fish. In laboratory experiments utilizing reverse osmosis water enriched with nitrate, Southern toad tadpoles metamorphosed earlier as nitrate concentr ation increased while in experiments utilizing spring water, tadpoles reared in high nitrate (30 mg/L) delayed metamorphosis, suggesting that water source was a factor beside s nitrate concentration affecting Bufo development (Edwards et al. 2006). In their study of adult male mosquito fi sh collected from springs with varying nitrate concentrations, Edwards and Guillette (2007) obser ved that nitrate concentration (up to 5 mg/L) were significantly correlated w ith decreased total sperm counts, although other characteristics such as pH and temperature also affected physiological condition of the mosquito fish, again suggesting that nitrate alone may not have been responsible.281 Secondary Consumers

Fish

Most fish utilizing spring ecosy stems fall in the category of secondary and or higher level consumers. Walsh and Williams (2003) examined fish and mussel species diversity in sixteen Florida springs and spring runs. For the purposes of this review, their da ta were compared to ambient nitrate nitrogen concentrations reported from fifteen of those springs. Although there were trends for lower fish and mussel species nu mbers with higher nitrate levels, they were not significant. Fish species diversity appears to be relatively constant at Silver Springs, as evidenced by the following: Hubbs and Allen ( 1943) reported thirty-fiv e species in the 1940s, Walsh and Williams (2003) collected tw enty-nine species in 2002, and Munch et al. (2006) observed thirty-three species in 2004-05. Although the number of fi sh species doesn’t appear to have declined in correlation with increased nitrat e, the biomass of fishes in Silver Springs has declined precipitously over the past 50 years (estimated 96% reduction since the 1950’s Odum study). The observed decline in fish biomass at Silver Springs (primarily channel catfish and striped mullet) may be hypothesized to be indirec tly affected by elevated nitrate concentrations and their apparent contribution to plant co mmunity changes and resulting lower overall ecosystem net productivity during this 50-year period. However, at the trophic level of consumers such as herbivorous and omnivorous fis h, the list of possible environmental stressors is lengthy, including limited access by riverine and marine species due to the construction of the Rodman Reservoir and dam in 1968 (Munch et al. 2006). This discussion illustrates the need to consider the complexity of spring ecosystems wh en designing experiment al studies and drawing conclusions from incomplete data sources.

Higher-Level Consumers

Ecosystem level changes that have the potential to affect higher level consumers as a result of nitrate enrichment are largely hypoth etical. If elevated nitrate con centrations cause qualitative or quantitative changes in primary producers, these could be reflected in quantitative changes in higher trophic levels including fish (catfish, gar, sunfish, bass, and pickerel), reptiles (alligators, softshell and snapping turtles), birds, and mammals (otters and manatees). Another potential impact to higher level consumers could be nitrat e toxicity resulting in methemoglobinemia or "blue-baby" syndrome. This syndrome is known to affect mammals; hence there may be a reasonable risk of impacting manatees whic h are known to drink spring water. A novel hypothesis has been presented to explain the apparently high number of perinatal deaths ( i.e. stillborn manatee calves) which account for 22% percent of all manatee deaths between 1992 and 2000 (USFWS 2001). Dr. W.T. Haller (UF prof essor, personal communication 2004) has suggested that if pregnant mana tee feed on hydrilla which has elevated tissue concentrations of nitrate or drink nitrate rich spring water their developing in-vitro calves may be susceptible to methemoglobinemia. This hypothesis warrants testing given the potential of impacting a threatened species due to nitrate contamina tion of ground water and spring ecosystems in Florida.282 Humans

A number of springs are ut ilized as a source for co mmercially bottled water ( e.g. Gainer Springs in Bay County, Ginnie Springs in Alachua County, and Silver Sp rings in Marion County (Scott et al. 2002)). Given the widespread occurrence a nd generally increasing magnitude of nitrate contamination in Florida’s springs, there may be the potential of spring wa ter utilized for human consumption to approach or even exceed current USEPA drinking water standards of 10 mg/L. The USEPA (2007) has good information concerning the effects of nitrate on human health ( e.g. http://www.ead.anl.gov /pub/doc/nitrate-ite.pdf , accessed December 2007). According to this USEPA source, nitrate is a normal component of the human diet, with th e average daily intake from all sources estimated at 75 milligrams. In h ealthy adults, about 5% is converted to nitrite by bacteria in saliva and additi onal bacterial conversion of n itrate takes place inside the alimentary tract and the stomach. A pH of the gastric fluid greater than five promotes the growth of nitrate-reducing bacteria and increases the conversi on of nitrate to nitrite. This is especially a concern for infants, whose gastrointestinal sy stems normally have a higher pH than those of adults. Nitrites and the compounds they can form with proteins ( e.g. nitrosamines), have been implicated in causing stomach cancer, but evidence remains inconclusive. Although nitrates alone are relatively nontoxic, their conversion to nitr ite in the human digestiv e system can in turn react with hemoglobin in the blood and crea ting methemoglobin, a form of the protein hemoglobin that is unable to transport oxygen. This results in a decrease d capacity of the blood to transport oxygen causing a c ondition known as methemoglobinemia. Although some amount of methemoglobin in blood is common, levels of 10% can cause the skin and lips to take on a bluish tinge (cyanosis), levels above 25% can cause weakness and a rapid pulse, and levels above 50 to 60% can cause loss of consciousness and death. It appears that infants are much more sensitive than adults to nitrates/nitrites, as essentially all deaths from nitrate/nitrite poisoning have occurred in infants; however a dults are susceptible to long-term exposure to lower levels of nitrates and nitr ites which can cause diuresis (an increase in the amount of urine, and starchy deposits and hemorrhaging of the sp leen). Given the potential for human health impacts, the USEPA has designated a drinking wate r standard of 10 mg/L for nitrate and 1 mg/L for nitrite. At this time, only a minority (< 2%) of Florida springs currently have nitrate concentrations great er than 10 mg/L.

ECOSYSTEM FUNCTION

Community Metabolism

We define the term community metabolism to include the characterization of the primary producers (autotrophs) and consumers (primary, s econdary, etc. heterotrophs and decomposers) which collectively comprise the living compon ent of spring ecosystems (see Chapter 1). Because primary producers generate oxygen through photosynthesis and both producers and consumers use oxygen through re spiration, the measurement of oxygen (or carbon dioxide) dissolved in the water can serve as a proxy for community metabolism. In addition, the ability to characterize changes in oxygen per unit area a nd per unit time allow rates of change in metabolism to be estimated further strengthening this techniques’ ability to document ecosystem responses. An example of the daily cycle in dissolved oxygen (DO) is shown in Figure 2, which283 shows average daily DO change at Silver Spri ngs, Florida over three different time intervals separated by about 50 years.

0 1 2 3 4 5 6

0 2 4 6 8 10 12 14 16 18 20 22
Time (hr)DO (mg/L) WSI Average (2004 - 05) Knight Average (1978 - 79) Odum Average (1952-54)

WSI (2004-05) = 19.8 Knight (1978-79) = 21.6 Odum (1952-55) = 15.5

Figure 2 – Comparison of Silver Springs average diurnal oxygen curves at the 1,200-m downstream station with the resultant estimated gr oss primary productivity in units of ppm-hrs. Higher total GPP estimates for Knight (1978-79) and WSI (2004-05) are a result of increased wetted area in Silver Springs Run above the 1,20 0-m station due to the addition of the Back Channel Extension in the late 1970s. Estimated G PP rates corrected for the total aquatic area declined from about 15.7 to 15.6 g O 2 /m 2 /d in the 1950s and the late 1970s, to about 11.2 g O 2 /m 2 /d during 2004/2005 (from Munch et al. 2006).

Gross and Net Productivity

Odum (1957a) measured nitrate and gross primar y productivity (GPP) in eleven Florida springs in 1955, which reveal a weak but negative relati onship between nitrate concentration and GPP (R 2 = 0.14). The recent findings of lower GPP in Silver Springs at higher nitrate nitrogen concentrations are consistent with the patterns re vealed by Odum’s earlier research (Table 1). A more recent study of community metabolism has b een completed as part of the Wekiva River Pollutant Load Reduction Goal evaluation conduc ted by the St. Johns River Water Management District, in which spring runs from the Wekiva River System with high nitrate levels were compared to spring runs with minimal nitrate le vels (WSI 2007a). The Wekiva River and Rock Springs Run were found to have higher nitrate nitrogen concentrations (averaging 0.69 and 0.84284 mg/L respectively) compared to the reference streams located in the Ocala National Forest (Juniper Creek – 0.05 mg/L and Alexander Springs Creek – 0.02 mg/L). Within the spring runs, the upstream nitrate concentrations in the Wekiva River and in Rock Spri ngs were significantly higher than downstream concentr ations (Wekiva – 1.1 vs. 0.36 mg/L; Rock Springs Run – 1.1 vs. 0.55 mg/L, respectively). Gross primary productiv ity and PAR efficiency were found to be significantly lower in all of these spring-fed stre ams with higher nitrate n itrogen concentrations (Figure 3 and Table 2, WSI 2007a). This trend is consistent with th e observations made at Silver Springs during the fifty-year retr ospective study described by Munch et al. (2006).

Table 1. A comparison of gross primary productivit y (GPP) over a fifty-year time period for Silver Springs (from Munch et al. 2006). Odum (1957b) Knight (1980) Munch et al. (2006)

Date GPP (gO 2 /m 2 /d) Date GPP (gO 2 /m 2 /d) Date GPP (gO 2 /m 2 /d) 2/19/1953 12.4 8/31/1978 19.3 Feb-04 8.2 3/7/1953 14.0 10/5/1978 13.6 Mar-04 11.4 3/25/1953 17.5 12/13/1978 7.8 Apr-04 13.2 1/7/1954 10.1 3/7/1979 10.7 May-04 13.9 5/23/1954 24.4 4/15/1979 16.8 Jun-04 12.7 7/12/1955 12.1 5/16/1979 23.4 Jul-04 13.6 8/11/1955 19.7 6/19/1979 20.7 Aug-04 12.3 7/17/1979 11.2 Sep-04 10.9 8/15/1979 17.1 Oct-04 11.7 Nov-04 9.8 Dec-04 8.5 Jan-05 8.6 Feb-05 11.1 Mar-05 10.8 Average 15.7 15.6 11.2285 y = -3.8294x + 5.7311
R 2 = 0.4286

0 2 4 6 8 10 12 14

0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 NOx-N (mg/L)Gross PrimaryProductivity (g O2/m2/d)

WR-1
WR-2
RSR-1
RSR-2
ASC-1
JC-1

y = -3.4485x - 1.5075
R 2 = 0.1463

-20 -15 -10 -5 0 5 10

0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 NOx-N (mg/L)Net PrimaryProductivity (g O2/m2/d)

WR-1
WR-2
RSR-1
RSR-2
ASC-1
JC-1

y = -0.3993x + 7.2451
R 2 = 0.0038

0 2 4 6 8 10 12 14 16 18

0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 NOx-N (mg/L)Community Respiiration(g O2/m2/d) WR-1
WR-2
RSR-1
RSR-2
ASC-1
JC-1

y = -0.8162x + 1.1131
R 2 = 0.2797

0.0 0.5 1.0 1.5 2.0 2.5 3.0 3.5 4.0

0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 NOx-N (mg/L)P/R Ratio WR-1
WR-2
RSR-1
RSR-2
ASC-1
JC-1

y = -0.2031x + 0.355
R 2 = 0.2882

0.0 0.1 0.2 0.3 0.4 0.5 0.6 0.7 0.8

0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 NOx-N (mg/L)Ecological Efficiency(g O2/mol) WR-1
WR-2
RSR-1
RSR-2
ASC-1
JC-1

Figure 3 – The relationship between nitrate and community metabolism parameters from six spring run segments, ASC is Alexander Springs Cr eek, JC is Juniper Creek, RSR is Rock Springs Run (1 is upstream, 2 is downstream), WR is We kiva River (1 is upstream, 2 is downstream). Eight sampling events for each spring segment (from WSI 2007a).286 Table 2 . Summary comparison of ecological indices between the spring run segments from Wekiwa Springs and Rock Springs and the refe rence spring runs from, Alexander Springs and Juniper Springs (from WSI 2007a). Wekiwa Springs Rock Sp rings Reference Springs Parameter Units SEG 1 SEG 2 Average SEG 1 SEG 2 Average Alexander Springs Juniper Springs Average

Dissolved Oxygen mg/L 1.93 6.81 4.37 5.67 6.40 6.03 5.54 6.81 6.17

pH s.u. 7.47 7.71 7.59 7.84 7.67 7.76 7.97 7.82 7.90 Specific Conductance μS/cm 316 426 371 242 257 250 898 1,645 1,272

Total Nitrogen mg/L 1.20 1.36 1.28 1.31 1.15 1.23 0.25 0.17 0.21

Nitrate + Nitrite N mg/L 1.01 0.364 0.686 1.13 0.554 0.841 0.051 0.024 0.038

Total Phosphorus mg/L 0.130 0.118 0.124 0.095 0.101 0.098 0.049 0.024 0.037

Soluble Reactive Phosphorus mg/L 0.120 0.105 0.112 0.084 0.084 0.084 0.041 0.016 0.028

Community GPP g O 2 /m 2 /d 2.12 5.39 3.76 1.41 1.00 1.20 5.71 5.67 5.69

Community NPP g O 2 /m 2 /d -3.50 -3.78 -3.64 -5.85 -8.59 -7.22 -0.39 -0.52 -0.46

Community Respiration g O 2 /m 2 /d 5.62 9.18 7.40 7.26 9.58 8.42 6.10 6.19 6.15

P/R ratio unit less 0.44 0.63 0.54 0.23 0.11 0.17 1.16 0.96 1.06

Community GPP Efficiency g O 2 /mol 0.22 0.29 0.26 0.09 0.09 0.09 0.31 0.45 0.38

Community GPP Efficiency % 6.06 3.57 4.82 1.87 3.63 2.75 1.82 3.05 2.44

Factors influencing stream metabolism have been estimated in eight streams from several biomes in North America by Mulholland et al. (2001). These streams were not spring runs, but were relatively free from human disturbance and had di ssolved inorganic nitroge n concentrations less than 0.15 mg/L and soluble reactive phosphorus con centrations of 0.014 mg/L or less. These streams had GPP estimates ranging from less than 0.1 to 15 g O 2 /m 2 /d, community respiration (CR) estimates ranging from about 2.0 to 11.1 g O 2 /m 2 /d, and photosynthetically active radiation (PAR) efficiencies ranging from 0.042 to 0.45 g O 2 /mol. Mullholland et al. (2001) reported that GPP and NPP were not correlated with either dissolved inorganic nitr ogen or soluble reactive phosphorus concentrations but we re significantly correlated with PAR, although CR was correlated with soluble reactiv e phosphorus concentrations.

Other estimates of primary production from spri ng ecosystems include those made by Duarte and Canfield (1990a), who published plant community and productivity data fo r thirty-one Florida springs. Productivity data were not directly co mparable to data collected by Odum (1957a, b), Knight (1980), WSI (2006), or Munch et al. (2006) as they were ba sed on a short term, rapid assessment technique rather than on a full diur nal cycle. Still Duarte and Canfield (1990a)287 observed no correlation between total nitrogen (which was mostly in the form of nitrate) or total phosphorus with the biomass or productivity of submersed aquatic vegetation (SAV) in the spring runs they examined. SAV standing crop and maximum daily produc tivity were correlated with degree of shading by shore line vegetation. These results sugge st that light may have been the primary limiting fact or during their study.

Respiration

Changes in community respiration are likely to be influenced by nutrie nts if nutrients are influencing the abundance or pr oductivity of primary producers. Changes in primary producers would in turn affect the consumer or decompos er community and hence the respiration of this part of the ecosystem. The practical differe ntiation of autotrophic versus heterotrophic respiration is not feasible in field studies. From Florida spring studies that have compared community respiration rates to water column nutr ient concentrations, it appears that a negative correlation does exist ( i.e. Munch et al. 2006; Figure 3, WSI 2007a), but the correlation is weak. Community respiration was also observed to in crease downstream in the Wekiva River and in Rock Springs Run as nitrate concentrations decr eased. Community respiration also declined at Silver Springs at the same time nitrate ni trogen concentrations were increasing (Munch et al. 2006).

Productivity/Respiration Ratio

The productivity/respiration (P/R) ratio is comm only used to classify ecosystems as either autotrophic (P/R > 1) or heterotr ophic (P/R < 1) depending on wh ether they are net producers or net consumers of organic matter (Odum 1956). Florida spring ecosystems were noted by Odum (1956) to have P/R ratios greater than one, indicating net production of organic matter. Estimates from Silver Springs illustrate that during the winter of 1952/1953 and March 1954, this ecosystem had P/R ratios of 2.9 and 7.0 resp ectively. The effects of elevated nitrate concentrations on the P/R ratio are revealed by the results of Munch et al. (2006), in that periodof-record (February 2004 to March 2005) average P/ R ratio at Silver Springs was estimated as 1.06, with the monthly maximum P/R ratio be ing 1.26 in June 2004 and a minimum monthly average of 0.73 in February 2004. In a comp arison of spring run segments with nitrate concentrations ranging from natural ambient leve ls (< 0.05 mg/L) to el evated levels (> 1.0 mg/L), P/R ratios were found to decline significantly in re lation to increasing nitrate concentrations (Figure 3, WSI 2007a). This ne gative correlation suggests that the P/R ratio can be used to evaluate nutrient impairment of spring ecosystems.

Photosynthetic Efficiency

The efficiencies of photosyntheti c primary producers have been me asured most comprehensively at Silver Springs, where Odum (1957b) reported an average value of 1.09 g O 2 /mol of PAR (ranging from 0.79 to 1.53), while Knight (1980) reported an average value of 1.06 g O 2 /mol of PAR (ranging from 0.61 to 1.50), and Munch et al. (2006) reported an average value of 0.95 g O 2 /mol of PAR (ranging from 0.07 to 2.71). The Munch et al. (2006) period-of-record average photosynthetic efficiency of 0.95 g O 2 /mol, is equivalent to an estimated PAR efficiency of 7.6% using estimated conversion factors and PAR efficiency was observed to vary seasonally with the288 highest monthly average value of 10.5% (1.30 g O 2 /mol) measured in December 2004 and the lowest value of 5.5% (0.68 g O 2 /mol) measured in June 2004. These estimates of photosynthetic efficiency at Silver Springs indicate a slight reduction in the magnitude of this parameter over the past 50 years although differences in the numbe r of estimates make it difficult to determine if this change is significant and if it has any relation to increasing nitrate concentrations. Photosynthetic efficiency estimates for other spring ecosystems have recently been made for Alexander, Juniper, Rock, and Wekiwa Springs with average values of 0.32, 0.45, 0.09, and 0.26 g O 2 /mol of PAR respectively (Figure 3, WSI 2007a). These estimates were derived from a similar number of daily measures collected during the same time periods and illustrate a reduction in photosynthetic effici ency correlated with increasi ng nitrate concentrations.

Community Metabolism Conclusions

As the previous examples illustrate, traditional measures of describing community structure which use biomass metrics allow few generalizati ons about the response of primary producers to elevated nutrient conditions in stream systems. This dilemma supports the use of ecosystemlevel metabolism functional measures, such as GPP, NPP, P/R ratio and photosynthetic efficiency, which integrate the overall functional response of th e ecosystem components and may reduce variance caused by site-specific conditions. The observations summarized in this chapter of reduced primary productivity an d ecological efficiency in corre lation with increasing nutrients may have multiple explanations.

First, Florida’s spring run ecosystems were likel y adapted to naturally lower ambient nutrient conditions and higher nutrient leve ls could represent an imbalanc e or stress; this explanation follows the subsidy-stress hypothesis of Odum et al. (1979). Support for this concept was provided in an ecosystem-level metabolism study of Silver Springs, where ecosystem production and efficiency were also reduced in inverse proportion to rising nitrate nitrogen concentrations (Munch et al. 2006).

A second possible mechanism that might help to explain a decline in ecosystem primary production could be reduced macrophyte photosynthesi s resulting from increased epiphytic algae growth and shading as a result of elevated wate r column nutrients. While epiphytic growths can reduce the light energy available to host plants, some research found these effects were primarily observed at low light levels and did not affect host photosynt hetic rates at typical midday intensities As spring waters are generally clear and light transm ission is excellent (Duarte and Canfield 1990b), epiphytic reductions in light availability to vasc ular plants ma y not reach the level of causing host plant death. In an experime ntal manipulation of epi phytes and light levels, Asaeda et al . (2004) noted that both epiphytic algae and low light reduce gr owth and production; plants tended to optimize low light-created shade by changing their physiology and morphology, but were unable to do the same when epiphytic al gae-created shade occurred on the leaf surface.

A third alternative explanation to reduced pr imary productivity in macr ophytes in relation to elevated nitrate concentrations was provided by Boedeltje et al. (2005) who reported high water column nitrate nitrogen concentrations ca n significantly reduce the growth of ammoniapreferring species such as Potamogeton alpinus , because the uptake and reduction of nitrate has a much higher energy and carbon requirement th an ammonia uptake a nd assimilation (Runge289 1986, Marschner 1995). Boedeltje et al. (2005) hypothesized that nitrate-dominated nitrogen assimilation may lead to strong me tabolic disturbances in specie s adapted to ammonia uptake by the roots (such as P. alpinus ) resulting in the observed reduction in growth under high water column nitrate concentrations , although further experiments ar e needed to a ssess potential negative effects on species found in the springs of Florida.

Although the causative mechanisms wh ich result in reduced primar y productivity an d ecological efficiency in correlation with increasing water column nutrient concentrations (especially nitrate and perhaps phosphorus) remain unknown; the aver ages, extremes, and seasonal patterns for these ecological measures are becoming more apparent as additional ecosystem metabolism data are being collected in Florida’s springs. Year -to-year measures of ecosystem metabolism are remarkably consistent compared to structural measures such as plant or fish populations and provide a valuable database for comparison to ot her springs, spring runs, a nd streams in Florida.

HUMAN AND AESTHETIC USES

Visitor Satisfaction

There are little or no published data that relate recreational satisfaction from Florida springs to their trophic status. It is readil y apparent that people prefer to sw im in clear, clean water rather than turbid, colored, or phytoplan kton-rich waters. It is likel y that highly abundant macrophytes discourage water recreation as well . This presents a dilemma, in that, native benthic vascular plants represent a natural and ecologically impo rtant component of spring ecosystems. Species such as Vallisneria americana or Sagittaria kurziana are typically not viewed as nuisance solely due to their presence.

Economic Impacts

As detailed in Chapter 1, the number of vis itors attending the springs in Florida and their cumulative economic impact is substantial enough to impact local economies. If recreational attendance were to decline due to real or percei ved diminished appeal of a spring system, there would be associated declines in the economic impacts genera ted. Preservation of spring ecosystems should not be based solely on an economic basis, unle ss considerations for environmental services are factored in as well, but they provide support towards these systems protection. At this time insufficient quantitativ e public use data exist from most springs to develop correlations with incr easing nutrient concentrations.

Effects of Eutrophication on Human Recreation

The type of recreation will obviously influence to lerance for the presence of SAV, with water dependent recreation perhaps having lower tole rance levels than non water dependent contact activities. SAV can attract waterfowl, wading birds, and other wildlife which would affect perceptions of persons engaged in observation of wildlife as well. Limited data on recreational satisfaction in relation to primary producers exis ts, but Holland and Cichra (1994) reported that what respondents liked best abou t their trip to Rainbow River wa s: clear, clean water (35%), scenery (19%), quiet, relaxing (17%), and seeing wildlife a nd fish (11%) and that what290 respondents liked least about their trip: rainy we ather (13%), motorboat noise (12%), vegetation and algae in water (10%), and river float too long (10%). These results suggest that while water clarity is very important, there is tolerance for submersed aqua tic vegetation among the surveyed recreationalists, possibly because this vegetation may have been viewed as a natural component of a healthy ecosystem. Declines in water clarity as a result of reductions in stream flow with corresponding increases in water residence time and phytoplankton abundance would likely be associated with decreased recr eational satisfaction. Similarly, increases in the abundance of vascular plants or macroalgae are likely to be viewed negatively by recreational us ers of spring ecosystems, given that people do not like to sw im through or stand in vegetation. If the abundance of filamentous algae increases as a result of eutrophication in a spring ecosystem, there would likely be a decline in recreational user satisfaction. Real or pe rceived declines in the aesthetics of a spring ecosystem as a result of eutrophication would likely result in less human recreational use and visitation, with resulting declines in the economic impact. It is important to note that human uses of the terms, such as natural or healthy , are subjectively defined, and that an educational component aimed to identify th ese biases would serve both the managers and users of springs and other ecosystems.

DISCUSSION

In the past century, many of Florida’s spring ecosystems have experienced changes which include declines in discharge, increases in conductivity, increased recr eation, and colonization by non-indigenous species. However, the widespread increase in nitr ate concentrations during this time period continues to be one of the key elements that spring ecosystems have had to adjust to, and it has been hypothesized that the increasing concentration of that form of nitrogen is a principal causative factor fo r observed changes in ecosystem metabolism. Strong (2004) examined temporal trends in wa ter quality characteristic s in 109 Florida springs. Parameters that were generally found to be in creasing were specific conducta nce, alkalinity, hardness, and nitrate. Only pH concentrations were observed to be declining. There were no apparent trends in phosphorus concentrations. Mean nitrate concentrations in sp ring samples generally increased from about 0.43 mg/L before 1977 to about 1.13 mg/L since 1990. Strong’s (2004) study suggests that state-wide trends for declinin g water quality of springs are occurring. An illustration showing the range of nitrate plus nitrite concentrations (NO x -N mg/L, although nitrite is generally a minimal component of these a ggregate values) for 130 Fl orida springs compiled from 2006 Florida Geological Survey data is shown in Figure 4. This figure presents the wide range of NO X -N concentrations that Florida spring s exhibit and suggest s that 83% of the surveyed springs have NO X -N concentrations greater than 0.1 mg/L, which is a commonly assumed background level for groundwater nitrate levels in Florida.291 0.0 0.5 1.0 1.5 2.0 2.5 3.0 3.5 4.0 4.5 5.0 5.5

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Min 0.004 Max 5.10 Spring ID Spring ID Spring ID Beecher Spring 1 Fenney Spring 45 Cedar Head Spring 89 Copper Spring 2 Hornsby Spring 46 Wakulla Tubing D-Tunnel 90 Green Cove Spring 3 Horn Spring 47 Gum Spring Main 91 Holton Creek Rise 4 Cypress Spring 48 Little River Spring 92 Orange Spring 5 Starbuck Spring 49 Wakulla Tubing C-Tunnel 93 Suwannee Springs 6 Turtle Spring 50 Double Spring 94 Newport Spring 7 Salt Spring (Hernando) 51 Jackson M ill Pond Spring 95 Waldo Spring 8 Hunter Spring 52 Wakulla Tubing B-Tunnel 96 Warm Mineral Spring 9 Ellaville Spring 53 Rainbow Spring #6 97 Santa Fe Spring (Columbia) 10 Falmouth Spring 54 Silver Spring Main 98 Nutall Rise 11 Alapaha River Rise 55 Hart Springs 99 Alexander Springs 12 Wacissa Springs #2 56 Otter Spring 100 Silver Glen Springs 13 Wakulla Tubing A/K-Tunnel 57 Bubbling Spring 101 Washington Blue Spring (Choctawhatchee) 14 Black Spring (Jackson) 58 Wekiwa Springs 102 Steinhatchee River Rise 15 Wakulla Tubing K-Tunnel 59 Rainbow Spring #1 103 Santa Fe River Rise (Alachua) 16 Wakulla Tubing A/D-Tunnel 60 Deleon Spring (Volusia) 104 Welaka Spring 17 Holmes Blue Spring 61 Ginnie Spring 105 Williford Spring 18 Jackson Blue Hole Spring 62 Reception Hall Spring 106 Juniper Springs 19 Homosassa #1 63 Devils Ear Spring (Gilchrist) 107 Columbia Spring 20 Sanlando Springs 64 Blue Grotto Spring 108 Fern Hammock Springs 21 Suwanacoochee Spring 65 Madison Blue Spring 109 Treehouse Spring 22 Gator Spring (Hernando) 66 Rainbow Spring #4 110 Salt Springs (Marion) 23 Homosassa #2 67 Rock Springs 111 Crays Rise 24 Chassahowitzka Spring Main 68 Mearson Spring 112 Washington Blue Spring (Econfina) 25 Citrus Blue Spring 69 Gilchrist Blue Spring 113 Big Spring (Big Blue Spring) (Jefferson) 26 Levy Blue Spring 70 Manatee Spring 114 Morrison Spring 27 Mission Spring 71 Sun Springs 115 Spring Creek Rise #1 28 Allen Mill Pond Spring 72 Telford Spring 116 St. Marks River Rise (Leon) 29 Chassahowitzka Spring #1 73 Owens Spring 117 Gainer Spring #1C 30 Magnolia Spring 74 Troy Spring 118 Spring Creek Rise #2 31 Homosassa #3 75 Buckhorn Spring 119 Tarpon Hole Spring 32 Rock Bluff Springs 76 Crystal Springs 120 Gainer Spring #3 33 Bugg Spring 77 Lafayette Blue Spring 121 Poe Spring 34 Blue Hole Spring (Columbia) 78 Running Springs 122 Shepherd Spring 35 Springboard Spring 79 Unknown 11365 123 Brunson Landing Spring 36 Volusia Blue Spring 80 Lithia Springs Major 124 Gainer Spring #2 37 Branford Spring 81 Hays Spring (Jackson) 125 Ponce De Leon Springs 38 Siphon Creek Rise 82 Jackson Blue Spring 126 Sulphur Spring (H illsboro ugh) 39 Little Springs (Hernando) 83 Shangri-La Springs 127 Rhodes Spring #1 40 Weeki Wachee Main Spring 84 Lafayette Ruth Spring 128 Rhodes Spring #2 41 Ichetucknee Head Spring (Suwannee) 85 Fanning Springs 129 Rhodes Spring #4 42 Guaranto Spring 86 Apopka Spring 130 Natural Bridge Spring 43 Wakulla Spring 87 Beckton Springs 44 Gadsen Spring 88

Figure 4 – Illustrates the average NO X

-N concentrations (mg/L) for 130 Florida springs (data from Florida Geological Survey, Scott et al. 2004).

A similar illustration showing the range of tota l phosphorus concentrations (TP mg/L), for 130 Florida springs compiled from 2006 Florida Geol ogical Survey data is shown in Figure 5. Although total phosphorus (the aggreg ate of all forms of phosphorus ) data are presented, in the case of groundwater, this total phosphorus is almo st entirely composed of dissolved inorganic forms of phosphorus such as orthophosphate or soluble reactive phos phorus. This figure suggests that the total phosphorus concentrations of spring water are less variable than the wide range of NO x -N concentrations that Florida sp rings exhibit. Groundwater phosphorus concentrations are largely a function of th e associated geology and many of the higher phosphorus values are observed from springs which overlie natural phosphatic deposits.

Of these 130 springs, 80% have total phosphorus concentrations greater than 0.025 mg/L, a value which has been defined as a break point sepa rating mesotrophic and eutrophic lakes (Table 3, Forsberg and Ryding 1980). Although othe r indices of trophic status exist ( e.g. the Florida Trophic State Index created by the Florida Depa rtment of Environmen tal Protection), their application to lotic systems has been problematic. This is principally due to the index’s inability to incorporate macrophytes and benthic algae th at are not accounted for by standard methods which measure water column chlorophyll content in non-flowing systems.292 Because phosphorus levels have not been incr easing in Florida’s springs, even though the groundwater is regionally rich in phosphorus, (Str ong 2004), it can be suggested that the majority of Florida springs would have been historically nitrogen limited, based on known ratios of nitrogen to phosphorus demand by plants ( e.g. Redfield ratio, Redfield et al. 1963). However, in the last several decades, a great majority of springs appear to have been impacted by high nitrate concentration (83% exceed 0.1 mg/L NO x ) and the resulting nitrogen to phosphorus ratios have become skewed to favor phosphorus limitation. Th is view should not be misinterpreted by the reader to suggest that nitrate is not important because it may no longer be limiting on a ratio basis. Instead it should be viewed that there is such an over abundance of nitrogen (principally in the form of nitrate) rela tive to historic ambient conditi ons, that an un-natural phosphorus limited state has evolved. While efforts to limit the input of phosphorus to aquatic ecosystems should be continued, the supply of this element to Florida’s springs appears to be linked to geological conditions which ar e largely beyond human contro l. Nitrate loading from anthropogenic activities can, however, be moderate d and hence represents our best ability to effect a nutrient reductio n strategy to benefit Flor ida spring ecosystems.

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Min 0.008 Max 1.90 Spring ID Spring ID Spring ID Spring ID Spring ID Weeki Wachee Main Spring 1 Holmes Blue Spring 27 Wakulla Tubing C-Tunnel 53 Rhodes Spring #4 79 Otter Spring 105 Little Springs (Hernando) 2 Ichetucknee Head Spring (Suwannee) 28 Juniper Springs 54 Wacissa Springs #2 80 Welaka Spring 106 Guaranto Spring 3 Chassahowitzka Spring Main 29 Wakulla Tubing B-Tunnel 55 Nutall Rise 81 Hart Springs 107 Gainer Spring #3 4 Gainer Spring #1C 30 Gilchrist Blue Spring 56 Natural Bridge Spring 82 Rock Springs 108 Washington Blue Spring (Econfina) 5 Morrison Spring 31 Mearson Spring 57 Reception Hall Spring 83 Bugg Spring 109 Gainer Spring #2 6 Hunter Spring 32 Troy Spring 58 Blue Hole Spring (Columbia) 84 Crays Rise 110 Gator Spring (Hernando) 7 Homosassa #3 33 Tarpon Hole Spring 59 Crystal Springs 85 Siphon Creek Rise 111 Green Cove Spring 8 Manatee Spring 34 Spring Creek Rise #2 60 Lafayette Blue Spring 86 Fenney Spring 112 Hays Spring (Jackson) 9 Ponce De Leon Springs 35 Falmouth Spring 61 Alexander Springs 87 Sulphur Spring (Hillsborough) 113 Magnolia Spring 10 Unknown 11365 36 Bubbling Spring 62 Branford Spring 88 Poe Spring 114 Salt Springs (Marion) 11 Brunson Landing Spring 37 Spring Creek Rise #1 63 Steinhatchee River Rise 89 Wekiwa Springs 115 Black Spring (Jackson) 12 Gum Spring Main 38 Copper Spring 64 Mission Spring 90 Alapaha River Rise 116 Gadsen Spring 13 Homosassa #2 39 Running Springs 65 Lithia Springs Major 91 Holton Creek Rise 117 Springboard Spring 14 Rainbow Spring #6 40 Wakulla Tubing A/D-Tunnel 66 Ellav ille Spring 92 Suw annee Springs 118 Washington Blue Spring (Choctawhatchee) 15 Fern Hammock Springs 41 Madison Blue Spring 67 Telford Spring 93 Starbuck Spring 119 Newport Spring 16 Salt Spring (Hernando) 42 Blue Grotto Spring 68 Sun Springs 94 Big Spring (Big Blue Spring) (Jefferson) 120 Warm Mineral Spring 17 Wakulla Spring 43 St. Marks River Rise (Leon) 69 Deleon Spring (Volusia) 95 Sanlando Springs 121 Williford Spring 18 Rainbow Spring #1 44 Lafayette Ruth Spring 70 F anning Springs 96 Rainbow Spring #4 122 Jackson Mill Pond Spring 19 Jackson Blue Hole Spring 45 Levy Blue Spring 71 Suwanacoochee Spring 97 Santa Fe Spring (Columbia) 123 Jackson Blue Spring 20 Shangri-La Springs 46 Rhodes Spring #2 72 Volusia Blue Spring 98 Treehouse Spring 124 Double Spring 21 Wakulla Tubing D-Tunnel 47 Devils Ear Spring (Gilchrist) 73 Hornsby Spring 99 Santa Fe River Rise (Alachua) 125 Turtle Spring 22 Wakulla Tubing K-Tunnel 48 Silver Spring Main 74 Owens Spring 100 Columbia Spring 126 Chassahowitzka Spring #1 23 Apopka Spring 49 Allen M ill P ond Spring 75 Shepherd Spring 101 Silver Glen Springs 127 Little River Spring 24 Cedar Head Spring 50 Ginnie Spring 76 Rock Bluff Springs 102 Homosassa #1 128 Cypress Spring 25 Citrus Blue Spring 51 Horn Spring 77 Waldo Spring 103 Beecher Spring 129 Beckton Springs 26 Wakulla Tubing A/K-Tunnel 52 Rhodes Spring #1 78 Orange Spring 104 Buckhorn Spring 130

Figure 5 - Illustrates the average to tal phosphorus concentrations (mg TP/L) for 130 Florida springs (data from Florida Geological Survey, Scott et al. 2004).293 Table 3 . Lake trophic state classification syst em based on Forsberg an Ryding (1980).

Trophic State Chlorophyll (mg/L) Total Phosphorus (mg/L) Total Nitrogen (mg/L) Water Clarity (ft) Oligotrophic < 0.003 < 0.015 < 0.40 > 13.1 Mesotrophic 0.003 – 0.007 0.015 – 0.025 0.40 – 0.60 8.2 - 13.1 Eutrophic 0.007 – 0.04 0.025 – 0.10 0.60 – 1.50 3.3 – 8.2 Hypereutrophic > 0.04 > 0.10 > 1.50 < 3.3

The best evidence suggesting a decline in the h ealth of Florida’s spring ecosystems comes from studies that have been conducted over a half-century time period in Silver Springs. Both the direct measurements and the estimated system metabolism analyses indicate that the Silver Springs ecosystem may be considerably less productive than it was fifty years ago (Munch et al. 2006). This appears to be a consequence of lowe red gross primary productivity and efficiency of light utilization. This result appears to be count er-intuitive to the observa tion that nutrient levels (nitrate nitrogen) which support plant growth have increased in Silver Spri ngs. It is generally thought that an increase in nutrients will stimulate both gross and net primary productivity in aquatic ecosystems (Wetzel 2001). However, spring ecosystems are characterized by flowing waters which continually re-supply nutrient dema nds of primary producers. H.T. Odum (1957a) theorized that prior to modern anthropogenic impacts, Florida’s springs existed as a balanced aquatic ecosystem which had evolved to maxi mize community metabolism, and the relatively recent increase of nutrients in the form of nitr ate may have resulted in a stress rather than a subsidy (E.P. Odum et al. 1979). The subsidy-stress hypothe sis may be the best model to describe the response of spring ecosystems to increasing nitrate c oncentrations; however, controlled field experiments n eed to be made in order to test this hypothesis.

One of the most comprehensive examinations of the effects of nitrate on spring ecosystem structure and function is the 50-year retr ospective study of Silver Springs by Munch et al. (2006). The key findings from th at study are reproduced here:

• Sagittaria kurziana remains the dominant submersed aquatic plant species in Silver Springs and represents one of the main physical features of the ecosystem.

• Biomass estimates for submersed aquatic plants in the summer season were not significantly different from estimates made by Odum in the early 1950s. However, estimates for winter biomass were 31% lo wer than Odum’s, who reported no seasonal difference in submersed aquatic plant biomass.

• Biomass estimates for the epiphyte community in the summer were approximately threefold higher than those reported by Odum, while winter values were not significantly different between the two studies.

• The largest disparity between the Munch et al. (2006) estimates of primary producer community biomass and those of Odum from the 1950s was the substantial increase in biomass for the benthic algal mat community. While Odum discounted the importance of algal mats, in terms of biomass, the Munch et al. (2006) study indicated that this primary producer category had biomass estimates simi lar to those observed for epiphytes and294 submersed aquatic plants. However, it is impor tant to note that these estimates may not be an accurate indicator of primary productiv ity attributable to benthic mats, since mat biomass includes large proportions of bacter ia, fauna, dead algae, and other detrital material.

• Total species richness for birds, fish, and reptiles in the Munch et al. (2006) study were similar to historical r ecords at Silver River.

• Estimated annual average fish li ve-weight biomass in the Munch et al. (2006) study has declined in Silver Springs since Odum’s study in the early 1950s by about 96%; and by 61% since Knight’s 1978-79 st udy (Knight 1980). These d eclines in biomass were primarily due to large reductions in a few sp ecies (channel catfish, mullet, and gizzard shad), while the remaining fish species were found in similar abundance across the fiftyyear span. It is likely that declines in channel catfish, mullet, and gizzard shad abundance are the result of diminished access to the St . Johns River and the Atlantic Ocean due to the Kirkpatrick (AKA Rodman) Dam

• Annual average gross primary productivity (GPP) declined from about 15.6 grams oxygen per square meter per day (g O 2 /m 2 /d) in the 1950s and late 1970s to about 11.2 g O 2 /m 2 /d during the current study, a decline of about 27%.

• Community respiration also declined from about 14.8 g O 2 /m 2 /d during the earlier studies to about 10.9 g O 2 /m 2 /d during the Munch et al. (2006) study, a 26% reduction.

• The resulting net community primary pr oductivity declined from about 1.0 g O 2 /m 2 /d in the 1950s, to 0.80 g O 2 /m 2 /d in the late 1970s, to about 0.42 g O 2 /m 2 /d during the Munch et al. (2006) study, a decline of about 59% over the past fifty years.

• The production to respiration (P /R) ratio remained relatively consistent between the three studies, ranging from about 1.11 during Odum ’s study in the 1950s to about 1.06 during Knight’s study in the late 1970s, and 1.06 in the Munch et al. (2006) study.

• Ecological efficiency declined from about 1.09 gram of oxygen per mol (g O 2 /mol) of Photosynthetically Active Radiation (P AR) during Odum’s study to about 0.94 g O 2 /mol of PAR during the Munch et al. (2006) study, a dec line of about 13%.

• Average particulate export rates were found to be 72% lowe r during the Munch et al. (2006) study compared to data published by Odum (1957). These particulates are a representation of the net syst em productivity, and their decline strengthens the evidence that overall productivity has declined in Silver Springs.

The Munch et al. (2006) study observed that the percenta ge of developed lands within the springshed (5 mile radius) was positively correlated with actual nitrate concentrations and spring discharge was found to be inversel y correlated with nitrate. The resulting co rrelation models were used in turn to predict future nitrogen loading to the Silver Springs ecosystem if the future land use development estimates for 2055 are realized (Table 4).295 Table 4 . Modeled changes in the nitrogen concen tration of Silver Springs water based on projected changes in land use and corresponding nitrogen load ing rates (Munch et al. 2006).

Year Observed Spring Flow MODFLOW Simulated Spring Flow Observed Nitrogen

Load (lbs/yr) Land Use/Land

Cover Model Est.

N Load (lbs/yr) Observed Spring N Concentration (mg/L) Land Use/Land Cover Model Est. N Concentration (mg/L) N Concentration Change (Model Predictions, %)
1957 640.0 716.5 94,416.0 399,054.10 0.10 0.38 N/A
1979 778.0 710.8 814,898.6 802,633.06 0.71 0.76 102.75
1995 720.0 708.3 955,962.0 1,036,198.93 0.90 0.99 29.56
2005 680.0 687.6 1,057,606.7 1,120,813.63 1.07 1.10 11.42
2055 N/A 687.6 N/A 1,760,000.00 N/A 2.02 84.00

The apparent decline in ecosystem metabolism and efficiency observed in Silver Springs over the past fifty years may be the result of a several ecosystem-level factors wo rking in concert or at odds. Four possible hypotheses were offe red to explain these observations (Munch et al. 2006):

• Decreased usable solar radiation is reaching the level of submersed aquatic plants in Silver Springs due to increased shading by a growing tree canopy (natural wetland forest succession) along the river and this reduced i nput of solar energy may have lowered GPP and PAR efficiency.

• GPP may have declined as areas of optim al submersed aquatic vegetation growth ( Sagittaria kurziana beds with adapted periphytic algae) have been replaced with benthic algal mats, possibly due to increasing nitrate nitrogen concentrations, flow decreases, or physical factors related to human uses.

• Decreases in flow rate and water velociti es due to natural climatic conditions or consumptive groundwater uses may have reduced the previous subsidy needed for maximum plant/periphyton growth a nd higher GPP and PAR efficiency.

• Decreased top-down consumer control of the primary producers and lowered GPP is resulting from lower fish/consumer populations below optimal grazing densities, possibly due to obstruction of fish migration by the Kirkpatrick Dam, or lower minimum daily dissolved oxygen concentrations, or due to indirect effects of nitrate on consumers. The role of striped mullet ( Mugil cephalus ) in the top-down control of the periphytic growths on submersed vascular plants in Silver Springs was reported by Allen (1946, p. 32), who wrote: “ These attractive fish are very prominent, always swimming about, forever feeding on algae off the blades of underwater grasses [ Sagittaria kurziana ] . They start at one end of the blade of grass, sucking up food as they slowly swim up toward the other end .”

Overall, the Munch et al. (2006) study indicates that multiple factors may have resulted in the observed ecosystem changes at Silver Springs ove r the past 50 years. These factors include reductions in the volume of spring discharge, in creases in riparian shading, reductions in the abundance of herbivorous fish, and increases in nitrate concentra tion and load. The authors of296 the Munch et al. (2006) study suggest that the corresponding increases in nitrate concentration over this time period cannot exclusively be respons ible for this degradation, but that nitrate pollution has likely contributed to this decline. Given the anticipated future increases in nitrate pollution, additional concern due to this factor re mains warranted. Support for the concept that nitrate can act as both a subsidy at low concentr ation and a stressor at high concentrations has been provided by the WSI (2007a) study in the Wekiva River and Rock Springs Run which observed negative correlations between measur es of ecosystem productivity and nitrate concentrations. Based on these lines of availabl e evidence, nitrate alone may not be the sole factor negatively impacting Florida spring ecosys tems, but it is the single most important water quality parameter that has changed during the pe riod of observed declines in spring ecosystem structure and function.

RECOMMENDATIONS FOR FURTHER RESEARCH

Despite a wide variety of spri ngs’ research, there remains a significant knowledge gap between the real and perceived threats that nitrate pollu tion plays on the ecology of spring ecosystems. This situation is not unique to Florida, as Bernot and Dodds (2005, p. 442) observed that for North American streams: “ There is a great need for long-ter m studies of nitrogen additions in lotic ecosystems and clear distinc tions need to be made between ecosystem responses to shortterm or periodic increases in nitrogen loading and alterations in ecos ystem functions due to chronic nitrogen loading. ” In addition to the general res earch questions to be answered, Florida’s springs represent impor tant cultural and economic reso urces which also provide realtime windows into the condition of Florida’s mo st essential natural resource – high-quality groundwater.

With these caveats in mind, we suggest that nutr ients are not the only stressors of concern in spring ecosystems. Recent work in the headwate r areas of Wekiwa and Rock springs indicates that ecosystem metabolism can be negatively imp acted by high levels of human recreation and by flow reductions (WSI 2007a, 2007b). Many, if not most, of Florida’s springs are threatened by these additional stressors as well as by in vasions from non-indigenous plant and animal species, aquatic weed management activity, erosi on and deposition of sediments, and increasing stormwater inflows from surrounding urbanized ar eas. Future research to be conducted should attempt to effectively understand the separate and synergistic effects of these multiple forcing functions and the associated res ponses of our spring ecosystems.

With the multiple stressors affecting spring ecosystems, any metric used to monitor their health should to the fullest extent possibl e be integrative. The single best way to characterize these systems would be through estimates of community metabolism and its constituent parameters: gross primary productivity, net primary productiv ity, community respira tion, productivity to respiration ratios, and photosynthetic efficiency. These metrics can be relatively easily estimated using upstream-downstream change s in dissolved oxygen concentr ations and have the best ability to describe ecosystem level functions, an d their value is enhanced when combined with measures of fauna and flora abundance and wa ter chemistry, other ty pical single-component ecosystem metrics.297 Proposed Research Priorities

The following springs’ research priorities are suggested:

• Recognize the importance of the collection of water quality samples, particularly for dissolved oxygen (diel changes) and nitrates, on a periodic schedule, from major springs. Continue the collection of discharge estimates from major springs (all first and second magnitude and a representative sample of springs with lower discharge rates), as declining discharge represents a major threat to springs. These types of data are regularly collected by local, state, and federal ag encies, but improvements in the spatial and temporal sampling could be realized as we ll as improved assembly, analysis, and reporting of the resulting data.

• Continue to expand and refi ne the development of Land Use and Land Cover (LULC) datasets, which in conjunction with improved infrared and color imagery, will greatly increase the relative accuracy in the LULC classifications for springsheds. In conjunction with accurate estimates of nitr ogen loading per LULC category, assessment of springshed land use changes on the nutrien t concentration of associated spring discharge could be made. Thus , rather than solely reacting to spring pollution after the source is in place, the ability to predict a nd as a result to limit or prevent springshed pollution would be enhanced.

• Better utilize current technology of in situ data loggers in a repr esentative set of major springs ecosystems so as to provide regul ar or continuous records of changing water quality conditions, including temperature, dissolved oxygen, pH, and specific conductance (and other parameters including nu trients as instruments and technologies become available). These measures, particular ly if made in conjunction with levels of ambient solar radiation and spring discha rge would allow the direct estimation of ecosystem metabolism and photosynthetic effici ency. Given the inter-specific range of physical and chemical conditions that Florida springs curre ntly exhibit, simultaneous estimates of ecosystem metabolism would a llow development of a response surface for comparison to changing environmental conditi ons (forcing functions). For example, multivariate analysis could be applied to de termining the relative importance of a number of forcing functions on ecosystem meta bolism, including identification of the concentration or intensity of each variable ( e.g. , nitrate) as it transitions from a subsidy to
a stressor.

• Quantify invertebrate, vertebra te, and floral diversity a nd abundance on a three-to-five year schedule at 1 st and 2 nd magnitude springs so that trends in these ecosystem components can be observed. This sampling s hould be conducted on springs that are also being monitored for ecosystem metabolism a nd should be scheduled on a monthly basis for at least one year during e ach three-to-five -year period so as to elucidate seasonal trends.

• Conduct in situ mesocosm experiments which quantif y the role that bottom-up processes ( i.e. nutrient limitation) vers us top-down processes ( i.e. herbivore grazing) have on the growth and abundance of primary producers, particularly epiphytes and filamentous algae. The primary question to be answered is whether future increases or decreases in nutrient concentrations will result in res ponses by the primary producer community.298 • Better monitor and inventory human use activitie s in a representative group of important springs on a routine basis so that optimal (as opposed to high) levels of human recreation can be maintained without degradation of th e springs’ resource. Human recreation levels or uses which are identified to be impairi ng spring health or degr ading spring aesthetics should be modified through adaptive manage ment strategies. The most successful example of this balance may be Ichetu cknee Springs State Park which annually accommodates hundreds of thousands of in wa ter recreationalists while maintaining sustainable populations of submersed macrophytes.

Recommendations

Based on the synthesis of information provided con cerning the effects of nutrients on springs in Florida, the following recommendations are provided:

• It is recommended that the Florida Depa rtment of Environmental Protection (DEP) Springs Initiative continue funding applied research under field c onditions in order to develop a consistent base of scientific knowledge for spring ecosystems. A critical need exists to establish a comprehensive baseline database for a large number of artesian spring ecosystems, including their ecosystem metabolism, trophic structure, and key forcing functions. A substantial increase in the annual funding for this research is justified by the ecological and economic importa nce of artesian springs in Florida.

• It is recommended that counties with significant springs’ resources utilize existing governmental agencies to develop springs prot ection plans that address such components as land development regulations and land acquisition, along with other groundwater and spring protection measures in order to prev ent further and, if po ssible, reverse the observed increases in nitrate-nitrogen c oncentrations. Regulatory programs and tax incentives should be utilized to reduce nitrogen loads to the ar tesian groundwater in karst areas of the state. These actions have the ability to reduce other ground water pollutants introduced through anthropogenic activities and would contribute to high quality ground water recharge which would be widely beneficial.

• It is recommended that the DEP with s upport from the Florida Legislature develop applicable standards that define nutrient impairment in artesi an springs. Springs that meet these impairment metrics should be moved as quickly as possible into the state’s Total Maximum Daily Load (TMDL) program. Nitr ate concentrations significantly above historical spring levels could be used as a criterion for lis ting under the Impaired Waters Rule. For springs that are above recomm ended thresholds for nitrate nitrogen concentrations, these TMDLs could provide a basis for enhanced nitrogen-removal standards for municipal and on-site wastewat er facilities as well as for stormwater management systems and the active and pro-ac tive enforcement of nitrogen mass loading limits by existing and future point and non-point sources in the aff ected springsheds.299 REFERENCES

Allan, J.D. 1995. Stream Ecology, Structure a nd function of running waters. Chapman and Hall. 388 pp.

Allen, E.R. 1946. Fishes of Silv er Springs, Florida. No publis her. Copyright E. Ross Allen, Silver Springs, Florida. 36 pp.

Asaeda, T., M. Sultana, J. Manatunge, and A. Ablimit. 2004. Colonization and growth of epiphytic algal communities on Potamogeton perfoliatus under two different light regimes. New Zealand Journal of Marine and Freshwater Re search 38: 585–594.

Bachmann, R.W., B.L. Jones, D.D. Fox, M.V. H oyer, L.A. Bull, and D.E. Canfield, Jr. 1996. Relations between trophic state indicators and fish in Florida (USA) lakes. Canadian Journal of Fisheries and A quatic Sciences 53: 842-855.

Bernot, M.R. and W.K. Dodds. 2005. Nitrogen retention, removal, and saturation in lotic ecosystems. Ecosystems 8: 442-453.

Boedeltje, G.A., J.P. Smolders, and J.G.M. Roel ofs. 2005. Combined effects of water column nitrate enrichment, sediment type and irradiance on growth and foliar nutrient concentrations of Potamogeton alpines . Freshwater Bi ology 50(9): 1537-1547.

Bothwell, M.L. 1985. Phosphorus limitation of lotic periphyton growth rates: An intersite comparison using continuous flow troughs (Thom pson River System, British Columbia). Limnology and Oceanography 30: 527-554.

Bristowe, J.M. and J. Whitcombe. 1971. The role of roots in the nutriti on of aquatic vascular plants. American Journal of Botany 58: 8-13.

Busch, D.E. 1978. Successional changes associated with benthic assembla ges in experimental stream. Ph.D. Thesis. Oregon State University. Corvallis, OR. 91 pp.

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Carignan, R. and J. Kalff. 1980. Phosphorus sour ces for aquatic weeds: water or sediments. Science 207: 987-988.300 Chambers, P.A., E.E. Prepas, M.L. Bothwell, a nd H.R. Hamilton. 1989. Roots versus shoots in nutrient uptake by aquatic macrophytes in flowi ng waters. Canadian Journal of Fisheries and Aquatic Sciences 46: 435-439.

Chen, C., D. Yin, B. Yu, and H. Zhu. 2007. Effect of Epi phytic Algae on Photosynthetic Function of Potamogeton crispus. Journal of Freshwater Ecology 22(3): 411-420.

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Holland, S.M. and C.E. Cichra. 1994. Human and environmental dimensions of the recreational use of Blue Run and Rainbow Springs State Park, Dunnellon, Florida. Final Report. Submitted to the Florida Department of Na tural Resources, Division of Recreation and Parks. 275 pp.

Howarth, R.W. 1988. Nutrient limitation of net primary produc tion in marine ecosystems. Annual Review of Ecology and Systematics 19: 898–910.

Hoyer, M.V. and D.E. Canfield, Jr. 1991. A phosphorus-fish standing crop relationship for streams? Lake and Reser voir Management 7(1): 25-32.

Hubbs, C.L. and E.R. Allen. 1943. Fishes of Silver Springs. Proceedings of the Florida Academy of Sciences 6: 110-130.

Hynes, H.B.N. 1970. The ecology of running waters . University of Toronto Press. Toronto.302 Knight, R.L. 1980. Energy Basis of Control in Aquatic Ecosystems. Ph.D. Dissertation. University of Florida. 200 pp.

Kurz, R.C., D.C. Woithe, S.K. Notestein, T.K. Frazer, J.A. Hale, and S.R. Keller. 2004. Mapping and Monitoring Submerged Aquatic Ve getation in Ichetucknee Springs – 2004. Final Report. Prepared for Suwannee River Water Management Dist rict. pp. variously numbered.

Marschner, H. 1995. Mineral Nutrition of Higher Plants. pp. 232–242. Academic Press, San Diego, CA.

Meade, M.E. and S.A. Watts. 1995. Toxicity of ammonia, nitrite, a nd nitrate to juvenile Australian crayfish, Cherax quadricarinatus . Journal of Shellfish Research 14: 341-346.

Mosisch, T.D., S.E. Bunn, P.M. Davies, and C.J. Marshall. 1999. Effects of shade and nutrient manipulation on periphyton grow th in a subtropi cal stream. Aquatic Botany 64(2): 167- 177.

Mulholland, P.J., C.S. Fellows, J.L. Tank, N.B. Gr imm, J.R. Webster, S.K. Hamilton, E. Martí, L. Ashkenas, W.B. Bowden, W.K. Dodds, W.H. McDowell, M.J. Paul, and B.J. Peterson. 2001. Inter-biome comparison of factors cont rolling stream metabolism. Freshwater Biology 46: 1503-1517.

Mulligan, H. and A. Baranowski. 1969. Growth of phytoplankton and vascular aquatic plants at different nutrient levels. Proceedings- Inte rnational Association of Theoretical and Applied Limnology 17d: 802-810.

Munch, D.A., D.J. Toth, C. Huang, J.B. Davis, C.M. Fortich, W.L. Osburn, E.J. Phlips, E.L. Quinlan, M.S. Allen, M.J. Woods, P. Cooney, R. L. Knight, R.A. Clarke, S.L. Knight. 2006. Fifty-year retrospective study of th e ecology of Silver Springs, Florida. Publication Number: SJ2007-SP4. St. Johns Ri ver Water Management District, Palatka, FL.

Nixon, S.W. 1995. Coastal marine eutrophication: A definition, social causes, and future concerns. Ophelia 41: 199-219.

Notestein, S.K. 2001. Physical, Chemical , and Vegetative Characteristics of the Chassahowitzka River. M.S. Thesis. Univer sity of Florida. Gainesville, FL. 86 pp.

Notestein, S.K., T.K. Frazer, M.V. Hoyer, and D. E. Canfield, Jr. 2003. Nutrient Limitation of Periphyton in a Spring-Fed, Coastal Stream in Florida, USA. Journal of Aquatic Plant Management: 41: 57-60.

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Odum, H.T. 1956. Primary Produ ction in Flowing Waters. Limnology and Oceanography 1: 102-117.303 Odum, H.T. 1957a. Primary Produ ction Measurements in Eleven Florida Springs and a Marine Turtle-Grass Community. Limnol ogy and Oceanography 2: 85-97.

Odum, H.T. 1957b. Trophic structur e and productivity of Silver Sp rings, Florida. Ecological Monographs 27(1): 55-112.

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# 7

Managing Nutrient Inputs to Flor ida Springs: The Legal Framework

Richard Hamann Center for Governmental Responsibility Levin College of Law University of Florida

Summary The legal framework for managing the nutrient po llution of Florida springs is potentially as broad as the scope of human activities that cont ribute nutrients to the sp ringshed. Regulations on the discharge of wastewater ar e key elements. The federal Clean Water Act and Safe Drinking Water Act provide a set of manda tes and incentives for state programs. The focus of the Clean Water Act is protecting surface waters through th e regulation of point source discharges under the National Pollutant Discharge Elimination Syst em (NPDES). Discharges to groundwater may be regulated under this program if there is a “significant nexus” to th e quality of navigable surface waters. The Safe Drinking Water Act protects public water supplies by establishing minimum criteria for drinking water quality a nd requiring states to regulate the underground injection of pollutants. Groundwater quality st andards for nutrients are thus focused on the protection of public health. Florida has been delegated NPDES permitting authority by the Environmental Protection Agency (EPA) and im plements an Underground Injection Control (UIC) program that is consistent with the SDWA . In addition, Florida regulates certain sources of nutrient pollution that do not fall within the dire ct jurisdiction of the federal program, such as nonpoint sources, agricultural discharges, a nd additional discharges to groundwater.

The regulatory program requires discharges to achieve effluent limitations based on the application of specified levels of technology or to achieve water quality standards, whichever is more stringent. Domestic wastewater treatmen t plants are generally only required to use secondary treatment, except in ar eas where the Legislature has required advanced wastewater treatment (AWT). Concentrated Animal Feeding Operations (CAFO) are currently regulated under state rules pending the adoption by EPA of a new federal rule. Those stormwater dischargers subject to NPDES permitting are required to reduce the discharge of pollutants to the Maximum Extent Practicable.

Florida’s surface water quality standards must be reviewed and approved by EPA every three years. They consist of designated uses, narrative and numeric criteria for each of those uses, and moderating provisions. There is al so a general limitation on the de gradation of any waters and a prohibition on the degradation of designated Ou tstanding Florida Waters (OFW). A narrative nutrient standard prohibits altering nutrient concentrations “so as to cause an imbalance in natural populations of aquatic fl ora or fauna.” Numeric standard s can also be developed, most commonly as Total Maximum Daily Loads (TMDL).306 TMDLs can be adopted by either EPA or the Fl orida Department of Environmental Protection (DEP). Florida’s process begins with determini ng whether a particular waterbody is “impaired” based on “objective and credible data, studies and reports” demons trating it does not meet water quality standards for a specific cr iterion, such as nitrogen or phosphorus. If there is sufficient data demonstrating the concentration of the spec ific pollutant causing the impairment, it can be added to the verified list. A TMDL is then calculated and the load reasonably and equitably allocated to the various sources and basins co ntributing pollutants. The DEP then assembles a group of stakeholders to deve lop a Basin Management Acti on Plan (BMAP) to develop strategies for implementing the adopted TMDL. NP DES permits may be revi sed as necessary to implement the BMAP. Other dischargers may be required to reduce their discharge “to the maximum extent practicable”. Nonpoint sour ces must demonstrate compliance by either implementing Best Management Practices (BMP ) or conducting water quality monitoring. There are additional powerful incentives for implemen ting BMPs. BMPs for nonagricultural sources may be adopted as rules by the DEP or a water management district. Only the Florida Department of Agriculture and Consumer Services (DACS) has the authority to adopt them for agricultural sources, although they must be verified by the DE P or a water management district.

Stormwater is a significant source of pollutants th at can be regulated as a point source, because much of it is collected into pi pes or channels, or as a nonpoint source, because much of it comes from diffuse sources and activities. The NP DES program regulates many of the larger stormwater systems and stormwater associated with industrial activities. Systems subject to NPDES permitting must be periodically reviewed and ways to reduce pollution must be considered in permit renewals. Most stormwat er systems in Florida, however, are either unpermitted..regulated or regulated under an Environmental Resource Permit program or its predecessor. The emphasis of ERP permitting is on the design and construction of stormwater systems. Although they must be operated and maintained, there is no program for periodic review for compliance or enhancements. Because even systems that are constructed in compliance with the current standards discharge more pollutants than previously assumed, the DEP is now developing a unified stormwater rule. Many local governments also have their own stormwater programs.

Septic tanks and other Onsite Treatment and Disposal Systems (OSTDS) are another major potential source of nutrient pollu tion. The Department of Health (DOH) has adopted regulations for the construction and siting of septic tanks and drainfields. However, these systems are inefficient at removing nitrogen, a nutrient of concern in springsheds. In some areas where nutrients are a concern, such as the Florida Keys, performance-based systems have been required. Local governments can adopt more stringent requirements.

Local governments are required to periodically revise and impleme nt comprehensive plans that can provide a framework for local springs pr otection. Local comprehensive plans and land development regulations can limit the intensit y and design of land development to limit the contribution of nutrients to springs. They can provide for improved stormwater and wastewater management. The Florida Legislature has specifi cally required local governments in the Wekiva Study Area to amend their comprehensive plans to enhance springs protecti on. It has also limited the ability of local governments to regulate agricultural land uses and the conversion of307 agricultural land to urban uses. Local governments are li able for compensating landowners who show an “inordinate burden”. A Consumer Fertilizer Task Force has recommended new restrictions on the ability of local government s to regulate the application of fertilizers.

REGULATION OF DISCHARGES

Clean Water Act

The Federal Water Pollution C ontrol Act Amendments of 1972 3 created one of our nation’s most comprehensive environmental statutes. It was en acted in response to wi despread water pollution, including nutrient overenrichment, and the failure of the states and weak federal institutions to remedy it. The goal of the FWPCA was to “protect and restore the chemical, physical and biological integrity of the nation’s waters 4 ” through a combination of research, planning, construction and regulation. The Act was amended in 1977 and renamed the Clean Water Act 5 (CWA), but the fundamental structure and policie s have remained intact for over 35 years.

The CWA prohibits any discharge of pollutants that is not authori zed by a permit or otherwise in compliance with the Act 6 . Discharges may be authorized through a permit issued by EPA or a state under a program called the National Polluta nt Discharge Elimination System (NPDES) 7

. The issuance of a permit depends on compliance with the most stringent effluent limitations based on either technology or water quality. Te chnology-based effluent limitations require the use of specified levels of techno logy for the prevention or removal of pollutants regardless of the impacts to water quality. Water quality-based effluent limitations impose additional limits on discharges to prevent violations of state water quality standards.

Not every source of nutrients to springs, howev er, qualifies as a regulated discharge under the definitions and exemptions of the CWA. A discharge is defined as “any addition of any pollutant to navigable waters fr om any point source” 8 . Nutrients certainly fall within the definition of a pollutant and many common sources of nutrients fit the definition of a “point source ” as a “discernible, confined and disc rete conveyance” of pollutants 9 . Point sources might include wastewater treatment plants, i ndustrial sources, concentrated animal feeding operations, or stormwater systems. Agricultural stormwater and irrigation return flows are expressly exempted from regulation, even though they might have th e physical characteristics of a point source 10 . A spring itself might be consider ed a point source under this defi nition if it served to directly convey pollutants to navigable waters.

3 PL 92-500, codified at 33 U.S.C. §§1251-1387 (2007).
4 33 U.S.C. §1251(a) (2007).
5 PL 95-217, Section 1.
6 33 U.S.C. §1311(a) (2007).
7 Id., §1342.
8 Id., §1362(12).
9 .Id., §1362(14).
10 Id.308 Although the term “navigable waters” was define d in the Act as “waters of the United States 11 ”, the extent to which discharges to groundwat er are subject to the CWA is not clear 12 . Several judicial opinions held that Congress intended to exclude groundwater from the scope of regulation under this statute 13 . Other decisions have found juri sdiction over groundwater that is hydrologically connected to surface waters 14

.
The Supreme Court’s recent decision in Rapanos v. United States 15 has further confused the scope of CWA jurisdiction in a case involving th e discharge of dredged or fill material to wetlands. A plurality held that there was no jurisdiction over waters that were not relatively permanent streams or standing bodies of water and connected wetlands. One concurring justice, with the deciding vote, held that jurisdicti on required a “substantial ne xus” to waters that are navigable 16 . In interpreting these opinions, the 11 th Circuit Court of Appeals recently held there might be no CWA jurisdiction over a discharge of pipe processing wastew ater to a perennial stream that connected to navigable waters 17 . A “mere hydrologic connection” was insufficient. A “significant nexus” must be proved in a new trial, presumably by showing how the discharge could significantly affect the chem ical, physical or biological inte grity of the navigable waters.

A similar test would have to be met for any discha rge of nutrients to groundwater to be subject to CWA regulation. There would have to be a rela tively direct hydrologic c onnection to navigable waters and there would have to be proof that the discharge signi ficantly affected the quality of

11 33 U.S.C. §1362(7)
12 See generally, Thomas L. Casey, III, Reevalua ting “Isolated Waters”: Is Hydrologically Connected Groundwater “Navigable Water” Under the Clean Water Act?, 54 Ala. L. Rev. 159 (2002).
13 Rice v. Harken Exploration Co., 250 F.3d 264 (5 th Cir 2001); Village of Oconomowoc Lake v. Dayton Hudson Corp., 24 F.3d 962, 965 (7th Cir. 1994); Town of Norfolk v. United States Corps of Engineers, 968 F.2d 1438, 1451 (1st Cir. 1992); Exxon Corp. v. Train, 554 F.2d 1310, 1322 (5th Cir. 1977); Umatilla Water Quality Protective Ass'n, Inc. v. Smith Frozen Foods, Inc., 962 F.Supp. 1312, 1318 (D. Ore. 1997); Kelley v. United States, 618 F. Supp. 1103 (W.D. Mich. 1985).
14 U.S. Steel Corp. v. Train, 556 F.2d 822, 852 (7th Cir. 1977); Coldani v. Hamm, 2007 U.S. Dist. LEXIS 62644 (E.D.Cal. 2007); Idaho Rural Council v. Bosma, 2001 U.S. Dist. LEXIS 9519 (D. Idaho 2001); Washington Wilderness Coalition v. Hecla Min. Co., 870 F.Supp. 983, 990-991 (E.D. Wash. 1994); Sierra Club v. Colorado Refining Co., 838 F.Supp. 1428, 1434 (D. Colo. 1993); Williams v. PipeLine Co. v. Bayer Corp., 964 F.Supp. 1300, 1320 (S.D. Iowa ); Idaho Rural Council v. Bosma, 143 F.Supp.2d 1169, 1180 (D. Idaho 2001); McClellan Ecological Seepage Situ ation (MESS) v. Weinberger, 707 F. Supp. 1182, 1196 (E.D.Cal.1988); Friends of Santa Fe County v. LAC Minerals, Inc., 892 F. Supp. 1333, 1357 (D. New Mexico 1995).
15 126 S.Ct. 2208 (2006).
16 The requisite nexus could be shown if the subj ect wetlands “either alone or in combination with similarly situated lands in the region, significantly aff ect the chemical, physical, and biological integrity of other covered waters more read ily understood as ‘navigable’”. 126 S Ct. 2248. Those effects cannot be “s peculative or insubstantial” Id.
17 U.S. v. Robison, 505 F.3d 1208 (11 th Cir 2007).309 those surface waters. The 9 th Circuit Court of Appeals recently found such a significant nexus in a case involving the discha rge of wastewater to a pond adjacent to a river 18 . Because the wastewater seeped into the river through a grou ndwater connection, affected the quality of the surface water and had a significant ecological connection through bir d, mammal and fish populations, it was a regulated water 19

.

State Discharge Permitting

The DEP has assumed primary responsibility from EPA for implementation of the NPDES permitting program 20 . For discharges to groundwater th at do not have a sufficiently close connection to a surface water, however, only stat e law applies. Florida’s regulatory authority over discharges that affect groundwater is more extensive than that of the federal Clean Water Act. The Florida Air and Water Pollution Contro l Act prohibits the construction or operation without a permit of any stationa ry installation reasonably expect ed to be a source of water pollution 21 . An “installation” is any “s tructure, equipment or facility which may emit . . .water contaminants. . . 22 ” “Pollution” is “the presence in the . . .waters of the state of any substances, contaminants, . . .or manmade or human-induced im pairment of . . .waters, or alteration of the chemical, physical, biological, or ra diological integrity of . . .wat er in quantities or at levels which are or may be potentially harm ful or injurious to human health or welfare, animal or plant life, or property or which unreas onably interfere with the enjoymen t of life or property, including outdoor recreation unless aut horized by applicable law. 23 ” Finally, “waters” “i nclude, but are not limited to, rivers, lakes, streams, springs, im poundments, wetlands, and all other waters . . .including underground waters 24 .” The distinction in federal law between “point” and “nonpoint” sources and between “navigable waters” and groundwater is thus not very significant in Florida 25 . The state, unlike EPA, has clear authority to regulate virtually any discharge to any
water.

Several sections of Florida st atutes authorize the DEP to a dopt rules and issue permits for

18 N.Cal. Riverwatch v. City of Healdsburg, 496 F.3d 993 (9 th Cir. 2007).
19 496 F.3d 1000-1001.
20 Fla. Stat. §303.0885 (2007); Rule 62-620, FAC. DE P assumed NPDES authority for industrial and domestic wastewater discharges in 1995 and stormwater discharges in 2000. See generally, Terry Cole, National Pollutant Di scharge Elimination System Program, Ch. 12.2, Florida Environmental and Land Use Law (June 2007).
21 Fla. Stat. §403.087(1) (2007).
22 Id., §403.031(1).
23 Id., §403.031(7)
24 Id., §403.031(13)
25 If a discharge is regula ted under Florida law, but not subjec t to the Clean Water Act, however, then certain provisions adopted to allow th e state to qualify fo r NPDES permitting do not apply. Id., §403.0885(2). For example, non-NPDES discharge permits can be issued for up to ten years, whereas NPDES permits are limited to five years. Id

. , §403.087(3). Unlike most other state permits, NPDES permits are not subject to issuance by default if the state fails to take action within ninety days. Id

. , 403.0885(3).310 sources of pollution 26 . If wastewater is discharged direct ly to groundwater vi a a well, then an Underground Injection Contro l (UIC) permit is required 27 . Generally, dischargers are required to provide “reasonable assurance” that water quality standards will not be violated and that appropriate levels of treatment are being implem ented. The rules applicable to certain significant sources are set forth below.

Domestic Wastewater

Domestic wastewater must generally receive a minimum of secondary treatment before it is discharged. Additional treatment may be required to meet water quality standard s, depending on where the effluent is discharged. Discharges to groundwater can be made through a variety of systems, with different levels of public access 28 . Treated wastewater is increasingly in demand for irrigation of areas where the public has access, e.g. golf courses and lawns. For dischargers to groundwater through sprayfields or rapid infiltration basins, the practical limit on nitrogen has generally been set by the Maxi mum Contaminant Level (MCL) for drinking water of 10 mg/l measured at the edge of the site or zone of discharge.

Advanced wastewater treatment has been required in areas where the Legislature has determined that the risks of regul atory uncertainty and delay should be reduced in favor of imposing the most protective requirements. For example, wastew ater discharged to such estuaries as Sarasota Bay, Tampa Bay, Charlotte Harbor and their tri butaries must meet advanced waste treatment standards 29 . Special requirements have recently been imposed on domestic wastewater discharges in the Wekiva Study Area that may serve as a model for ot her areas contributing recharge to springs 30 . These requirements were based on recommendations developed pursuant to the Wekiva Parkway and Protection Act 31 . The DEP first determined that 0.2 mg/l should be adopted as an initial target fo r nitrogen concentrations at th e spring vents, based on unimpacted reference sites. The DEP also mapped the recharge areas according to their relative vulnerability.

26 Id., §§403.061(14), .087(2), (5), .088(2) (2007). Th e rules governing permitting are at Fla. Admin. Code r. 62-4 (2007).
27 Fla. Admin. Code r. 62-528 (2007).
28 Rule 62-610 provides different criteria depe nding on whether treated water feeds a reuse system or is applied to the land in slow or high rate systems with restricted or open

access. 29 Fla. Stat. §403.086(1)© (2007). For nutrients, the AWT standard is Total Nitrogen, 3 mg/l and Phosphorus, 1 mg/l. Id., §403.086(4).
30 Fla. Admin. Code r. 62-600.550 was adopted April 13, 2006.
31 Chapter 2004-384, Laws of Florida codified at Fla. Stat. §369.318(1) (2007). The Wekiva Study Area was delineated by the Florida Le gislature, Id. §369.316, and includes large parts of the springsheds of Rock Spring, We kiwa Spring and other springs feeding the Wekiva River. For the basis of recommenda tions specific to wastewater discharge see Florida Department of Environmental Pr otection, A Strategy for Water Quality Protection: Wastewater Treatment in the Wekiva Study Area (December 2004). For background information on the various planning and coordination activi ties related to the Wekiva Study Area see http://www.dca.state. fl.us/fdcp/dcp/wekiva/wekivaact/index.cfm (visited December 14, 2007).311 The rule thus adopted three protection zones, wi th the Primary Protection Zone considered “Most Vulnerable”. The rule prohibits new or expanded sprayfields or rapid infiltration basins and absorption fields within the Primary Protection Zone 32 . Large existing facilities are limited to 3.0 mg/l Total Nitrogen. Increasing concentrations of nitrogen ar e allowed in areas of less vulnerability, for systems that have lower rates of applicati on and for smaller systems whose costs of additional treatment are much higher. Dischargers can qualify for relief from these requirements by making “an affirmative demonstr ation, based on relevant water quality data, physical circumstances, or other credible informa tion, that the discharge of reclaimed water is protective of surface and ground water quality with respect to the target nitrate-nitrogen level of 0.2 mg/l, as N, for the spring vent. . . 33 ”

Industrial Discharges and Concentrated Animal Feeding Operations (CAFO)

Industrial wastewater is any regulated wastewater that is not domestic wastewater 34 . Industrial dischargers must comply with technology-based effluent lim itations developed for various industrial categories and adopted as rules, as well water qual ity-based effluent limitations 35 . New sources must meet new source performance sta ndards (NSPS). Existing sources are required to implement the best available technology econo mically achievable (BAT). For conventional pollutants such as Biological Oxygen Demanding materials (BOD5) or Total Suspended Solids (TSS), a less stringent effluent limitation is a pplicable, termed best conventional pollutant control (BCT).

Concentrated Animal Feeding Operations (CAFO) are a particular type of industrial discharger that generates substantia l quantities of nutrients 36 . The national regulations adopted by EPA have been challenged by both industrial and environmental interests 37 . Final revisions have not yet been adopted 38 . Florida will be required to implemen t the final federal rules, but meanwhile such facilities are governed by rule 40 C.F.R. pt. 412. Facilities that ar e not subject to NPDES permitting are governed by rule 62-670, F.A.C. 39

32 Fla. Admin. Code r. 62-600.550(4)(a)(2007). Tec hnically these are term ed “rapid-rate or restricted access slow-rate land application systems”.
33 Id., 62-600.550(7). This provision a ppears to allow the discharger to demonstrate that the discharge will achieve compliance with the targ et goal, not to challe nge the goal itself.
34 Id., 62-660.200(2)(e).
35 Id., 62-660.400. DEP has adopted by reference th e rules developed and adopted by EPA under the Clean Water Act.
36 See generally, James H. Andreasen, Concentrat ed Animal Feeding Operations: A Program in Transition, Natural Resources and Envi ronment 45 (Spring 2007). CAFOs produce 500 million tons of manure annually. Id. at 45.
37 Waterkeeper Alliance v. U.S. Envt’l Pr otection Agency, 399 F.3d 486 (2d Cir. 2005).
38 71 Fed. Reg. 37744 (June 30, 2006). The existing rules are at 40 C.F.R. pt 412.
39 The application of the rule, however, has been affected by court order in Save Our Suwannee v. Fla. Dept. of Envt’l Protection, ER FALR 04:181, Case No. 2001-CA-001266 (Fla. 2d Cir. Ct Mar. 5, 2004). See Susan L. Stephens, Florida’s Waters: Signi ficant Cases in the Works, Section Reporter (March 2005). E nvironmental and Land Use Law Section, The Florida Bar. Available at http://www.el uls.org, visited February 23, 2008. DEP is312 WATER QUALITY STANDARDS

Florida has established classifications a nd standards for both ground and surface water 40 . One of the primary regulatory criteria for any discharge is that it not cause or contribute to violations of water quality standards 41

.

Groundwater Quality Standards

The DEP generally regulates di scharges of pollutants to gr oundwater and has established a scheme of groundwater clas sification and standards. 42 The standards for groundwater are primarily designed to protect human health 43 . For example, all groundwater must be “free from” substances in concentrations which are carcinogen ic, mutagenic, teratogenic, or toxic to human beings 44 .. The only protection for non-human organisms in the groundwater is for plants, animals and organisms native to the soil that are respon sible for treating the discharge. A groundwater discharge is also prohib ited from causing violations of the acu te toxicity standard for surface water if it reaches the surface 45 , cannot cause a nuisance, and cannot interfere with the reasonable beneficial use of adjacent waters. A nutrient discharge that causes the growth of algae resulting in allergic react ions by recreational users coul d violate these criteria. The difficulties of proving a causal relationship, however, are substantial.

Other standards applicable to groundwater depend on the classification. Generally, the groundwater that discharges to springs is likely to be Class G-II and for this classification the most important standard is the maximum c ontaminant levels (MCL) for drinking water 46 . The only nutrient listed in the MCL is nitrate/nitrite which is limited to 10 mg/l 47 . Groundwater quality standards thus protect spring discharges agai nst only the highest levels of nutrient contamination..

engaged in rulemaking to revise rule 62-670 and adopt new rules for certain types of animal feeding operations. See draf t rules 62-620, 62- 621.500(3) & (3)(a).
40 Florida, like other states, must review its surface water quality standards every three years and submit any changes to EPA for review and approval. 33 U.S.C. §1313; 40 CFR Part 131.
41 Fla. Admin. Code r.62- 4.030, -4.070, -4.240(4) (2007).
42 See generally, Cynthia Christ en, Groundwater Protection: An Overview, Ch 11.1, Florida Environmental and Land Use Law (Feb 2002).
43 Fla. Admin. code ch 62-520 (2007).
44 Id., 62–520.400(1)(b).
45 Rule 62–302.200(1) is the acute toxicity standard.
46 Rules 62–520.420; 62–550.310 (primary);62–550.320 (secondary). The federal Safe Drinking Water Act (SDWA), 42 U.S.C. §§300f-300j-26 (2007), requires states to regulate the quality of water delivered for human cons umption through water supply systems. EPA adopts Maximum Contaminant Levels (MCL) as minimum criteria for drinking water. The SDWA also requires the st ates to regulate the discharg e of pollutants to underground sources of water and MCLs ha ve become commonly used to limit the contamination of groundwater and provide criteria for the remediation of contaminated sites.
47 Fla. Admin. Code r. 62-550.310 and Table 1. The level can be up to 20 mg/l under some circumstances.313 Surface Water Quality Standards

Once groundwater reaches the surface at a spri ng, however, surface water quality standards apply and a discharge of nutrien ts to groundwater cannot be perm itted if it would cause or contribute to a violation of surface water quality standards 48 . Florida’s surface water quality standards include a polic y statement on nutrients 49 . It states,

The Department finds that excessive nutrients (total nitrogen and total phosphorus) constitute one of the most severe water quality problems facing the State. It shall be the Department's policy to limit the introduction of man-induced nutrients into waters of the State. Particular consideration shall be given to the protection from further nutrient enrichment of waters which are presently high in nutrient concentrations or sensitive to further nutrient concentrations and sensitive to further nutrient loadings. Also, particular consideration shall be given to the protection from nutrient enrichment of those presently containing very low nutrient concentrations: less than 0.3 milligrams per liter total nitrogen or less than 0.04 milligrams per liter total phosphorus.

Water quality standards have two components, a classification of the designated use and standards applicable to that class of water. In addition, there are both a general nondegradation standard and special criteria applicable to Outstanding Florida Waters (OFW). The “moderating provisions” are also considered part of state water quality standards 50

.
Most spring discharges in Florid a will be Class III waters, whose designated use is “Recreation, Propagation and Maintenance of a Healthy, Well -Balanced Population of Fish and Wildlife”. A discharge cannot interfere with the designated use 51 . A discharge of nutrients that interfered with recreational swimming, for example, or the propagation of wildlife, would violate this criterion. In addition, a discharge cannot inte rfere with an existing use, which “may be different or more extensive than the designated use” 52

.
There are both minimum water quality standards, applicable to all classes, and numeric and narrative standards specific to each class. Th e minimum criteria prohibit discharge components which produce nuisance or toxic conditions 53 . To the extent that nutri ent discharges could have such effects, they would violate the minimum standards. In that ca se they would also be likely to

48 Id., 62-4.070; 62-302.300(15).
49 Id., 62-302.300(13)
50 The moderating provisions include exemptions from water quality standards, Rule 62-4.243, mixing zones, rule 62-4.244 and site sp ecific alternative cr iteria, rule 62-302.800. Dischargers who are violating water quality st andards can also continue to operate under the terms of an administrative order, apply fo r a variance or seek reclassification of the affected waters. See generally, Jennifer L. Fitzwater, Relief from Florida Water Quality Criteria, 12.4, Florida Environmen tal and Land Use Law (August 2003).
51 Fla. Stat. §403.021 (2007); Fla. Admi n. Code r. 62-302.300(11), (14) (2007).
52 Id., 62-302.300(14). A discharger can, however, show th at a particular use is not attainable and overcome this prohibition.
53 Fla. Admin. Code r. 62-302.500(1).314 violate the specific narrative standards for nutrients.

There is a general prohibition on any substances, including nutrients, “in concentrations which result in the dominance of nuisance species” 54 . The narrative standard specific to nutrients states, “In no case shall nutrient concentrations of a body of water be altered so as to cause an imbalance in natural populations of aquatic flora or fauna. 55 ” Determining the point at which an alteration of nutrient concentrations causes an imbalance can be extraordinarily difficult and conflicted. For example, in 1988 the United States sued the State of Florida and the South Florida Water Management Distri ct, alleging that operation of th e water management system in South Florida caused an im balance of flora and fauna 56 . In 1992 the lawsuit was settled 57 , but it was not until 2004 that a numeric criter ion for phosphorus was finally adopted 58

.
Florida’s nondegradation policy and implementing rules also control nutrient discharges. In addition to protecting existing and designated uses , the policy prohibits an y degradation of water quality by a new discharge or the expansion of an existing discharge unless the degradation is “necessary or desirable under federal standards and under circumstances which are clearly in the public interest” 59 Degradation, for these purposes, incl udes “Man-induced nutrient enrichment (total nitrogen or total phosphorus). . . 60 ” Factors for determining wh ether a discharge is clearly in the public interest are set forth at rule 62-4.242 (1). Nutrient enrichment can be allowed by the DEP after considering and ba lancing these factors.

A related nondegradation standard protects wate rs that are designated as Outstanding Florida Waters (OFW) and thus entitled to the highest levels of protection 61 . Many of Florida’s springs are designated as OFWs in Rule 62-302.700, F.A.C .. The standard applies to discharges or activities that are located within the boundaries of an OFW or that contribute to significant degradation of the OFW. Existing ambient wate r quality cannot be lowere d as a result of the proposed activity or discharge 62 . The term “existing ambient wate r quality” is defined as the water quality that could reasonabl y have been expected, based on th e best scientific information

54 Id., 62-302.530(46).
55 Id., 62-302.530(47)(b).
56 United States v. S. Fla. Water Manageme nt Dist., Case No. 88-1886 (S.D. Fla.).
57 United States v. S. Fla. Water Mana gement Dist., 847 F.Supp. 1567 (S.D.Fla.1992). Enforcement of the resulting consent decree has been similarly controversial. U.S. v. S.F.W.M.D., 373 F.Supp.2d 1338 (S.D. Fla. 2005).
58 Fla. Admin. Code r. 62-302.540 (2007). For a compre hensive review of the issues and process see Miccosukee Tribe of Indians v. Dept of Envt’l Protection, Final Order, DOAH Case No. 03-2872RP (June 17, 2004). Approval of the standards by EPA was also challenged. Miccosukee Tribe of Indians v. U. S., 2006 WL 648055 (S.D. Fla. 2006).
59 Fla. Admin. Code r. 62-302.300(17)(2007).
60 Id., 62-302.530 (47) (a) (Table).
61 Id., 62-302.700, 62-4.242(2). A slightly more stringe nt rule applies to Outstanding National Resource Waters, which are given additional pr otection against degradation to the extent required by federal rule 40 C.F.R. 131.12. Fla. Admin. code r. 62-4.242(2)(3)(2007).
62 Id.,62-4.242(2)(a)2. The activity must also be “clearly in the public interest”. Id. Note that the rule contains numerous exceptions and qua lifications to these requirements.315 available, to have existed during the baseline year for the particular OFW (i.e. the year prior to designation) or the year pr ior to a permit application 63

.
Specific numeric nutrient standards can be establis hed in several ways. Th e Legislature required the DEP to establish a numerical crit erion for phosphorus in the Everglades 64 . Rule 62-40 provides for the development by water management districts or local governments of pollutant load reduction goals as part of watershed management plans or basin-specific criteria development 65 Most numeric criteria, however, are be ing developed as Total Maximum Daily Loads (TMDL), discussed below. The DEP also has an ongoing project to develop numeric criteria for nutrients that can be applied throughout the state 66

.

TOTAL MAXIMUM DAILY LOADS

The Clean Water Act required the states to asse ss their waters to identify those areas where compliance with technology-based effluent limitati ons is not sufficient to achieve water quality standards 67 . For those waters, the state is required to establish a total maximum daily load (TMDL) for relevant pollutants at a level sufficien t to meet the standards or EPA is required to establish one. EPA and the states effectively ignored this requirement until citizen groups had prevailed in a number of lawsuits to fo rce compliance with the statutory mandate 68 . The result in Florida was a 1999 consent decree requiring EPA to establish TMDLs for almost 2000 waterbodies by 2011 69 . The Florida Legislature responded by enacting the Florida Watershed Protection Act 70 , which established a process for developing and implementing TMDLs by the state. Florida now has a very comprehensive process for assess ing waterbodies, establishing TMDLs and developing Basin Management Acti on Plans (BMAP) to a ddress water quality violations 71 . TMDLs and BMAPs are being developed to address nutrient pollution in many areas. Whenever the state fails to develop a TMDL that meets fede ral criteria or the requirements of the consent decree, then EPA must establish one 72 . The Florida Legislature has directed the DEP to take action to develop and implement TM DLs for several specific waterbodies, including

63 Id.,62-4.242(2)©
64 Fla. Stat. §§373.4592(2)(m), (4)(e)(2007). The statut e provided for a default standard of 10 ppb to be used if the scientific uncertainties could not be resolved sufficiently to adopt a rule by 2004. Id., §373.4592(4)(e)2.
65 Fla. Admin. Code r. 62-40.210 (20), .430 (2007). The Districts are aut horized to develop strategies for meeting water quality st andards in Surface Water Improvement and Management (SWIM) plans. Fla. Stat. §373.453(2)(e) (2007).
66 http://www.dep.state.fl.us/water /wqssp/nutrien ts/index.htm.
67 33 U.S.C. §1313(d) (2007).
68 OLIVER A. HOUCK, THE CLEAN WATER ACT TMDL PROGRAM: LAW, POLICY AND IMPLEMENTATION (2d Ed. 2002). Environmental Law Institute, Washington.D.C. ISBN 1-58576-038-2.
69 Florida Wildlife Federation v. Carol Browne r, Case No. 4:98CV356-WS (N.D.Fla.).
70 Fla. Stat. § 403.067 (2007).
71 DEP TMDL documents are available at http ://www.dep.state.fl.us/water/tmdl/index.htm.
72 EPA TMDL documents are accessible at http://www.epa.gov/OWOW/TMDL/.316 Lake Okeechobee and its tributaries 73 , the Caloosahatchee 74 , the St Lucie 75 and the Wekiva 76

. Legislation has been proposed to require th e development of TMDLs for additional springs 77

.

Impaired Waters

For the purpose of developing TMDLs, the DEP has assigned all of the state’s watersheds to one of five groups, which it reviews every five years. The cycle begi ns by determining whether there is sufficient quality data to prove that a partic ular waterbody is “impaired”, as defined in the impaired waters rule 78 , to justify placement on the planni ng list for further evaluation and assessment. A waterbody is impaired if there is “objective and credible data, studies and reports” demonstrating that it does not meet water quality standards for a specific criterion 79 . The rule establishes stringent quality standards for the data that can be used in this assessment. If numeric criteria have been estab lished, then narrative or biological criteria cannot be used as the basis for showing impairment unless the DEP demonstrates the numer ic criteria are not adequate 80 . If impairment is based on narrative or biological criteria, “the specific factors concerning particular pollutants shall be identified prior to a total maximum daily load being developed for those criteria. . . 81 ” For interpreting the narrative nutrient criteria, the rule sets forth specific guidance:

Trophic state indices (TSIs) and annual m ean chlorophyll a values shall be the primary means for assessing whether a water should be assessed further for nutrient impairment. Other information indi cating an imbalance in flora or fauna due to nutrient enrichment , including, but not limited to, algal blooms, excessive macrophyte growth, decrease in the dist ribution (either in density or areal coverage) of seagrasses or other submerged aquatic ve getation, changes in algal species richness, and excessive diel oxygen swings, shall also be considered 82

.
Specific nutrient criteria are also specified for streams 83 , lakes 84 and estuaries 85

.

73 Fla. Stat. §§ 373.4595(1)(e)-(h), (l), ©, (4) (2007).
74 Id., § 373.4595(5).
75 Id.,§ 373.4595(l), ©, (4).
76 Id., § 369.318(1), (8).
77 Proposed Florida Springs Protection Act, SB 2394 (02/28/2008). Earlier proposed legislation also relied on a modified TMDL proce ss. See e.g., SB 2368 (03/08/2005), 2005 Regular Session, The Florida Senate. Use of the TMDL program to protect Florida springs was proposed by students in the Conservation Clin ic at the University of Florida Levin College of Law. See http://www.law.ufl. edu/conservation/sp rings_narrative.shtml (visited April 22, 2008).
78 Fla. Admin. Code r. 62-303 (2007).
79 Fla. Stat. § 403.067(3)(b) (2007).
80 Id., ©.
81 Id

. 82 Fla. Admin. Code r. 62-303.350(1) (2007).
83 Id

. , 62-303.351.
84 Id

. , 62-303.352.317 The Verified List

The next step is to determine whether to list the waterbody on the “verified list”. For a water to be placed on this list due to nu trient enrichment, the DEP must ha ve five years of data sufficient to develop a site-specific threshold for when an im balance in flora or fauna occurs or to apply the generic thresholds for streams, lake and estuaries 86 . To add a water to the verified list, the DEP must be able to “specify the pollutant or pollutants causing the impairment and the concentration of the pollutant(s) cau sing the impairment. 87 ” For waters impaired by nutrients, the DEP must identify whether it is limite d by nitrogen, phosphorus or both 88 . Even if the water meets these criteria, it cannot be added to the verified list if the DEP determines that “existing or proposed technology-based effluent limitations and other pollution control programs under local, state, or federal authority are sufficient to result in the a ttainment of applicable water quality standards” in the future, with “reasonabl e progress” by the next cycle 89 . The verified list must be submitted to EPA for review and approval an d is subject to challenge in state administrative courts as a final order 90

.

Calculation and Allocation

The development of a TMDL is the next step in the process for those waterbodies that reach the verified list. For each pollutant specified in the verified list, the DEP must calculate the amount that can be received from all sources without exceeding water quality standards 91 . In calculating the TMDL, the DEP must account fo r seasonal variations and “incl ude a margin of safety that takes into account any lack of knowledge concerning the relationship between effluent limitations and water quality 92 .” The calculated load must then be allocated to the various sources contributing pollutants. The DEP may allocate loadings betw een point and nonpoint sources, among basins and among sources. The allo cation must be “reas onable and equitable” and based on considerati on of ten listed factors 93 . TMDLs are adopted as rules by the Secretary. Detailed allocations may be left to the Basin Management Action Plan.

85 Id

. , 62-303.353. Open coastal waters are also included.
86 Id

. , 62-303.450.
87 Id

. , 62-303.710(1).
88 Id

. , 62-303.710(4).
89 Id

. , 62-303.600. An example of such a case is the Suwannee River Partnership. Ay ş in Dedekorkut, Suwannee River Part nership: Representation Inst ead of Regulation, Ch 2 in JOHN T. SCHOLTZ AND BRUCE STIFTEL, EDS., ADAPTIVE GOVERNANCE AND WATER CONFLICT: NEW INSTIT UTIONS FOR COLLABORATIVE PLANNING (2005). Resources for the Fu ture, Washington, D.C. ISBN 1-933115-18-1.
90 Winston K. Borkowski, Total Maximum Daily Loads in Florida– The New Millennium, 12.5-5 in Florida Environmental and Land Use Law (June 2007).
91 Fla. Stat. § 403.067(6)(a)2 (2007).
92 Id.
93 Id., §403.067(6)(b).318 Basin Management Action Plans (BMAP)

Basin Management Action Plans (BMAP) are th e foundation for implementation of an adopted TMDL. BMAPs are intended to integrate existing water management strategies to achieve water quality standards and equitably allocate po llutant reductions. They must include an implementation schedule, a basis for evaluating the plan’s effectiven ess, feasible funding strategies and mechanisms to address future increases in pollutant loading. Phased implementation is allowed. Credits may be grante d to sources that reduced pollutant loads prior to development of the BMAP. The plan must in clude provisions for monitoring water quality and revising the plan in response to the results 94

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The BMAP may include such nonregulatory m easures as public works construction, land acquisition, education, waste minimization, pollu tion prevention, interagency agreements and further planning 95 . The BMAP must also affect permitting. NPDES permits may be reopened to incorporate conditions to implement the plan 96 . If the NPDES permit regulates stormwater discharges, the obligation of the permittee is to implement “best management practices or other management measures” “to the maximum extent practicable. 97 ” For other permitted, nonagricultural dischargers, the po llutant reduction actions in the BMAP must be implemented “to the maximum extent practicable. 98 ”

Nonpoint sources are also subject to the BMAP, thus bringing many of them within a regulatory framework for the first time. Nonpoint discharg ers identified in the BMAP must demonstrate compliance by either implementing best management practices (BMP) or conducting water quality monitoring 99 . A further incentive for implementing BMPs is that the disc harger cannot be required to implement measures by “permit, enforcement action, or otherwise. . . 100 ” BMPs provide a presumption of compliance with state water quality standards and release from liability for reimbursement of the Water Quality Assu rance Trust Fund for contamination-related
expenses 101 . BMPs thus set a maximum limit on regulatory requirements.

Mandatory BMPs must be adopted as rules by the DEP, a water management district or the Department of Agriculture and Consumer Services (DACS) 102 . The DEP and the water management districts have authority to adopt BMPs for “nonagricultural, nonpoint pollutant sources” and DACS has au thority to adopt them for “agri cultural pollutant sources”. The

94 Id., §403.067(7)(a)5.
95 Id., §403.067(7)(b)1.b-f.
96 Id., §403.067(7)(b)2.a.
97 Id, §403.067(7)2.b.
98 Id, §403.067(7)2.b.f.
99 Id., §403.067(7)(b)2.g.
100 Id., §403.067(7)(b)2.i.
101 Id., §403.067(7)(c)3. The shield also extends to research proj ects that are developing or demonstrating BMPs. Id.
102 Id., §403.067(7)©. The term “BMP” means what this section actually refe rs to as “interim measures, best management practices, or other measures necessary to achieve the level of pollution reduction established by the department. . .” Id.319 effectiveness of all BMPs must be veri fied at representative sites by the DEP 103 . An initial verification based on “best profe ssional judgement” that BMPs are “reasonably expected to be effective” must be made by the DEP before BMPs are adopted as rules 104 . If water quality problems persist despite the implementation of BMPs, they must be reevaluated and, if modifications are warranted, they must be implemented within a reasonable time period 105

.

STORMWATER

Stormwater conveys nutrients to both ground and surface waters, an d one of the objectives of stormwater regulation has been to control those inputs. In Florid a, stormwater may be regulated by any combination of the DEP, a water manage ment district or a local government, depending on where in the state the discharge is located, the date the discharge was permitted, the activity generating the stormwater and other permits that may be required for the facility.

Florida first enacted a comprehensive stormwater rule in 1982 to regula te the construction of new stormwater systems 106 . Florida’s water management distri cts also exercised authority over the construction and operation of surface water mana gement systems under Part IV of the Water Resources Act of 1972. To streamline permitting and eliminate duplicati on, the regulation of stormwater in most areas of the state today has been consolidated into an Environmental Resource Permit (ERP), issued by either the DEP or a water management district 107 . Older permitting criteria still apply, however, to ma ny existing developments or ongoing vested development activities. A further complexity stem s from the phasing in of requirements for some stormwater dischargers to obtain NPDES permits . In order to achieve more uniformity in stormwater regulation and greater effectivene ss in controlling nutrients, the DEP has begun the process of developing a new statewide stormwater rule 108

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The Water Resource Implementation Rule sets forth the goals for stormwater permitting 109 . In addition to controlling pollution, the program is intended to maintain and restore salinity and flow regimes and groundwater rech arge. It established the policy th at if a stormwater system is designed to meet the criteria of the relevant rule s, there is a rebuttable presumption that it will

103 Id., 403.067(7)(c)3
104 Id.
105 Id., 403.067(7)(c)4
106 Fla. Admin. Code r. 62-35 (2007). The Depart ment of Environmental Regulation (DER), DEP’s predecessor, implemented the rule.
107 The division of responsibility between DEP and the water management districts is governed by operating agreements. ERP permitting authority may also be delegated to local governments. Many local governments also regulate stormwater under their own authorities. Stormwater utilities have beco me a popular means of paying for stormwater system maintenance and improvements.
108 DEP is proposing to develop Rule 62-347. A draft is not yet available. See http://www.dep.state.fl.us/wat er/rules_dr.htm#erp (visited 12-20-07). See also Chad Kennedy, DEP, Florida’s Unified Stormwater Rule Status, Presentation to Lake Okeechobee WRAC, November 2007 (Powerpoint slides available at sfwmd.gov).
109 Fla. Admin. Code r. 62-40.431(2)(a) (2007).320 not cause or contribute to violations of state water quality standards and will achieve an 80% reduction in the average annual load ing of pollutants. For stormwat er discharges to Outstanding Florida Waters (OFW), the goal is a 95% reducti on. Each of the water management districts has rules regulating the construction a nd operation of stormwater systems 110 . The criteria adopted by the Districts vary significantly, but all rely on a combination of retention, detention and filtration to meet the pollutant load reduction goals. A recen t review of the criter ia and site conditions, however, determined that current permitted designs only remove 40-50% of the nutrients 111

. There may thus be no rational basis as to nutrients for the presumption.

A problem with the current criteria relevant to springs protection is that the goals for pollutant removal are intended to protect su rface water quality. The retention of stormwater that recharges groundwater may contribute to ach ieving pollutant reduction goals for discharge to surface water, but it does so at the cost of transferring nutrients to the groundwater system. Only the St Johns River Water Management District has adopted any speci al criteria for groundwater protection. In a Sensit ive Karst Area designated for part s of Alachua and Marion Counties special criteria apply to the c onstruction and maintenance of stormwater basins to minimize the potential for the development of sinkholes or solu tion pipes to convey stormwater directly to the underlying aquifer 112 . These criteria primarily help to maintain whatever treatment is provided by the system as designed and constructed. Defici encies in treatment ca pability are somewhat addressed by requirements for ad ditional soil in stormwater ba sins. In the Wekiva Recharge Area, special basin criteria require the retention of three inches of runoff on certain soils, but do not enhance the removal of pollutants conveyed to groundwater 113

.
ERP permits apply to both construction and opera tion of the permitted system. Once a system is permitted, and constructed as designed, however, the operation of the system is licensed in perpetuity. Provisions for continuing inspectio n and maintenance are relatively weak. Most importantly, there are no means for requiring older systems to be upgraded to incorporate additional pollution control measures 114 . The NPDES stormwater permitting system addresses that deficiency for a limited class of stormwater systems and, for discharges to nutrient impaired waters, the Basin Management Action Plans (B MAP) of the TMDL program may provide a means for requiring increased pollutant reductions.

110 St Johns River Water Management District, I d., 40C-42 (stormwater), 40 C-41 (basin criteria), 40C-44 (agricultural surface wate r); South Florida Water Mana gement District, Basis of Review for Environmental Resource Perm it Applications (July 22, 2007); Northwest Florida Water Management Di strict, Id., 62-346; Southwest Florida Water Management District, Id., 40D-4, 40, 400 a nd Basis of Review; Suwanee River Water Management District, Id., 40B-4.
111 Harvey H. Harper and David M. Baker, Eval uation of Current Stormwater Design Criteria within the State of Florid a (June 2007). Final Report to Florida Department of Environmental Protection, DE P Contract No. S0108.
112 Fla. Admin. Code r. 40C-41.063(7) (2007).
113 Id., 40C-41.063(3)(a).
114 The general prohibition on causing or contributing to violations of water quality standards may still apply to permitted systems, but that is difficult to enforce.321 NPDES Stormwater Permitting

Many of the facilities th at collect and convey stor mwater to navigable wa ters clearly meet the federal definition of a “point source”. Neverthele ss, because such systems often collect pollutants generated by many other sources and are cont rolled by government entit ies, Congress and EPA delayed requiring NPDES permits for many years. In 1990 EPA began to phase in a regulatory
program 115 . Phase I required NPDES permits for larg er municipal separate storm sewer systems (MS4), large construction sites and some industr ial sites. In 1999, Phase II extended the program to smaller municipalities and construction sites 116 . Authority to implemen t the program has been delegated to the DEP, which has adopted NPDES stormwater rules 117 . These rules require monitoring, the development of a Stormwater Management Program (SWMP) and the reduction of pollutants discharged to the Maximum Extent Practicable 118 . Because MS4 stormwater permits cannot be issued for longer than 5 years 119 , permit renewal provides an opportunity not currently available through the ERP program to require improved performance and correction of problems.

SEPTIC TANKS

Septic tank and drainfield system s, a type of Onsite Sewage Treatment and Disposal System (OSTDS), are used to tr eat the wastewater of ap proximately 1/3 of Flor ida’s population and are a major source of artificia l groundwater recharge. 120 . OSTDS are regulated by the Department of Health under authority intended to ensure they do not “adversely affect the public health or significantly degrade the groundwater or surface water 121 .” Traditional sept ic systems can be installed in subdivisions of up to four lots per acre, provided the lots are also served by a public water supply and other criteria are met; if there is no public wa ter supply, the minimum lot size is one half acre. Septic systems must be sized ac cording to the projected load and meet standards for the quality of the soils, depth to the wate r table, and setbacks from wells and surface water bodies. Generally, there must be 42 inches of su itable soils below the bot tom of the drainfield 122
and it must be at least two feet above the wet season water table 123 . The drainfield must be set back at least 75 feet from a permanent surface wa ter and 15 feet from a swale or retention area. These standards provide for treatment of wast ewater by the soil, which is generally very effective in removing pollutants, with the exception of nitrogen. The criteria can be substantially relaxed for lots in older subdivisions 124 . Variances may also be granted to relieve hardships where there is no reasonable alternative for waste disposal and the onsite system will not

115 55 FR 47990 (Nov. 16, 1990).
116 64 FR 68722 (Dec 8, 1999).
117 Fla. Admin. Code r.62-624 (2007) (MS4).
118 Id., 62-624.500.
119 Id., 62-624.300(5).
120 Florida Department of Health, Nitrogen Impact of Onsite Sewage Treatment and Disposal Systems in the Wekiva Area 1-20, 3 (June 30, 2007).
121 Fla. Stat. §381.0065(1) (2007). Fla. Admin. Code r. 64E-6 provides standards for OSTDS.
122 Id., 64E-6.006(1).
123 Id., 64E-6.006(2)
124 Fla. Stat. §381.0065(4)(g)(2007).322 adversely affect health or “significantly degrade the groundwater or surface waters 125 .”

Alternatives to traditional systems that are de signed by engineers to meet specified performance standards must be approved 126 . Alternatives include aerobic treatment systems and so-called performance-based treatment systems. The owner of such a system must have a maintenance contract with an approved entity and an operating permit from the County Health Department 127

. Whenever a central sewer system becomes availabl e, the owner of an OSTDS can be required to connect to it 128

.

More restrictive criteria for OSTDS have been adopted by statute for the floodways of the Suwannee and Aucilla Rivers 129 and the Florida Keys 130 . Recommendations for improved regulation of OSTDS have been made for systems installed in the Wekiva Study Area 131

.
Local governments can adopt more stringent regula tions on construction, performance, setbacks or maintenance of OSTDS through local comprehe nsive plans and land development regulations. For example, in the Green Sw amp Area of Critical State Concern, two counties have adopted ordinances requiring periodic maintenance. Waku lla County requires higher levels of nitrogen removal by installation of performance-based systems.

LAND USE PLANNING AND REGULATION

The loading of nutrients to a spring is closely related to the type, intensity and design of land uses in the springshed. Regulatory programs aff ect land use, but do not provide the kind of comprehensive planning and control of land use that is required to ma nage nutrient inputs. Florida’s growth management system emphasizes the development and implementation of local government comprehensive plans to control land use.

Florida’s Local Government Comprehensive Planning and Land Development Regulation Act 132 (hereinafter “Growth Management Act”) requires every local government in the state to adopt, periodically revise and implement a local comp rehensive plan. The local comprehensive plan must be “consistent” with the goals, objectives and policies of the St ate Comprehensive Plan 133 and the strategic regional policy plan of the relevant Regional Planning Council 134 . The Florida

125 Id., §381.0065(4)(h).
126 Id., §381.0065(4)(j).
127 Id., §381.0065(4)(j)4.
128 Id., §381.00655.
129 Id., §381.0065(4)(t)(2007). The regulatory floodway is defined to include the 10 year floodplain.
130 Fla. Admin. Code r. 64E-6.017-.0182 (2007).
131 Florida Department of Health, Nitrogen Impact of Onsite Sewage Treatment and Disposal Systems in the Wekiva Area 1-20, 16-17 (June 30, 2007).
130 Fla. Stat. ch. 163, Pt. II (2007).

133 Id., ch 187.

134 Consistency of the local comprehensive plan w ith the state comprehensive and the strategic323 Department of Community Affairs (DCA), which oversees the process at the state level, has adopted criteria in Rule 9J-5 for making such determinations. The Act requires that local land development regulations be consistent with a nd implement the goals, objectives and policies of the adopted comprehensive plan 135 . It also requires that land de velopment be consistent with the adopted plan 136

.
A local government’s comprehensive plan must in corporate various elements potentially relevant to protecting springs and springsheds. Four of the most pertinent elements are: a Future Land Use Element; a Sanitary Sewer, Solid Waste, Dr ainage, Potable Water, and Natural Groundwater Recharge Element; a Conservation Element; and an Intergovernmental Coordination Element. The Future Land Use element must include protections for potable water wellfields and protection of environmentally sensitive lands 137 . The Sanitary Sewer, Solid Waste, Drainage, Potable Water, and Natural Groundwater Recharge Element requires identification of natural drainage features/groundwater rech arge areas, assessment of current land use regulations related to these issues, and objectives and policies for imp lementation of land use regulation to protect drainage and recharge functions 138 . The Conservation Element must identify natural resources, including groundwater, and incorporate objectives and policies to conserve such resources 139

. Finally, the Intergovernmental Coordination Element requires analysis of current intergovernmental coordination, specific objectives for futu re coordination, and policies addressing each objective 140 . This element could serve a crucial role in encouraging the intergovernmental coordination nece ssary for effective protection of springsheds that lie in more than one local government. Coor dination also must exist to adequately account for springs protection during water supply planning and under water manage ment district regulatory programs. Sources of water supply adequate to provide for planned growth must be identified. Furthermore, when the Department of Commun ity Affairs (DCA) gives assistance to local governments with their comprehensive plan , DCA must consider , among other things, groundwater recharge.

Local governments must evaluate and update th eir comprehensive plans every seven years 141

. This evaluation includes consideration of water management district regional water supply plans

regional policy plan exists if the local plan is compatible with and furt hers those plans. It is “compatible” if it does not conflict and it “furthers” a plan if it takes “action in the direction of realizing” the goal or po licies of the plan. Id., § 163.3177(10)(b). Consistency with regional water supply plan s is also effectivel y required. Id., §3177(6).

135 Id.,§163.3194, .3202.

136 Id.,§163.3194(1).

137 Id., §163.3177(6)(a) (2006). Fla. Ad min. Code r. 9J-5.006 (2007).

138 Fla. Stat. § 163.3177(6)© (2007. Fla. Admin. Code r. 9J-5.011(2007).

139 Fla. Stat. § 163.3177(6)(d) (2007). Fl a. Admin. Code r. 9J-5.013 (2007).

140 Fla. Stat. §163.3177(6)(h) (2007. Fla. Admin. Code r. 9J-5.015 (2007).

141 Fla. Stat. §163.3191 (2007).324 and any groundwater issues identified as “major” 142 . Water management districts, the DEP, and DCA all have an opportunity to comment on draft evaluations of local government comprehensive plans, thus giving them a chance fo r input to help protect springs. Only the DCA, however, has the authority to determine that an amendment is not in compliance with the Act 143

. The DCA and the DEP have published a handbook for local governments to use in developing policies for springs protection 144 . The DCA is also developing a model land development code for springs protection 145

.
The state has assumed a stronger role in supervis ing local land use decisions in certain areas of the state. The Governor and Cabinet, sitting as the Administration Commi ssion, is authorized to designate “areas of critical state concern” in order to pr omote protection of and reverse deterioration of water resources in those areas 146 . The statute authorizing areas of critical state concern specifically refers to environmen tal resources and aquifer recharge areas 147 . Local comprehensive plans and land development regulati ons must be consistent with Principles for Guiding Development adopted by the Administration Commission 148 . The protection of water quality has clearly been an important goal for designating ACSC such as the Florida Keys, Big Cypress, and Green Swamp 149 . While designation as an area of critical state con cern could help preserve springs, only five percent of the state can be so designated, thus limiting the usefulness of this tool in its current form.

The Florida Legislature can also specifically compel local governments to adopt policies to protect certain springs, In the Wekiva Parkway a nd Protection Act, the Le gislature required each local government within the Wekiva Study Ar ea to amend appropriate elements of its comprehensive plan to implement a master stor mwater management plan and “establish land use strategies that optimize open space and promote a pattern of development on a jurisdiction-wide basis that protects the most effective recharge ar eas, karst features, and se nsitive natural habitats including Longleaf Pine, Sand Hill, Sand Pine, and Xeric Oak Scrub 150 .” Legislation proposed to protect Silver and Rainbow Springs would require local governments to “adopt local land

142 Id., §163.3191(1)© (2007).

143 Id., §163.3184 (2007). DCA’s determination is, howev er, subject to challenge before an administrative law judge and an ultimate de termination, based on that record, by the Administration Commission.

144 Fla. Dept. of Envt’l Protection and Fla. Dept . of Community Affairs, Protecting Florida’s Springs: Land Use Planning Strategies and Best Management Practices (November 2002). Available at http://www.dca.state.fl .us/fdcp/dcp/publicati ons/index.cfm (visited Jan 8, 2008).

145 http://www.dca.state.fl.us/fdcp/DCP/s prings/index.cfm (visited Jan. 8, 2008).

146 Fla. Stat. §380.05 (2007)..

147 Id., §380.05(2)(a) (2007).

148 Id., § 380.05(1)(b)2

149 http://www.dca.state.fl.us/fdcp/dcp/a csc/index.cfm (visited Feb 13, 2007).

150 Fla. Stat. §369.321(3) (2007).325 development regulations and other local requir ements to minimize harmful impacts to the springs 151 .”

State statutes can also limit the authority of lo cal governments to plan and regulate for springs protection. For example, agricult ural activities do not fall within the scope of “development” for the purposes of comprehensive plans and thus are not regulated by comprehensive plans 152

. Local governments may not enforce new regulations for agricultural activiti es already regulated by or subject to best management practices prom ulgated by DEP, the Depa rtment of Agriculture and Consumer Services (DACS), a water ma nagement district, or a federal entity 153 . Thus, while agriculture is a significant cont ributor of nutrients to springs and springsheds in many areas of the state, counties and municipalities have li mited authority to impr ove their regulation of agricultural practices.

Local governments have also been prohibited from restricting th e density of residential and commercial development in so-called “agricultural enclaves 154 ” and cannot impose an “inordinate burden” on any landowner without payment of compensation 155

.
One area of current controversy co ncerns the authority of local governments to regulate the use of fertilizers within their ju risdictions. Many local governments have adopted or proposed regulations to limit nonagr icultural fertilizer use 156 . Apparent concern for regulatory overreaching and confusion led the 2007 Legislature to create a Consumer Fertilizer Task Force 157

. Among the charges to the was the following:

Recommending methods to ensure local or dinances are based on best available data and science and to achieve uniformity among local government ordinances where possible, unless local ordinance variations are necessary to meet mandated state and federal water quality standards 158

.
The Task Force report recommends legislation to authorize local governments to regulate fertilizer use by adopting a specified “model ordinance” 159 . Local governments could adopt more

151 Proposed legislation 10-16-07, a dding section 369.406(1), F.S.

152 Fla. Stat. § 380.04 (2007), referred to by Id., §163.3164(6).

153 Id., §§ 163.3162, 823.14(6); J-II Investments v. Leon Cty., 908 So. 2d 1140 (Fla. 1st DCA 2005).

154 Fla. Stat. §§163.3162(5), (33) (2007).

155 Id., §§70.001-.80.

156 See the undated compilation of “local fertilizer use ordinances” at http://consensus.fsu.edu/Fert ilizer-Task-Force/legislati on.html (visited April 22, 2008).

157 S. 10, Ch. 2007-232, Laws of Florida, codified at Fla. Stat. §576.092 (2007).

158 Id., §576.092(2)(a)4.

159 Florida Department of Agriculture and Consumer Services, Florida Consumer Fertilizer Task Force, Final Report to the 2008 Florida Le gislature, Appendix # 4 (January 15, 2008)326 stringent limitations only under the following conditions: - They have verified impaired waters and are facing existing or possible TMDL requirements (under state and federal laws); or - They have verified harm to human health or harm to the environment that warrants additional consumer fertilizer requirements; or - That they will improve water quality or prevent future impacts of consumer fertilizers on the environment 160 . The 2008 Legislature is considering bills to implement the Task Force recommendations 161

.

Available at http://consensus .fsu.edu/Fertilizer-Task-Force /index.html (visited April 22, 2008).

160 Id.

161 S2352 and H1267, 2008 Regular Session, Florida Legislature.327 

# 8

Synthesis of Nutrients and Springs: Kn owns, Unknowns, Research Priorities and Management Implications

Mark T. Brown Center for Environmental Policy and Department of Environmenta l Engineering Sciences University of Florida

Summary This summary of the state of the knowledge rela ted to nutrient effects on spring ecosystems and biota draws from the proceeding seve n chapters of this report. It represents a synthesis of the most important findings; greater details can be f ound within each of the chapters. This synthesis chapter is organized into three parts: 1) a summary of the “knowns” regarding nutri ent effects on springs and spring biota, 2) the unknowns and suggestions for future research, and 3) suggestions for a research and adaptive ma nagement program for spring ecosystems in Florida.

Florida springs are undergoing ch ange. It is clear that nutri ent loads within springsheds, particularly nitrate-N, from anthropogenic sour ces are increasing. These loads have apparently increased the quantities of n itrogen in shallow and deep ground waters. Nitrogen, but not phosphorus, concentrations in the discharges of ma ny springs have increased in the past several decades. There are numerous other biotic and abioti c factors that have also changed in some, but not all springs, such as increased recreation al use, decreased discharges, decreased O 2 concentrations in discharges, in creased aquatic weed control, invasive species, and increased salinities. By themselves or in combination, these factors may explai n the changes Florida’s springs have experienced in the last several decades.

There is a general consensus in the management community that nitrate-N is an important cause of change in spring ecosystems and biota. In summary, control of nitr ate-N appears warranted, indeed critical, since nitrate-N has been linked to many detrimental impacts in both ecological and human systems (obviously this is true of all watersheds, not just the karst region s of Florida). The fact that nitrate-N is likely to be nearly conservative in the groundwater, and since much of the increased loads are believed to be of human origin, suggests that efforts to control nitrate loads should be focused at the springshed scale and focus on its control pr ior to delivery to the soil and ultimately to the aquifer.

Few studies have been able to provide a threshold value for nitr ate-N at which impacts begin. The literature suggests that in addition to nitrate-N, discharge and dissolved oxygen, which show short-term and perhaps long-term correlation with nitrogen should receive more attention in the328 future as potential causes of ecosystem change in springs. Other stressors in combination or separately may be important causes of change in spring ecosystems as well.

After reviewing the literature, an adaptive mana gement and research program that faces these uncertainties directly and uses them to learn and change is proposed. This proposal combines management interventions with targeted monitori ng to make them more useful as scientific experiments that form the basis for learning and adaptive change. We call for a grand experiment through the creation of Adaptive Sp ringshed Action Programs (ASAPs ) as a way of developing a statewide response to the declines in ecosystem health observe d of many springs. The proposed ASAPs will develop an environment within wh ich meaningful research, education, social learning, and management can occur for the be tterment of Florida’s spring ecosystems.329 INTRODUCTION

This review of the literature was conducted in response to a request by the Florida Springs Initiative of the Florida Department of E nvironmental Protection (DEP) for a summary and interpretation of the available information on th e effects of nutrients on spring organisms and systems. DEP’s mandate was to prepare a re view of what is known about the interactions between nutrients and springs, what information gaps exist that inject substantial uncertainty into management and restoration of springs, and what policy and research coul d be used to address these gaps. This summary of knowledge related to nutrient effects on spring ecosystems and biota represents a synthesis of the most important findings from the literature related to Florida springs along with important finding s studies in other aquatic systems.

To address the complexity of spring ecosystems, we developed a multi-chapter review conducted by several units at the University of Florida (U F) and Wetland Solutions, Inc.. The UF Water Institute provided project management, and other groups were responsible for different chapters in the literature review:

1. Springs as Ecosystems – Wetland Solutions Incorporated 2. Springshed Nutrient Loading, Transport a nd Transformations – UF –School of Forest Resources and Conservation 3. Biogeochemical Processes and Implications for Nutrient Cycling – UF –Department of Soil and Water Science 4. Nutrient Effects on Spring Flora and Fa una – UF Department of Fisheries and Aquatic Sciences 5. Invasive Aquatic Plants and Aquatic Plant Control in Florida Springs – UF –School of Forest Resources and Conservation 6. Effects of Nutrients on Spring Ecosys tems – Wetland Solutions Incorporated 7. Managing Nutrient Inputs to Florida Sp rings: The Legal Framework – UF –Center for Governmental Responsibility 8. Synthesis of Nutrients and Springs – UF –Center for Environmental Policy

This chapter, organized into three parts, summari zes the details from other chapters in an attempt to synthesize knowledge regarding effects of nutrients; recommends additional research for discussion with a broad group of stakeholders; and outlines an approach to future policy and management.

PART 1 - NUTRIENTS and SPRINGS: State of the knowledge

Fate, Transport, and Transformation: What we know

There is growing evidence that nitrate concentr ations are increasing in the water discharging from springs (Chapter 2). Indee d, increasing nitrate conc entrations in freshwater is not just a Florida or karst problem, but ra ther a global problem. Human popul ations have altered the global nitrogen cycle and other biogeochemical cycles th rough land use changes, fertilizer use, fossil fuel combustion and other pathways. With population and land us e change comes nutrient330 enrichment. Florida’s karst region has e xperienced unprecedented population growth and changes in land use over the past several decades, with a consequent transfer of nutrients to the relatively unprotec ted groundwater.

Among the nutrients of concern, nitr ogen, particularly nitrate, appear s to be most problematic in Florida’s karst region. There are four primary reasons for greater concern about nitrate-N than phosphorus. First, increases in concentrations of nitrate-N are nearly omnipresent in areas where anthropogenic loading to the land’s surface has occurred. Second, once in the ground water, nitrate-N appears to be transported as a conserva tive solute, with denitrif ication apparently being negligible (Chapter 2). Third, although Florida’s geology is natu rally rich in phosphorus, there does not appear to be a trend of increasing p hosphorus concentrations in spring discharges (Chapter 2, 4, and 6). Second, because springs are naturally rich in phosphorus, the majority of Florida springs may be historically nitrogen limited (Chapter 3, 6).

Given the importance of nitrate-N and its role as a driver of primary production in many aquatic systems, for better or for worse, the loading, fate, transport, and transforma tion of N appear to be the aspect of springs that has received the most study. Some general prin ciples and caveats have emerged from this work (Chapter 2):

• Emphasis on nitrate-N loading has been pr ioritized because of strong evidence for anthropogenic enrichment on groundwater N in the karst envi ronment of North Florida while contemporaneous evidence fo r broad-scale enrichment of P has not been accumulated

.

• Because there is no geologic store of N, all of the nitrogen emerging in spring vents was deposited on the land surface.

• Where data exists, evidence suggests that overall, the increase s in loading have resulted in 10 to 350 fold increases in concentrations of nitrate-N in spring discharges over the past 50 years. As a consequence of these increases in groundwater nitrate-N concentrations , downstream nitrate-N loads are increasingly rapidl y in many watersheds. For example, a two to three-fold increase in nitrate export to the Gulf by the Suwannee River has been documented over the past 5 years.

• Delineation of springsheds has direct bearing on loading calculations but delineation is not an exact science.

• The underground reservoir from which spri ng flow derives may be extremely large, and hence nominal residence times can be long and extremely variable between springs.

• Some aging measurements of water emergi ng from springs suggest, on average, it has spent between 10 and 30 years in the s ubsurface, but that a significant portion of water (30-70%) has residence times less than 4 years. Relative age contributions can vary from spring to spring.

• Biological fixation of nitrogen has not b een taken into account in most loading studies; an omission that may account for as much as 30 to 35% of the contemporary nitrogen budget.

• Based on differences between N loads applied to the land within springsheds and N in spring discharges, atte nuation withing springsheds appears to be significant331 (~90%), but this apparent attenuation may be the result of lag times between application and emergence.

• Estimates of nitrate and total nitrogen loading in various Florida springsheds suggest that over half the apparent loads to the land surface result from fertilizer application on either urban or agricultural lands. Additional loads result from livestock manure and wastewater septic systems. Atmosphe ric N deposition is generally less than 10% in most studies, but it can account for as much as 26% in some rural springsheds.

• Isotopic characteristics of nitrate-N em erging from springs appears to provide evidence of mixed sources (i.e., mineral fertilizers, organic wastes), but use of isotopes for management purposes require s flowpath tracking and consideration of transformations along the flow path.

Biogeochemical transformations of nutrients in spring systems (C hapter 3) have been little studied. What we know of elemental cycling and nutrient transformations is gleaned from the general literature on lotic environments:

• In general, transformations are most promin ent in biofilms, algal mats, and the hyporheic sediments resulting from a variety of processes including: organic matter decomposition, nitrogen fixation, nitrifica tion /denitrification, iron reduc tion, and sulfate reduction, among others. While these processes have been studied in other lotic environments, there is very little information regarding these processes in spring systems.

• In their historical natural state, most Flor ida springs contain high levels of bioavailable phosphorus (P) (Soluble Reactive Phosphorus [SRP] levels of between 30-60 ppb),

• There are suggestions that once established, algal mats may play a significant role in nutrient transformations and eventual nutrie nt storage, making reversal of current conditions where algae dominate sp ring primary producers problematic.

In summary, control of nitrate-N appears warrant ed, indeed critical, since nitrate-N has been linked to many detrimental impacts in both ecologi cal and human systems ( obviously this is true of all watersheds, not just the karst regions of Fl orida). The fact that nitrate-N is likely to be nearly conservative in the groundw ater, and since much of the incr eased loads are believed to be of human origin, suggests that efforts to control nitrate loads should be fo cused at the springshed scale and focus on its control prior to delivery to the soil and ultimately to the aquifer. Given that springsheds are heterogene ous surfaces of topography, soils, a nd vegetation, with subsurface confining layers that are not ub iquitous, detailed land use pla nning should figure prominently in controlling loads. Best Management Practices (BMPs) in confined regions are of value (although, there is some disagreement on the effec tiveness of BMPs), but in unconfined regions where groundwater recharges rapidly and there is little opportunity to store and treat stormwater on the surface, BMPs alone may be insufficient to prevent groundwater n itrate enrichment.

Effects of Nutrients on Spr ing Organisms and Systems: What we know

There is no question that Florida’s spring ec osystems are undergoing change. An apparent increase in the dominance of filamentous alg ae in many springs, losses of desirable submerged aquatic vegetation (SAV, e.g., native vascular species such as Vallisneria americana or332 Sagittaria kurziana ), increases in invasive vegetation forms, and anecdotal evidence of the disappearance of grazers and hi gher order animals all point to significant structural and functional changes in spring ecosys tems. It is a widely held belief that these observed changes are the result of anthropogenic enrichment of nutrient levels in groundwater discharges and the consequent eutrophication of th e spring ecosystem. Certainly, the correlative evidence (in the time domain) points in that direction, yet the apparent correlation between increased nitrate loading and increases in filamentous algae in spring ecosystems has only anecdotally provided evidence for a causative relati onship between these phenomena (Cha pters 4 and 6). Further, the belief that increased algal biomass has led to th e displacement of native macrophytes appears to be based largely on extrapolation from laborator y studies or qualitative observations (Chapter 4)

. Overall, this review of the literature suggest s, that there is little dire ct evidence linking increases in filamentous algae observed in many springs with increased nitrate delivery to spring systems (Chapter 4)

. This is not to say that causality does not exist, only that direct quantifiable evidence has yet to be demonstrated.

These changes in springs paralle l a eutrophication progression scheme proposed for shallow and relatively open coastal systems (s ee Duarte 1995). This scheme s uggests that increased delivery of dissolved nutrients to aquatic systems will fa vor production of fast-growing algae that extract their nutrients primarily from the water colu mn. These algae may ultimately out-compete and displace slower growing vascular plants that rely primarily on nut rients in sediment and require more light. In the scheme, algal growth can creat e feedback loops that exacerbate the shift to algal dominance and decouple the system from nutri ent concentrations in the water column. For example, increased organic matter or eutrophicati on can lead to decrease d oxygen levels that can stress vascular plants and displace or kill graze rs that might control algae. Although the changes in Florida springs appear consistent with this scheme, other information from spring systems raises questions, including: 1) the possibility that flux of nutrients in springs is a more potent driver of primary produc tivity than water-column concentra tions, 2) the potential that flux obviates nutrient limitation for most flora even at historical and low concentrations, and 3) the possibility that the observed parallels to the eutrophication scheme initially arose from changes to limiting factors other than nutrients, such as light (Odum et al . 1953; Odum 1957) or 4) the possibility that top-down effect s such as loss of higher trophi c level organisms have had a cascading effect on vegetative community structure.

Effects on Flora and Fauna

The general pattern that emerges fr om the review of the literature related to primary producers in Florida springs (Chapter 4) can be summarized as follows:

• Few studies of nutrient limitation for algae exist, and the availa ble results indicate that, if limitation exists, both nitrogen a nd phosphorus limitation can occur.

• There are no experimental studies that unequivocally demonstrate any submersed vascular plant in a Florida, spring-fed system is consistently limited by either nitrogen or phosphorus.

• There are a number of empirica l investigations that suggest the potential for a negative effect of epiphytes or macroalgae on vascular plants in Florida’s sp ring systems, although none clearly demonstrated it or clarified causal mechanisms.333 • Algal mat assemblages may be limited by the av ailability of dissolved inorganic carbon, in Florida’s springs.

• Algal mat thickness in Florida’s springs was weakly and positively correlated with nitrogen and phosphorus concen trations, but cover and biom ass were not significantly correlated with nutrient concentrations.

• Empirical data suggest that there is no statis tically significant link between surface water nutrients and production of macrophytes at this time; however, light availability, especially as mediated by riparian shading, ha s been implicated as a limiting factor for SAV growing in Florida’s spring-fed streams.

• The literature suggests that ni trate concentrations currently observed in the water column of many springs should not limit the growth rates of algae comprising benthic mats. However, this does not preclude the possibili ty of nitrogen limitation in large, welldeveloped algal mats, where access to nutrients may be impacted by diffusion rates and nutrient gradients within the mat.

• The possibility that phosphorus may be a limiti ng factor for algal growth in some springs has been highlighted by recent observations of high N/P ratios in plant and algal samples from several Florida springs.

• Specific conductance (salinity) has been shown to affect the structure of vegetative assemblages in several of Florida’s coasta l, spring-fed systems, perhaps masking the effects of nutrient enrichment.

• There is some evidence from a survey of sp rings in Florida that a negative relationship exists between dissolved oxygen concentr ations and benthic algal coverage.

Nutrients can affect fauna in springs through dir ect toxicity and indirect effects mediated by vegetation. In essence, effects can be generated in three primary ways: 1) toxicity, 2) changes to trophic webs, and 3) changes in habi tats. The literature indicates:

• Nitrate levels in 92% of 130 Florida springs have been reported to average less than 2 mg L -1 ; a value that is lower than most con centrations reported to cause effects on animals (Munch et al. 2006). However, nitrate concentratio ns exceed levels of concern in some springs, and they may reach levels of concern for short periods in other springs, especially after the appli cation of safety factors.

• Invertebrate assemblages vary among differe nt vegetated and unvege tated habitats, which suggests that changes in the types and am ounts of vegetation (e.g., replacing rooted, vascular plants with algal mats) may lead to changes in invertebrate assemblages.

• Differences in invertebrate assemblages were correlated with dissolved oxygen concentrations, which points to the im portance of factors beyond nutrients.

• Fish assemblages and standing stocks vary am ong different habitats, which suggests that changes in habitats may lead to changes in fish assemblages.

• Evidence of changes in fish assemblages base d on repeated surveys in springs that have experienced increases in nitr ate concentrations (e.g., Silver and Wekiva) is equivocal, with significant potential fo r confounding of drivers.

• Fish biomass and phosphorus concentrations were related in a nonlinear manner, with increases in fish biomass being correlated wi th increased phosphorus concentrations only for lower initial concentrations.

• Nonlinear interactions among fauna, flora and abiotic drivers are found in other aquatic systems, and they may operate in Florida’s springs.334 Thus, the effects of nutrients on flora and fauna in springs are likely to ar ise through interrelated and interacting mechanisms. The outcomes bear di rectly on the sustainable use and management of spring ecosystems.

Effects on Invasive Plants and Invasive Plant Effects on Springs

The literature from lake ecosystems demonstrates a positive relationship between nutrient levels and growth potential for non-native “invasive” plants. As a result , there is widespread concern that increases in nitrate-N in springs may be a primary factor driving invasive plant overgrowth. However, the available literature suggests that maximum stimulation of invasive plant growth by nitrate-N is likely to occur at the very low end of concentrations in Flor ida springs and thus the nutrient increases experienced in recent years may not portend increased likelihood of plant invasion. For instance, productivity measurements of Eichhornia crassipes (water hyacinth) and Pistia stratiotes (water lettuce) suggest that maximum grow th potential was possible at nitrate-N concentrations characteristic of historical background concentrations , but that the reason they did not take over may have been the result of downs tream export rather than nutrient limitation.

Overall the interrelationships be tween invasive plants and spring ecosystems can be summarized as follows:

• Available scientific studies have not definitiv ely determined a concentration of nitrate-N in springs that would effectively limit growth of E. crassipes, P. stratiotes , or Hydrilla verticillata (hydrilla).

• There is evidence that growth of H. verticillata and its ability to out-compete Vallisneria
americana (native tape grass) increases as a f unction of increased sediment N (albeit from tank experiments rather than in situ experiments).

• Field surveys in Florida lakes indicate a ma rginally significant relationship between dissolved total N and the extent of H. verticillata occurrence.

• In one long-term study of ecologi cal communities a ssociated with E. crassipes in the St. Marks River, faunal habitat valu es were demonstrated to be similar to that provided by native plant communities.

• Contrary to other ecosystem contexts in which large mats of E. crassipes have been clearly shown to depr ess DO, coverage by E. crassipes in the flowing wa ters of the St. Marks River did not result in significant oxy gen profile differences as compared to Sagittaria kurziana (strap-leaf sag) or other native plant communities.

• The literature suggests that aquatic plant cont rol techniques (e.g. herbic ide applications or mechanical harvesters) that are used to s uppress plant overgrowth have the potential to serve as severe disturbances that could promote succession towards algal dominated springs ecosystems.

Ecosystem Scale Effects

Frequently the changes observed in spring ecosystems are describe d as driven by “bottom-up” or “top-down” control proce sses (or combinations of both). Bo ttom-up control describes processes that originate at the bottom of the trophic ladde r and are driven generally by exogenous forcing functions such as, light, flow regime, temperatur e, or nutrient availability that act to control growth and productivity of primary producers. Top-down control, on the other hand, originates335 at the higher end of the trophic ladder and gene rally takes the form of endogenous effects like changes in grazing pressure, s eed dispersal, and other contro lling feedbacks from within. However, the exogenous verses endogenous effect s should not be taken to mean that top-down only implies factors originating within the ecosyst em. In fact exogenous drivers like the removal of top carnivores or the loss of grazers can ultimat ely result in top-down cas cades of impacts.

Detailed study of two spring eco systems, Silver Springs and th e Rock Creek, Wekiva Spring complex, suggests that there is evidence for ecosystem scale eff ects of nitrate-N enrichment. Using a subsidy stress theory to explain the decline in gross primar y production (GPP), it is suggested that increases in nitr ate-N act to increase productivity for a time, but then act as a stress to depress productivity above a certain threshold (Chapter 6) . One study presents compelling evidence for a decrease in overall productivity in Silver Springs from values measured in a 1950 study as compared to t oday. This declining ecosystem productivity documented at Silver Springs was highly co rrelated with increasing nitrate nitrogen concentrations during the 50-year period of ava ilable data. Declining spring flows, increased shading by riparian trees, and altered fish popula tions were also observed to be correlated with declining ecosystem production at Silver Spri ngs and could offer alternate or cumulative explanations of the observed ecosystem cha nges. The second study conducted in the Wekiva River and Rock Springs Run also found an invers e correlation between nut rient (total nitrogen and total phosphorus) concentra tions and ecosystem metabolism. These spring run ecosystems also had other significant environmental stre sses caused by humans, in cluding decreases in discharge, intensive exotic pl ant management efforts, and di sturbance due to recreational activities. Studies of whol e ecosystem responses to nutrients that would result in direct evidence that increased nutrient levels alone could result in decreased ecosystem productivity and/or photosynthetic efficiency are not available (Chapter 6).

The Legal Framework

The legal framework for managing spring ecosyst ems exists primarily as a collection of regulations for the discharge of wastewater s through the National Pollutant Discharge Elimination System (NPDES). Additionally, certain sources of nutrient pollution that do not fall within the direct jurisdiction of the federa l program, such as nonpoint sources, agricultural discharges, and additional discharges to groundwater are regulated by Florida through various programs (Chapter 7). Narrative nutrient standard for surface waters prohibits altering nutrient concentrations “so as to cause an imbalance in natural populations of a quatic flora or fauna.” Numeric standards can also be developed, as Total Maximum Daily Loads (TMDL), however, development of TMDLs for a particular wate rbody requires that it be “impaired”. Local comprehensive plans and land development regulations can be developed to limit the intensity of land development to minimize potentia l nutrients contributions to springs. They can also provide for improved stormwater and wastewater management.

Overall, the current framework for managing spri ng ecosystems is focused on water quality with little or no ability to manage holistically or adaptively. Springs management would benefit from a framework that includes options for integrating pr ogram initiatives of a variety of agencies and programs including: Florida Department of Envi ronmental Protection, Florida Department of Health, DEPs Bureau of Invasi ve Plant Management, Florida Water Management Districts,336 Florida Fish and Wildlife Conservation Commissi on, and local citizens re presenting important stakeholders.

PART 2: SPRINGS RE SEARCH PRIORITIES

We present recommendations for research before we discuss management implications of the findings of this literature review, because what we don’t know has serious implications for how and why we manage ecosystems. The management of complex, coupled human and natural systems, like Florida’s springs, necessarily requir es an “adaptive” approach, an approach that sees management as an evolving process rather than an end in itself. Management requires knowing, and while we can never know everythi ng, a baseline of knowledge is always a good idea prior to meddling. Therefore, a concerted effort to more thoroughly study and understand springs and their springsheds is warranted.

Florida’s 700+ springs are important resources that are undergoing signifi cant change. They are under assault from a variety of st ressors including reduced flows, increased nutrients (principally nitrate-N), human disturbance, changes in specific conducta nce, decreasing oxygen levels, invasions of non-native species, and so forth. While key elements of spring ecosystems resemble key elements of other lotic systems, Florida spring systems are unique, (indeed, even individual spring systems often differ in key elements) ma king generalizations usi ng key principles and across all springs to be dangerous at best.

Our review of the availa ble literature has shown that we do not have a systematic understanding of historic changes, fundamental processes, or mechanisms of ecosystem function sufficient to definitively manage springs and their contributin g areas. Yet manage we must in the face of significant uncertainty. It is with this uncer tainty in mind that we make the following recommendations regarding research needs. There are many areas of resear ch that if undertaken would yield important information that should lead to better understanding of spring ecosystems. While we present these suggestions one at a time, it is our conc erted belief that the research approach should be a systematic one. An appr oach that incorporates many of the individual suggestions into a holistic comparative, cross sc ale, cross stressor study of a half dozen springs. Obviously the springs need be carefully chosen.

Nutrient Fate, Transport, and Transformation: What we don’t know

Some of the most fruitful research that will im prove understanding of past changes and potentials for future management of spring ecosystems can be found in elucidating factors affecting fate, transport and transformation of nutrients at the scale of the springshed. We offer the following research objectives:

• Explicit accounting of nutrient budgets is essential to understanding attenuation in springsheds (or lack there of). Developing nutrient and water budgets for springsheds has already begun to yield a clear picture of source importance and will continue to help setting realistic target load reduction strategi es. It is a critical fa ctor necessary in the development of TMDLs for spring sheds. As part of a systematic research program, continued development of spatially explicit models of nutrient an d water budgets is337 strongly encouraged to be tter understand the reasons fo r the observed attenuation between springshed and vent.

• Understanding the effective springsheds and flow-paths is essential to establishing baseline conditions and for targeting management for i ndividual springs. The link between discharge and age is an important rese arch need because of its implications for regulatory minimum flows and levels and fo r better understandi ng of the time lags between management action and water quality response. Detailed studies that elucidate spatial variability in recharge and resulting spring discharge are needed for management of individual springs.

• Indications are that Florida’s karst geology does not have an unlimited capacity for P sorption and that future phosphorus release is a potential area of concern. As such, this should be an area of active research, with th e primary objective to determine the degree of sorption saturation along pr incipal groundwater flow-paths.

• Indications are that there may be significant spatial and tem poral variability in DO of spring water that may be related to larger scale springshed dynamics (e.g., organic matter loading of ground waters, ground water age). Res earch is needed to de termine if declines in DO or variability in DO levels are rela ted to the increase in the dominance of filamentous algae in many springs, either th rough negative impacts on grazers, or impacts on SAV respiratory functions.

Within-system nutrient dynamics are not well unde rstood and there is a ge neral lack of data on characterization of spring systems for bioge ochemical purposes. We propose the following studies that will elucidate biogeochemical pathways and pr ocesses, enhancing spring management in the long run

• Our understanding of nutrient dynamics in Fl orida spring systems resembles a black box approach, where changes in water colu mn nutrients provide evidence that biogeochemical transformations are taking place within the hyporheic, detrital, and epiphytic zones, but with little understanding of how or what rate these transformations are occurring.

• Carbon (C) sources and distributi on play a key role in struct uring the biogeochemistry of the spring system; therefore, better understanding is need ed regarding the sources and processing of C in spring systems and especially the role of aquatic plant management in the distribution and fate of C within particular systems.

• The role of internal nutrient supply and re lative bioavailability of various nutrient forms stored in sediments and water column is im portant to document as are the fluxes between sediment, vegetation, and water column.

• Research is needed to understand the fate of nitrogen within the sp ring system. This would include understanding of biotic upta ke processes (alga l/macrophyte uptake), importance of different N sources (NH 4 + , NO 3 - , DON), and the occurrence, rates and importance of biological N 2 -fixation. Whole spring run N bu dgets should be performed to better isolate these processes.338 Nutrient Effects on Flora and Fauna: What we don’t know

The effects of nutrients on spring flora and fa una always will be uncertain for two primary reasons: 1) we cannot afford the amount of sc ience that would come close to eliminating uncertainty and 2) unforeseen and unforeseeable re sponses will arise as emergent properties of these complex systems. In support of sustainabl e use and management of springs systems, we recommend a targeted research program comprisi ng inventories, baseline s, diagnostic studies and other scientific approaches, with choices driven primarily by management questions derived from human values. Given these relationships, input from diverse stakeholders becomes a requirement during development of both “answer able” questions, e.g., questions with welldefined scopes and appropriate levels of precisi on, and detailed research projects designed to provide useful answers. Issues that arose during this review se rve as starting points for such discussions. Key unknowns and examples of research directions are:

• A lack of understanding regard ing the roles of various type s of flora in nutrient budgets hampers management decisions regarding assim ilation of nutrients and setting of numeric criteria for nutrient loads or concentrations; and it can be addressed by diagnostic studies to determine assimilation of nutrients by vascular plants, m acroalgae, epiphytes, microphytobenthos or microalgae found in se diments, and other types of flora.

• A lack of understanding rega rding the links among nutrients and other drivers of primary productivity hampers efforts to predict impa irment, such as overgrowth or shading of vascular plants by algae, set criteria to reverse impairment, or make decisions about restoration; and it can be a ddressed by diagnostic studies th at determine how nutrients, other drivers, and interactions among drivers affect or limit production, competition, and other relationships among periphyton, macroalgae and vascular plants.

• A lack of understanding rega rding the form and magnitude of faunal responses to changes in vegetated habitats hampers efforts to predict impairment, such as loss of grazers or valued species, set criteria to reverse impairment, or make decisions about restoration; and it can be addressed by dia gnostic studies to 1) elucidate changes in composition and function of faunal assemblages related to changes in habitats and 2) determine palatability of flora to grazers, rates of grazing, and the form and magnitude of predation by primary consumers in an effort to identify flor a, grazers, and trophic links to be protected or restored by management actions.

• A lack of understanding regard ing the population and assembla ge effects of ammonia and nitrate toxicity hampers efforts to set criteria and safety factors; and it can be addressed by diagnostic studies to identify and elucidat e the toxic effects of ammonia and nitrate beyond physiological and indi vidual levels of orga nization (e.g., population and assemblage effects) in an effort to determin e threshold levels, safety factors and the need for management actions.

Invasive Plants and Springs: What we don’t know

It is surprising that more research has not b een conducted on the impacts of aquatic weed and invasive plant management on spring ecosystems in general and specifically as it relates to changes in trophic structure, nutrient cycling, and ecosystem productivity. The pervasive nature of these management techniques should provide st rong incentive for a more holistic approach to understanding their short and long term effects. The study of spring ecosystems offer ample339 opportunities to investigate the role of invasive plants and the possible detrimental effects of their management in the changes that are occu rring within spring ecosystems. We offer the following research topics related to invasive plants and their ma nagement in spring systems.

• Diagnostic studies (nutrient and organic ma tter budgets and cycling as well as flora and fauna community structure) before and after aquatic plant control in springs ecosystems are needed to better understand and charac terize the effects on sediment nutrient accumulation and ecosystem succession.

• Since the available literature suggests that even historically characteristic nutrient concentrations at Florida sp rings could support invasive pl ant growth, research into growth dynamics is needed to better target management interventions.

• The seemingly opposing views in the literature that invasive plants cause irreparable change to aquatic systems at all trophic levels a nd studies and observa tions that indicate major invasive plants often provide highly attractive habitat for key springs ecosystem fauna and may support some feedbacks that co uld be expected to reduce algal overgrowth needs further study. Are there circumstances that would potentially warrant the use or existence of invasive and/or non-native plants as a management intervention?

• Finally, given the wide range of concerns about possible algal se lection, habitat loss, faunal toxicology, and dissolved oxygen suppressi on that may be associated with aquatic plant control, a significant di agnostic program of study is wa rranted. This should include detailed in situ studies of the above under differi ng aquatic plant control regimes (including both mechanical and chemical treat ment). The funds for such studies should be included in aquatic plant cont rol budgets as a matter of course.

Ecosystem Scale Effects: What we don’t know

Spring ecosystems are exceedingly complex, but pot entially understandable based on theories derived from the study of general systems. They do not possess a singularity that would suggest that there is a one to one correspondence between a stressor (such as nitrat e concentrations) and a response (such as algal or plan t growth and abundance). We know that springs, like other ecosystems, have feedback and control pathwa ys that make predicting cause and effect relationships somewhat problematic. Yet that is exactly what management interventions require. Given X intervention, we expect Y result. Unfortuna tely the result is ofte n Z; something we did not intend at all.

Ecological monitoring of the multi-level effects of stresso rs on ecosystems is important for understanding the full implications of management interventions. It is not sufficient to monitor only the target of a particular in tervention, because freque ntly the target is not the true indicator of the problem we are trying to address with the intervention or the element that responds the quickest, or responds at all. Whole system monitoring is essent ial if management is to learn from past and refine future interventions. W ith this in mind we offer the following “whole ecosystem” research agenda.

• A critical need exists to establish a comprehensive baseline database for a large number of spring ecosystems, including their range of ecosystem me tabolism, trophic structure,340 and key forcing functions. Data need to be co llected at a spatial and temporal resolution to facilitate separation of transient behaviors from long-term trends.

• Elucidating changes in primary production, eco logical efficiency, species composition, and succession, that result from changes in forcing functions such as nutrient concentrations, dissolved oxygen, salinity, temper ature, and shading are necessary if we are to understand ecosystem management interventions. This can be accomplished through in-situ whole system studi es in a variety of springs.

• Elucidating effects of changes in herbi vory, which could lead to primary producer effects that appear similar in symptom to nutri ent effects is an important are of research since the strength of bottom-up vs . top-down effects is a central question to be answered.

• In-situ mesocosm experiments are also needed that involve controlled manipulation of one or more environmental forcing functions (temperature, discha rge, carbon dioxide concentration, nutrient concentr ations, oxygen concentration, specific conductance or salt contents, and grazing pressure) and the responses of individual spec ies and trophic levels. Any changes in individual or ganisms could then be related to changes in ecosystem structure and functioning and coupled with whole ecosystem studies (above) might help to elucidate feedbacks a nd heuristic potentials.

• Indicators and assessment tools for spring ecosystems are essen tial. Understanding change requires indices that are sensit ive enough to detect change. DEP’s Stream Condition Index has been developed and used to assess a wide variety of stream ecosystems of Florida, yet we are uncertain if it is applicable in its present form to Spring ecosystems. Research is needed to develop appropriate indices of biotic integrity for spring ecosystems.

Summary of Research Recommendations

During one of the public meetings that preceded the final production of this document, an attendee expressed extreme frustration because “w e scientists” were pontificating on the cause while the patient was sick and dying on the oper ating table...a particular ly vivid and apropos analogy and one that has stuck with us. To carry the analogy a bit further...we have the luxury of multiple sick patients, multiple poten tial diseases, and relatively low-risk of clinical trial error as a result of the ecological changes now being expe rienced by a diversity of Florida’s springs. We propose using this opportunity to explore the causes and long-term implications of these changes. Therefore we propose that a number of smaller spri ngs be selected that represent a variety of stressors acting separately and in combination as a means of te sting various hypotheses regarding the impacts of stressors on spring ecosystems and biota. Tests might involve, for instance, choosing springs with low vers us high DO, presence and absen ce of grazers, low versus high nitrate-N, absence versus presence of recrea tion and so forth; then changing environmental conditions such as nitrate-N concentrations or DO concentrations in spri ngs to support grazers, etc. These experiments should be accompanied by sufficiently fine-grained monitoring, both spatially and temporally, to discriminate the effects of manipulations from other potential interactions. The experiments also need to be carried out over sufficient lengths of time to elucidate the possibility for alternative stable stat es and hysteresis effects in which, even after long periods of time, the state of the spring syst em may be determined by its history. In other words removing a stressor may not be sufficient to “restore” a spring ecosystem to a former state341 as there may be considerable lag in spring ec osystems between removal of a stressor and the effect of its removal.

In addition to the above, we suggest a relatively intense monitoring campaign in a like number of springs that is designed to eluc idate spatial variability and tem poral dynamics of stressors and impacts. Most of the current spring monitoring data has a reso lution that makes discriminating spatial variability impossible and lacks the tem poral sensitivity to ac count for potential pulse events that may have significant impact (low DO/nitrogen/high sa linity events, for example).

Scanning though the list of resear ch suggestions, it is apparent that answers to the many uncertainties surrounding Florid a Spring management will not come quickly. A sustained research program will necessarily require a su stained source of funding. Here we suggest a simple source of funding that is tied directly to the resource a nd that is independent of water management district funding or state tax dolla rs....a source of funding that with appropriate social marketing can be thought of as a positiv e contribution to natural resource management instead of a tax. The term social marketing in this context means that through a marketing campaign we would enlist the cooperative buy-in of those affected as a contribution rather than imposing a tax.

The proposed source of funding is a “One-cent cont ribution” for sustainable springs research and management from each bottle of water sold in Florida. According to the International Bottled Water Association , total U.S. sales of bottled water volume surpassed 8.82 billion gallons, in 2007. That translates into an average of 104 sing le-serve bottles per pe rson per year. With 18.68 million people in Florida, that equals 1.94 billion bottles of water sold in Florida in 2007. If one cent per bottle were contributed to ward springs research, the sale of bottled water in Florida would generate $19.4 million/yr (this does not incl ude tourism sales of bottled water which could easily push the total available funding to nearly $21 million per year). If ma rketed correctly this could result in a sustainable sour ce of funding for research and management of Florida’s spring ecosystems, obviously some of the most important natural resources the State possesses. The authors recognize that this is no easy sell, for th e first reaction to the proposed funding source is that it looks like a tax. However, we emphasize that it is not a tax, we do not solicit legislative intervention to impose a ta x, rather we believe th at if presented correctly to the bottling industry they would be extremely willing to add a sentence on thei r bottling labels indi cating that with the sale of each bottle of water they are contribu ting to research that will ultimately lead to sustainable management of Florida’s springs.

PART 3: RECOMMENDATIONS for SPRINGS POLICY and MANAGEMENT

There is no question that Florida’s springs ar e important, unique resources and as with most natural resources in Florida, they are under consid erable pressure from changes in environmental drivers, direct human use, and indirect alteration of the chemi cal and physical environment in which they exist. Equally apparent is the fact th at many Florida springs are exhibiting change in biological components. While exact causes are un certain, in the previous sections, we document numerous stressors that might be grouped into at least four general classe s: physical disruption, changes in quantity of discharges, changes in chemical composition of discharges, and changes in populations of higher troph ic organisms which have top-down cascade effects on lower342 trophic levels. It is quite possi ble but highly unlikel y that the observed ch anges in biological indicators in some springs may be caused by a si ngle stressor. More probab le is the possibility that springs exhibiting change ar e affected by combinations of stressors, acting in a cumulative manner. In some systems, management interventions can be tailored to a particular stressor and the trajectory of change it produces . However, it is unlikely at th is time that our ability to understand causality between stress ors and indicators in spring ecosystems is sufficient to allow such a fine grain approach to management. Th us, it is our opinion that a holistic (or systems) approach to management is required.

A systems approach is nothing new; there has been a call for an ecosystem approach to management of natural resources for some time (Agee and Johnson 1987). Defined as the integration of ecological, social , and economic objectives for the planning and management of natural resources, early attempts at ecosystem management sought to maintain and enhance biological diversity and ecosystem integrity (Gru mbine, 1994). The focus on integration of these ecological objectives with the simultaneous attain ment of social and econo mic objectives is what distinguished ecosystem management from earlie r natural resource management approaches.

Recognizing the often pathological insistence of managers to st ick to one narrative and one management scheme with little or no flexibility for new information, resulted in a call for an approach that addressed not onl y a whole systems approach to management but an adaptive one as well (Holling, 1978; Walters, 1986).

The Springs Initiative: Pr otecting Florida’s Springs

In 2001, the Department of Environmental Protec tion (DEP) made springs protection a priority through the establishment of the Florida Sp rings Initiativ e (FSI). Since then, DEP committed several million dollars to fund sc ientific research, water quality and biological monitoring, education and outreach, landowner as sistance projects, and springs re storation (FSI, 2007). In the six years since its establishment, the FSI ha s accomplished much in the four areas of its responsibility:

• Research and monitoring;

• Landowner assistance in the form of planning, management, and restoration;

• Restoration/protection ac tivities and educational out reach in state parks; and

• Environmental education and outreach.

Cited as an “adaptive management approach” the resources and staff of the FSI are tailored to future management activities based on the past experiences of the program (FSI, 2007). While FSI’s approach involves change and might be calle d adaptive in the sense that the FSI tailors future activities to what has been “learned” from past activities, it is not Adaptive Management (AM) by the strict characterization suggested by the Resilience Allianc e (RA, 2008) and others (Holling, 1978, 1986; Gunderson et al. 1995; Ha bron, 2003: Folke et al. 2005). A casual understanding of Adaptive Manage ment leads to the belief th at monitoring management activities and occasionally changing them c onstitutes an Adaptive Management approach. Contrary to this commonly held belief, AM is much more than simply changing future activities based on past experience. This is not critic ism of the current approach of the FSI, only343 clarification since the following proposals for policy and management of Florida’s springs revolve around an AM strategy. We believe that the complexity of the issues as well as the complexity of the springs’ social ecological sy stem, comprised of a multi tude of actors from microbes to state government, requires a true AM approach. What follows then, are suggestions for establishing an AM program and framework for Florida springs.

Adaptive Management: An Ap proach to Springs Protection

An adaptive approach to natural resource manage ment involves exploring alternative ways to meet management objectives, pred icting the outcomes of alternativ es based on the current state of knowledge, implementing one or more of thes e alternatives, monitoring to learn about the impacts of management actions, and then usin g the results to update knowledge and adjust management actions (Murray and Marmorek, 2004) . AM is a dynamic process that requires learning and adapting. This means that an AM program incorporates , as an integral part of the program, management interventions that are developed and carried out as scientific experiments. Through the experiments, the AM team learns a bout the systems but more importantly, they are learning whether the intervention is likely to resu lt in the desired objective. As an experiment, the management intervention must be monitored and data that is gathered must be analyzed to yield information that can direct management. Finally, the management team must be flexible enough to incorporate new informa tion and adjust management if the experiment suggests that adjustment is necessary. A main organizational prin ciple of AM is that it requires partnerships of policymakers, managers, scientists, and other stak eholders who learn together how to create and maintain sustainable resource systems. In other wo rds, AM is a social as well as a scientific
process.

A Tragedy of the Commons

Unlike Hardin’s (1968) over simplified dilemma of the commons involving herders use of rangelands, Florida’s springs are nestled in a comp lex world of physical, ecological

, economic and political realities. As such, Hardin’s twofold institutional arrangement for protection , either centralized government or private property, cannot work to protect the “springs commons ” over the long run in the face of over exploitation. Th e complexity of springs issues do not lend themselves to traditional single focus resource mana gement that uses inform ed- trial-and-error of applying management interventions, watching the impacts, seeing what works, and then adjusting future interventions according to observ ations. Nor does their re solution lie in either private ownership or top-down govern ment control. Instead, the complexity of issues and causes must be addressed in an informed process of Ad aptive Management that is inclusive, holistic and responsive.

The springs’ commons have a multitude of players: from herders, to agriculturalists, to boaters, to government, to scientists each with their own understanding of the issues, causes, impacts, solutions and futures of Florida’s springs. Th e adaptive management framework not only must include all these stakeholders, but must have the capacity for social learning (Bandura 1963), or the ability for the entire fram ework to learn and exchange information that is linked to management decisions. Thus, central to the AM framework for springs is the development of a social fabric that is inclusive and that can address links between institutions, communities,344 knowledge, science, and power (Stringer et al. 2 006). This requires that information flows be horizontal, among groups, as well as ve rtical, among institutional levels.

Managing in the Face of Uncertainty

The National Research Council (20 04) defines adaptive management as a process that promotes flexible decision-making that can be adjusted in the face of uncertainties as outcomes from management actions and other events become better underst ood. The careful monitoring of outcomes both advances scientific understanding a nd helps adjust policies or operations as part of the iterative learning process. AM is not an e nd in itself or a trial and error process, but rather a means to more effective d ecisions and enhanced benefits through learning while doing.

Springs are complex, coupled natural and social systems, sometimes called social ecological systems. Their complexity lead s to uncertainty about the exact structure of the system, the relationships between components, and causes a nd effects of actions, all of which lead to significant uncertainty regarding the outcomes of management decisions. In the face of some ecological change (or crisis), the approach most often used is to develop an informed trial and error approach to management that first tries the most obvious intervention. The intervention generally is the result of a rigid idea of cause and effect, which leads to relatively inflexible management in the face of new information and often results in effects that were counter to expected, sometimes called “ecological surprises”. The alternative is the generation of several hypotheses for the observed change and then in terventions based on the hypotheses and designed as scientific experiments. If posed correc tly and followed up with sufficient monitoring, management interventions become the basis for learning, and if the management framework is sufficiently flexible, the basis for adaptive change.

In the face of changing spri ngs ecosystems, the causes of wh ich are somewhat uncertain, we must learn as we manage. The opportunity to impose an adaptive management framework on the process so that research gene rates new knowledge and informati on that flows di rectly from management decisions is irresistible and t oo good to pass up. Thus, we suggest a springs management framework that faces uncertainty head on and uses it to learn and change, instead of avoiding risk and uncertainty a nd preserving the management syst em. There is a fundamental difference in approach between management th at is adaptive and Adaptive Management; the former stresses preservation of the status quo, a voidance of risk, and concern with the costs of management. The latter accepts change and the uncer tainty that comes with it, and is willing to fund the processes necessary to avoid ignorance.

Adaptive Springshed Action Programs (ASAPs)

What is required are new institutions and institutional frameworks as well as new science and experimental frameworks. As a grand experiment, we propose th e creation of up to six Adaptive Springshed Action Programs (ASAPs). Calle d Action Programs instead of management programs, we wish to stress that these progra ms are more than management. They are new management, research, and education structures th at incorporate institu tional verticality (i.e., state, regional, and local institutions) along w ith a horizontal group structure of socio-economic stakeholders, managers, and scientists organized as a working group to adaptively manage the345 spring and its springshed. They are akin to BM APs but contrary to the BMAP/TMDL program, waters do not have to be declared impaired in order to implement an ASAP. All their implementation requires is the acceptance that a spring ecosystem is ch anging, a belief that through appropriate management it is possibl e to affect positive change, a willingness to implement interventions as scientifically valid e xperiments, the ability to monitor, collect and analyze data, and finally the read iness to change interventions should the data suggest it is appropriate to do so.

The springsheds should be chosen to represen t a cross-section of geographical, ecological, physiological, sociological, and polit ical settings. As a first cut, we suggest the following eleven spring systems, from whic h six could be chosen:

• Silver and Rainbow Springs in Marion County – two of the state’s largest and most economically important springs with rising nitrate concentrations

• Ocala National Forest Springs (Alexander, Juniper, and Silver Glen) – three of the State’s most protected springs that are curre ntly unaffected by any nu trient increases or flow reductions

• Wakulla Springs – an important north Florida spri ng with a history of springshed interventions to reduce nitrate-N.

• Kings Bay/Crystal River Complex – a complex of springs with a long history of management interventions re lated to invasive species

• Homosassa Springs Complex – a system with significant ba seline data and a recent loss of flora

• Chassahowitzka Springs Complex – a system adjacent to the Homosassa system with a similarly comprehensive baseline and less development in its springshed

• Weeki Wachee Springs – on the west coast of Flor ida the springshed has undergone significant development in the last two decades

• Ichetucknee Springs – a long history of recreational use with several different spring boils with differing chemistry a nd a newly urbanizing landscape

• Wekiwa /Rock Springs Complex – a highly urbanized springshed, with decreased discharged volumes, intense recreational use and important regulatory initiatives for protection

• Suwannee River Springs (Fanning and Manatee) – springs in an ag ricultural landscape with the highest nitrate N concentrations in the state

• Wacissa River Springs – springs in a relatively unde veloped rural landscape with minimal nutrient impacts

By choosing several spring systems simultaneous ly, we suggest that communication between spring ASAPs is just as important to manage ment, learning, and science as within ASAP communication. Frameworks for communication betw een them must be facilitated, encouraged, and even required. Since each spring systems is a test case in a broader study of adaptive management it is imperative that cross system communication, comparison and ultimately learning take place.

Part of ASAP framework should be workshops th at build a collaborative space for discussions between the policy, science, user, and manager gr oups. These discussions should focus, using a346 “systems approach”, on elucidating leading pr oblems and issues within each of the spring systems, developing hypotheses co ncerning the causes of problems and effects of interventions, and develop scenarios about the future. Th ese workshops should be a common form of participant interaction, feedback and formalization of the reflective/iterative processing of information that is required of the adaptive pr ocess. A deeper unders tanding of the resource issues, management options, and futu re scenarios are likely outcomes.

The current bag of spring management tool s (PLRGs, TMDLs, BMAPs, MFLs etc.) is appropriate and usable within the context of the ASAPs. Ultimately whatever tools are used, the allocation of responsibility for reducing impacts on springs (wheth er impacts are from nutrients, reduce discharges or human disturbance) will require negotiation. Who pa ys and how much are part of reducing impacts. The proposed ASAP s represent a democratic, responsive bottom-up management structure that can garner support and allocate responsibilit y without heavy-handed top-down regulation.

It should be noted that the ASAP approach sugges ted here is in addition to the needed research proposed in the previous section of this chap ter. There needs to be multiple layers to understanding and managing spring ecosystems with efforts at the large scale to control things that matter at that scale and science at the sm all scale to better unders tand which interventions are most likely to yield desired ecological effects. These are not mutually incompatible, and are likely wholly complementary. The research propos ed in the previous section is designed to provide baseline information a bout spring ecosystems and res ponses of spring ecosystems to stressors. The ASAP program suggested here is a novel approach to management that includes science and the ability to test hypotheses con cerning appropriate management interventions for problems believed to be at the root of changes in spring ecosystems. The proposed approach to management is similar to, and in fact could us e the existing spring work ing groups as a starting point. So far, the working groups seem to be less about management and more about exchanging information, as suggested in the following stat ement by a participant in spring working groups:

"Springshed working groups provide a marvelous opportunity for collaboration, exchange of ideas, and general education. As someone who has participated in a number of different worki ng groups, my impression is that while there is clear success in bringing stakeholders together and sharing information, there is much room for improvement in terms of facil itating bottom-up discussion about how to process and use this information movi ng forward. All too often, working group meetings are characterized by a series of presentations that are not followed by serious discussion, synthesis, and ar gument among stakeholders about the information presented, much less the articu lation of different ideas about how to proceed in the face of this information. Exchange of information is undoubtedly a good thing, but we need to come up with better ways to take advantage of multistakeholder expertise in progressively ev aluating the big picture at each spring and adapting the research and policy agenda over time. The interest and energy to do this is there. What we need is a formal recognition th at critical dialogue, collaborative development of resear ch and management priorities, and collaborative evaluation of research and management results should be integral parts of the working group model."347 Our proposal for the ASAPs would build on the working groups model, develop overall management goals and objectives, and then initiate a round of management interventions to test hypotheses concerning the causes of current appa rent declines in spring ecosystem condition.

The social framework of the ASAP team shoul d include a tripartite of individuals from government, science, and the citizenry at large. The ASAP team should be empowered to make management decisions and direct the relevant agencies to carry out the interventions. For instance, they may direct the appropriate agency to limit recreational use, or eliminate spraying of aquatic weeds as part of an overall manageme nt strategy to restore SAV vegetation in a spring run. The appropriate agencies as members of the ASAP team woul d develop the necessary policy and implementation strategy.

In summary, the key difference between the ASAP s and the current springs working groups is related to membership, organization, and empo werment. Through agreements between agencies, the ASAP will be empowered to take action after developing a program that includes goals and objectives, management interventions that are te stable and capable of being validated through collection of data, the collection and synthesis of that data, and finally, real time reflection on whether the interventions are succeeding. The DEP’s springs in itiative could coordinate the ASAPs. We propose the DEP utilize a neutral facilitator whose primary job would be the development of consensus between th e tripartite member s of each ASAP.

Adaptive Spring and Springshed Research

Research within the AM context is designed fo r the purpose of producing results that may be applied to and answer management questions. Just as this concept of research differs from “basic research” (investigations of a phenomenon without reference to particular human needs and wants) the science that drives the AM research is different. Science within the context of AM differs radically from classical notions of “pur e science” as it must re spond to often-conflicting group demands of managers, politicians, citizens, developers, environmentalists, and users. It must be able to operate under conditions of comp lexity, uncertainty and risk, and generally be applied over large spatial and temporal scales. In addition AM research must cross many ecological and organizational boundar ies increasing its complexity and the lead time required for accomplishing its objectives. Thus, the payoff from AM research should be measured over the course of years, not months. Often this is difficult for the management and policy arena to accept. Frequently policy makers want answers before science has even formulated the question, wishing for a synthesis of availabl e information with inductive infere nce to the particular case in point. Managers want to move forward with sc ientific certainty that their intervention will achieve the desired result. Science is left wa nting; wishing for the time to do the research, synthesize the results, a nd answer the questions.

The research within each of the ASAPs should ultim ately be formulated by their participants, as they each have their particular ecological is sues, management potentials, and socio-economic constraints. Overall, our unde rstanding of Florida sp ring systems would benefit from cross system comparison. It is important that the science/rese arch component foster the development of broad integrative studies that bring policy a nd management into focus and generate the kinds348 of knowledge and understanding that will inform and develop better management as well as social learning rather than just the accumulatio n of more information. In the absence of the ASAP system and the integrative program of st udy it would foster, th e individual research questions given in the preceding section of this ch apter remain as important research priorities. Most are significant questions about the contro lling factors of ecological change in spring ecosystems. However, the creation of ASAPs and the opportunities for real time research, education, management and social learning that they will afford is truly a grand experiment and the most important opportunity provided by the crises of ecological change now being experienced by Florida’s springs.349 REFERENCES

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