Summary and Synthesis of the Available Literature on the Effects of 

Nutrients

on Spring Organisms and Systems

Mark T. Brown & Kelly Chinners Reiss

Center for Environmental Policy

Department of Environmental Engineering Sciences

Matthew J. Cohen & Jason M. Evans

Forest Water Resources Laboratory

School of Forest Resources and Conservation

Patrick W. Inglett, Kanika Sharma Inglett & K. Ramesh Reddy

Department of Soil and Water Science

Thomas K. Frazer, Charles A. Jacoby & Edward J. Phlips

Department of Fisheries and Aquatic Sciences

Robert L. Knight & Sky K. Notestein

Wetland Solutions, Inc.

Richard G. Hamann

Center for Governmental Responsibility

Levin College of Law

Kathleen A. McKee

University of Florida Water Institute

28 April 2008

University of Florida Water Institute

570 Weil Hall

PO Box 116601

Gainesville, FL 32611-6601
ii

Executive Summary 

This summary of the state of the knowledge related 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 Florida Department of Environmental 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 University of Florida’s Water Institute, Center for 

Environmental Policy, School of Forest Resources and Conservation, 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 consulting 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 Implications for Nutrient Cycling 

4. Nutrient Effects on Spring 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 Florida Springs: The Legal Framework 

8. Synthesis of Nutrients and Springs 

DEP’s mandate was threefold: 1) to prepare a review 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 management, and 3) make recommendations for public 

policy and research to address these gaps. The following summary draws the most important 

highlights from each chapter according to this mandate. 
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Chapter 1: Springs as Ecosystems 

This chapter focuses on describing the abiotic and biotic components of spring 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 existing literature on spring ecosystems there 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 temperature 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 biological systems that are highly efficient at 

converting solar energy into useful productive work. 

• During the past century humans have been exerting an increasing variety of external 

and internal stresses on spring ecosystems in Florida, the most pervasive include: 

o Alteration of discharge regimes resulting from increased groundwater 

withdrawals, 

o Increased levels of nutrients, particularly nitrate-N in groundwaters discharging 

from springs, 

o Extensive recreational disturbance of springs and spring runs, and 

o Increased disturbance in the 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 can 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 further develop these concepts the 

following research directions are suggested: 

• Development of a holistic management strategy 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 such as monitoring of ecosystem metabolism may provide 

a more sensitive and quantifiable indicator of the response of springs to nutrient 

increases and the myriad other 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 Loading, Transport and Transformations 

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 discusses intrinsic 

vulnerabilities of karst areas to nitrogen loading, and examines the effects of changes land 

use/land cover and hydrology on various aspects of water quality decline. The chapter ends with 

an examination of the state of our understanding of springsheds 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 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 springshed management, 

• Karst landscapes are vulnerable to pollutants from anthropogenic activities at the 

surface. It is clear that the risks of aquifer contamination are greatly attenuated in regions 

of the state where regional confining layers (e.g., Hawthorn Formation) limit the rate of 

interaction between surface and Floridan Aquifer waters. 

• Water quality is changing in dramatic ways including: nitrogen enrichment, hardness, 

dissolved oxygen concentrations, and trace quantities of pharmaceuticals. 

• The preponderance of N pollution appears to be from fertilizer sources. While there are 

several reasons to treat this finding as an over-generalization, most of the accumulated 

evidence from mass balance computations and isotopic tracer studies suggests that 

mineral fertilizers, and therefore not septic tanks and wastewater sprayfields, are the 

principal sources of N pollution. However, since fertilizer use is highly diffuse and based 

on hundreds of thousands of individual decisions, control of N from point sources such as 

municipal wastewater effluent disposal activities 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 transformations injects substantial 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 mixture 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 water quality response variables. 

• What are the sources of N in springsheds? While there is accumulating evidence that 

the source of N is principally mineral fertilizer, there remains significant uncertainty 

about the interpretation of bulk 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 studies of other water quality parameters. 

• What is the assimilative capacity of the landscape 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 activities and declining water quality in spring 

ecosystems, evaluating the extent to which parameters are management-sensitive is 

relevant to both timelines for recovery and prioritization of management efforts. 

Chapter 3: Biogeochemical Processes and Implications for Nutrient Cycling 

The potential biogeochemical reactions and pathways occurring in springs are numerous, and 

have been studied in varying levels 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 literature related to biogeochemistry in Florida spring 

systems, and 3) a discussion of the needs and directions for future spring research. 

The Literature: What is Known

• In springs and spring runs, biogeochemical 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 biogeochemical processes on spring water 
vi

composition begins immediately upon emergence from the boil and continues throughout 

the entire length of a spring run. 

• Key interfaces involved in elemental 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 decomposition, nitrogen fixation, 

nitrification, denitrification, iron reduction, sulfate reduction, and associated processes 

involving phosphorus. 

• Organic matter compositions and distributions greatly affect the variety and 

importance of biogeochemical processes especially those requiring 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 physico-chemical parameters (e.g., 

dissolved oxygen). Some of the observed association of increased nitrate with increased 

algal growth could be the result of other 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 reactions that are occurring in these systems. 

• There is little information to ascertain 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 alternative 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 explain 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. heterotrophic) involved in spring processing/cycling 

of nutrients. 

o Basing the collection of water quality data on upstream and downstream locations 

of stream segments will improve our ability to understand processes and spring 

run characteristics contributing to nutrient uptake and transformation. 

o Characterization of sediments (e.g., nutrient storages, organic matter distributions, 

mineral composition) would enable better assessments of benthic and hyporheic 
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processes affecting current and future storages and fluxes of sediment nutrients 

(sediment memory). 

o Studies are needed to document the potential for aquatic plant management efforts 

to alter biogeochemical conditions (i.e., creating anoxic zones) and nutrient cycles 

(e.g., increasing denitrification loss of nitrate or increasing release of ammonium 

and phosphorus). 

o More studies are needed to document the involvement of individual system 

components 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). 

Chapter 4: Nutrient Effects on Spring Flora and Fauna 

Many of Florida’s springs and spring runs are enriched in nitrate due to broad-scale 

contamination of groundwater supplies, which fosters two primary and interrelated concerns 

regarding the effects on flora and fauna in spring systems: 

1) high nitrate concentrations can promote 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 native vascular plants (see 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 consistent with the proposed 

eutrophication progression scheme, which a cue that management 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) 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 concern in most spring 

systems. 
viii

o Toxicity needs to be factored into management of spring systems, especially 

given uncertainty factors that could 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, such as grazing pressure, habitat use 

and predation pressure. 

o The outcomes of ecological interactions clearly relate to sustainable management 

of spring systems 

The Literature: What is Not Known

• There are few quantitative data that definitively link changes comprising the 

eutrophication progression scheme to increased nitrate loads in spring systems 

o The eutrophication progression scheme predicts that nonlinear responses may 

have “decoupled” these systems from contemporary 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 ecosystems would benefit from compilation of 

inventories and collection of baseline data. Initial suggestions for discussion include 

inventories and baseline data collection designed to: 

o Measure nutrients, oxygen, carbon, light, flow, substrates, grazing, algae, plants, 

invertebrates, fish, and other key drivers and valued components of spring 

systems chosen with the best available information and, eventually, with guidance 

from the results of diagnostic studies. 
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o 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). 

o Sample synoptically or at spatial and temporal scales that support rigorous 

analyses of interactions (note that these inventories and baselines form the 

platform for adaptive management). 

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

• Diagnostic studies focused on bottom-up and top-down interactions are also critical. 

Initial suggestions include the following 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 drivers affect overgrowth, shading and other 

relationships among periphyton, macroalgae and vascular plants in an effort to 

ascertain the need for management actions, 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 plants that are hypothesized or known 

to play major roles in nutrient assimilation 

− including potentially limiting factors other than nitrogen and 

phosphorus and interactions among limiting factors that can 

significantly influence the outcomes of management (e.g., 

micronutrients, iron, oxygen, carbon, light, flow and substrate 

type) 

− exploring relationships and consequences across multiple levels of 

biological and ecological complexity, including uptake and other 

physiological 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 relationships form the core 

of an adaptive management approach) 

− elucidating the direction, magnitude and form of limitations (e.g., 

nonlinear and nonreversible) to identify suitable targets and goals 

for management actions. 

o Identify and elucidate changes in composition 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 grazers, rates of grazing, and the form and 

magnitude of predation by primary consumers in an effort to identify flora, 

grazers and trophic links to be protected or restored by management actions. 

o Identify and elucidate the toxic effects 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 hyacinth, 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 approximately $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 their control in Florida, 2) the growth potential 

of the nonnative plants in springs as a function of elevated nitrate-nitrogen concentrations, 3) the 

social and ecological consequences of aquatic plant overgrowth, 4) the ecological risks 

associated with current aquatic plant control methods, and 5) the potential benefits of alternative 

aquatic plant management approaches 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 Florida began with the introduction of water 

hyacinth, into the St. Johns River in the late 19th 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 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 present in Florida before European colonization, 

or was introduced by early Spanish settlers. Chemical control programs have maintained 

water lettuce at low levels throughout 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 associated 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 allelopathic compounds on algal dynamics in 

springs ecosystem are not presently known. 

• Ecosystem surveys indicate that water hyacinth, water lettuce, and hydrilla provide 

attractive habitat for crayfish, 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 currently 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 worthwhile in highly degraded springs ecosystems 

where these plants are established. 

• Biotypes of hydrilla that are resistant to fluridone, a systemic herbicide commonly used 

for hydrilla control in Florida lakes, have been documented in recent years. Thus, there 

is increased concern about the potential evolution of hydrilla strains that are resistant to 

Aquathol®, the contact herbicide most commonly used to control hydrilla in springs. 

• Establishment of selective biological control organisms is increasingly viewed as a 

priority for sustainable control 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 suitable for experimental introduction into other 

springs systems. 

The Literature: What is Not Known

• Although scientific literature indicates clear relationships between nitrogen 

enrichment and increased growth of water hyacinth, water lettuce, and hydrilla in nonflowing aquatic systems, the few studies available for springs and other flowing waters 

have 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 would be necessary to develop nitrate-nitrogen limitation values for these 

plants in springs conditions. 
xii

• Although the scientific literature suggests 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 the environmental and economic importance of 

artesian springs in Florida. An ecosystem approach is essential to provide a greater 

understanding of the relative interactions between the myriad physical, chemical, and biological 

fluxes present in springs and their normal responses to rising nutrient levels. A focused and 

logical research agenda will be critical to pulling these precious but threatened natural resources 

back from their current declining path. 

The Literature: What is Known

• The best evidence suggesting a decline in the 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 Springs ecosystem may be considerably 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 submersed aquatic plant biomass. 

o Biomass estimates for the epiphyte community in the summer were approximately 

three-fold higher than those reported by Odum, while winter values were not 

significantly different between the two studies. 

o 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. 

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 Knight’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 Rock Springs Run (WSI 2007a) (spring runs 

averaging 0.69 and 0.84 mg/L of nitrate nitrogen, respectively)found similar results as 

those observed at Silver Springs with significantly lower gross primary productivity and 

photosynthetic efficiency demonstrating an inverse correlation between rising nitrate 

concentrations and reduced aquatic ecosystem metabolism. 

The Literature: What is Not Known, with Emphasis 

on Future Research Directions

Despite a wide variety of springs’ research, there remains a significant knowledge gap between 

the real and perceived threats that nitrate pollution plays on the ecology of spring ecosystems. In 

summary the following recommended applied research activities are suggested: 

• Baseline data on ecosystem-level structure 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 range of forcing functions including discharge 

rates, groundwater chemistry, nutrient concentrations, and recreational 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 responses to a range of nutrient conditions under 

realistic and relevant spring environmental conditions; 

• Understanding of impacts and effects of management interventions are lacking.

Whole-spring manipulation studies are suggested to test the effects of possible 

management techniques such as controlling levels of springshed nutrient loading, human 

recreational activities, alternative control methods for invasive species, and estimation of 

optimal consumer carrying capacities (e.g., manatee density). 

Chapter 7: Managing Nutrient Inputs to Florida Springs: The Legal 

Framework

The legal framework for managing the nutrient pollution of Florida springs is potentially as 

broad as the scope of human activities that contribute 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 set 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 the National Pollutant Discharge Elimination 

System (NPDES). Discharges to groundwater may be regulated under this program if 

there is a “significant nexus” to the quality of navigable surface waters. 

• The Safe Drinking Water Act (SDWA) protects public water supplies by establishing 

minimum criteria for drinking water quality and requiring states to regulate the 

underground injection of pollutants. Groundwater quality standards 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 implements 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 

direct jurisdiction of the federal program, such as nonpoint sources, agricultural 

discharges, and additional discharges 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. Domestic wastewater treatment plants are 

generally only required to use secondary treatment, except in areas where the 

Legislature has required advanced wastewater treatment (AWT). Concentrated 

Animal Feeding Operations (CAFOs) 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. 

o 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. 

• Waters designated as Outstanding Florida Waters (OFW) have a general limitation on 

their degradation. A narrative nutrient standard prohibits altering nutrient concentrations 

“so as to cause an imbalance in natural populations of aquatic flora or fauna.” Numeric 

standards can also be developed, most commonly as Total Maximum Daily Loads 

(TMDL). 

• The TMDLs process begins with determining whether a particular waterbody is 

“impaired” and demonstrating it does not meet water quality standards for a specific 

criterion. If there is sufficient data demonstrating 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 equitably allocated to the various sources and 

basins contributing pollutants. A Basin Management Action Plan (BMAP) is developed 

for implementing the adopted TMDL. 
xv

• Stormwater is a significant source of pollutants 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 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 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 septic tanks and drainfields. 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 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 nutrients to springs. They can provide for 

improved stormwater and wastewater management. 

• The Florida Legislature may require special protections or enact other legislation that 

could protect springs and springsheds. The Legislature has specifically required local 

governments in the Wekiva Study Area to amend their comprehensive plans to enhance 

springs protection. It has also limited the ability of local governments to regulate 

agricultural land uses and the conversion of agricultural land to urban uses. A Consumer 

Fertilizer Task Force has recommended new restrictions on the ability of local 

governments to regulate the application of fertilizers.

Chapter 8: Synthesis of Nutrients and Springs: Knowns, Unknowns, Research 

Priorities and Management Implications 

This chapter represents a synthesis of the most important findings; greater details can be found 

within each of the chapters.. This synthesis chapter is organized into three parts: 

1) a summary of the “knowns” regarding nutrient effects on springs and spring biota, 

2) the unknowns and suggestions for future research, and 

3) suggestions for a research and adaptive management program for spring ecosystems in 

Florida. 

Part 1: NUTRIENTS and SPRINGS: State of the knowledge

There is growing evidence that nitrate concentrations are increasing in the water discharging 

from springs. Given the importance of nitrate-N and its role as a driver of primary production in 
xvi

many aquatic systems, N appears to be the aspect of springs that has received the most study. 

Chapter 8 summarizes the current state of knowledge concerning nutrients in springs. 

PART 2: SPRINGS RESEARCH PRIORITIES

Our review of the available 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 spring and the effects of 

nutrients and other stressors on spring biota. 

Overall we propose a rather holistic research agenda that includes diagnostic studies, paired 

comparative studies of springs, manipulation studies, and fine grained monitoring to elucidate 

cause and effect relationships and to test various hypotheses regarding the impacts of stressors on 

spring ecosystems and biota. In particular we suggest: 

• that a number of smaller springs be selected that represent a variety of stressors acting 

separately and in combination as a means of testing various hypotheses regarding the 

impacts of stressors on spring ecosystems and biota. Tests might involve, for instance, 

choosing springs with low versus high DO, presence 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 like number of springs that is designed to 

elucidate spatial variability 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 contribution 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 between 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 multitude 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 frameworks. 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, and then initiating a round of management 

interventions to test hypotheses concerning the causes of current apparent declines in 

spring ecosystem condition. 

• The social framework of the ASAP team should include a tripartite of individuals 

from government, science, and the citizenry at large.

• Research within the AM context is designed 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 large spatial and temporal scales. 

o AM research crosses many ecological and organizational boundaries increasing 

its complexity and the lead time required 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-economic constraints. 
xviii

TABLE OF CONTENTS 

Chapter 1 - Springs as Ecosystems.................................................................................... 1 

Summary................................................................................................................................. 1 

INTRODUCTION TO FLORIDA 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 ..................................................................................................... 34 
xix

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 Loading, 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 QUANTITY AND QUALITY ............................................................ 77 

Water Quantity...................................................................................................................... 77 

Nitrogen ................................................................................................................................ 79 

Phosphorus............................................................................................................................ 85 

Carbon................................................................................................................................... 87 

Other Water Quality Parameters........................................................................................... 89 
xx

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 TRANSFORMATIONS 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 and Implications for Nutrient Cycling

......................................................................................................................................... 135 

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

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

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

INTRODUCTION .................................................................................................................. 137 

SPRING ENVIRONMENT.................................................................................................... 137 

Spring water chemistry ....................................................................................................... 137 
xxi

Reactive Zones.................................................................................................................... 139 

Biofilm ................................................................................................................................ 140 

Benthic Algal Mats ......................................................................................................... 141 

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

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

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

Carbon Sources ............................................................................................................... 144 

Decomposition .................................................................................................................... 145 

Nitrogen .............................................................................................................................. 148 

Biological Nitrogen Fixation .......................................................................................... 149 

Nitrogen uptake............................................................................................................... 150 

Nitrogen mineralization .................................................................................................. 151 

Denitrification................................................................................................................. 152 

Alternative N pathways................................................................................................... 153 

Overall N Cycle .............................................................................................................. 156 

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............................................................................................................... 162 

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

Sulfide Oxidation............................................................................................................ 162 

Sulfate Reduction............................................................................................................ 163 

Overall S Cycle............................................................................................................... 164 

EFFECTS OF INCREASED NUTRIENTS ........................................................................... 164 

Carbon/Organic Matter Impacts ......................................................................................... 164 

Nitrogen Impacts................................................................................................................. 164 

Phosphorus Impacts ............................................................................................................ 165 

Sulfur Impacts..................................................................................................................... 166 

SUMMARY AND CONCLUSIONS ..................................................................................... 166 
xxii

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 .................................................................................................................. 182 

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 Nutrients and Fauna in Florida Springs.................... 207 

Nutrient Effects Operating Through Ecosystem Processes ................................................ 207 

Nutrient Effects Operating Through Toxicity..................................................................... 210 
xxiii

Toxicity of Ammonia...................................................................................................... 211 

Toxicity of Nitrate........................................................................................................... 213 

Summary of Nutrient Effects 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 .................................................................................................................. 234 

Major Invasive Aquatic Plants: HIstory and Control in Florida............................................. 235 

Water Hyacinth ................................................................................................................... 235 

Water Lettuce...................................................................................................................... 239 

Hydrilla ............................................................................................................................... 240 

NITRATE-NITROGEN AND NONNATIVE 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....................................................................................................... 247 

AQUATIC PLANT CONTROL AND ALGAL SUCCESSION ........................................... 248 

Herbicide Selection Hypothesis.......................................................................................... 248 

Copper............................................................................................................................. 249 

Diquat.............................................................................................................................. 249 

Endothall......................................................................................................................... 250 

Glyphosate ...................................................................................................................... 250 

2,4-D ............................................................................................................................... 251 

Attractor – Catastrophe Hypothesis.................................................................................... 251 
xxiv

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............................................................................................................. 280 

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.............................................................................................................. 289 
xxv

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 Inputs to Florida Springs: The Legal 

Framework................................................................................................................. 305 

Summary............................................................................................................................. 305 

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

Clean Water Act.............................................................................................................. 307 

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............................................................................................................. 316 

The Verified List............................................................................................................. 317 

Calculation and Allocation ............................................................................................. 317 

Basin Management Action Plans (BMAP)..................................................................... 318 

STORMWATER..................................................................................................................... 319 

NPDES Stormwater Permitting ...................................................................................... 321 

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

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

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

Research Priorities and Management Implications.................................... 327 

Summary............................................................................................................................. 327 

INTRODUCTION .................................................................................................................. 329 

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 Invasive 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 SPRINGS POLICY and MANAGEMENT ............. 341 

The Springs Initiative: Protecting 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 ....................................................................................................................... 349 
1

1 

Springs as Ecosystems 

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

1 Wetland Solutions, Inc. 

2 Department of Environmental Engineering 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 temperature and chemical conditions. Structurally, springs 

are best characterized by their 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 producers 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 exerting an increasing variety of 

stresses on spring ecosystems in Florida. These external and internal stresses include decreasing 

groundwater supplies and increasing levels of nutrients, extensive recreational disturbance, and 

management alterations. Due to their physical and biological complexity combined with their 

adaptations to provide optimal levels of system productivity, a holistic ecosystem-level approach 

to spring studies is necessary to provide a greater understanding of anthropogenic effects (such 

as nutrient enrichment) on spring ecosystems. Development of a management strategy for 

springs will be dependent upon a better understanding of the relative importance and synergism 

of direct stressors at the local level and indirect stresses that occur at the regional scale. Systemslevel metrics have promise for providing a better 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 availability of macro and micro nutrients and by the 

ambient groundwater temperature. Additional factors that may affect 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 period 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 archeological evidence, and by present day populations who utilize them 

for a variety of recreational purposes (Scott et al. 2002). There is little doubt that the intrinsic 

aesthetics of clear, cool water vigorously emanating from the underground will continue to 

interest and fascinate humans. Many of Florida’s oldest tourist attractions (e.g., Weeki Wachee 

mermaids, Homosassa fish bowl, and Silver Springs glass bottom boats – see Figure 1) utilize 

spring ecosystems as their main draw. In addition, numerous spring boil areas have been 

modified to facilitate swimming, recreation, and even “health spas”. Today all of the largest 

springs in Florida, whether privately or publicly 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 erosion, litter, etc.) due to visitation by 

ever increasing numbers of people. 

However, over the last two decades other more serious factors, with the potential to permanently 

alter Florida’s spring ecosystems, have been increasingly recognized. Two of the most apparent 

anthropogenic factors that may be causing significant changes in spring ecosystems are: 1) the 

reduction in groundwater supply through consumptive human withdrawals; and 2) the 

simultaneous pollution of groundwater, principally with nitrate-nitrogen, resulting from human 

land use changes, cultural practices, and general population growth. The full spectrum of the 

combined effects associated with flow reductions and nutrient enrichment on spring ecosystems 

is not yet clear. However, there is justifiable concern for potential negative consequences due to 

the obvious degradation of other aquatic habitats worldwide subjected to declining flows and 

water levels, and increased nutrient loading. The potential consequences of nutrient enrichment 

in springs include an increase in opportunistic primary producers and organic matter deposition, 

an increase in nuisance algae species biomass, a decrease in plant and animal productivity and 

diversity, a reduction in water quality and not insignificantly, a reduction of the aesthetics these 

ecosystems have long provided. Clearly, an effort to synthesize the current state of knowledge 

concerning the effects of nutrient 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 Silver Springs at the time of H.T. Odum’s 1957 

landmark study (illustration by Elizabeth 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 environmental characteristics generate natural 

controlled settings suitable for general ecological research. Springs research is broad and varied, 

ranging from descriptive classifications 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 methodologies. The first, the 

“reductionist” method, in which single species and their associated ecological controls are 

studied, in turn may be combined with other species-centric studies to assemble a working 

community model. This approach is generally limited 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 second 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 obtained concerning the fate of 
4

individual species and resources that 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 that our understanding of 

spring ecology and how to apply effective management and protection to spring ecosystems 

should be based on both types 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 other aquatic ecosystems. 

This environmental stability, in turn, results in less temporal complexity in springs than in other 

aquatic ecosystems and more consistent food webs based on relatively 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 condition, 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 resulting effects on groundwater quantity 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 “canary 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 ecosystems 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 there 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 metabolism, temperature, and major chemical 

constituents are “essentially constant” (Odum et al. 1953). The steady-state concept of springs 

was further expanded by Whitford (1956) in his observations that discharge, temperature, gas, 

and mineral content of spring waters were all relatively constant within individual spring 

ecosystems (although quite different between springs) and that only light varied greatly as a 

consequence of temperate latitudes. 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 steady-state systems. The stability of spring 

ecosystems must also be viewed in context of larger 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 before present) when sea levels were dramatically lower. Yet, 

there are no other ecosystems outside of the tropics and the ocean that have had such a long 

period of relatively constant external forcing functions and resulting community adaptation. 
5

As dynamic entities composed of physical, chemical, and biological components and fueled by a 

variety of external energy flows, springs and spring runs have all the qualifications 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 individual water quality parameters or plant and 

animal species lists. When available, additional information from springs and spring run 

ecosystems from other localities in the United States and elsewhere is utilized to emphasize, 

enhance, or contrast information from Florida springs throughout 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 (decimal 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 coastal areas to salt water influences. As global and 

national economies influence land use changes in this region of Florida, springs are affected. For 

example, agricultural products in the Suwannee River area of Florida 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 through their irrigation requirements and increased 

loads of nitrogen and other pollutants 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 are economic engines, 

attracting a high rate of international and interstate visitation (5% and 31.5% of visitors, 

respectively, Bonn 2004). These visitors spend money that fuels local economies that have the 

potential to affect the springs’ ecologies through mechanisms such as groundwater pumping, 

increased surface water runoff pollutant transport, 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 regional rainfall. The quantity and 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, springs are dependant upon their “springshed”, 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 springshed delineations of many of Florida’s largest spring ecosystems. 

These springshed delineations are held by the Florida Geological Survey (FGS) and were 

produced using multiple methods using many data sets including surface water and ground-water 

flows, potentiometric levels, subsurface conduit maps and dye trace studies and many other 

hydrogeologic information (FGS 2007). The inset portion of Figure 2 shows an image of the 

Ichetucknee Springs springshed located in north central Florida. A large number of 

anthropogenic activities within this springshed have the potential to affect the ecology of these 

springs. 

Figure 2 – Florida springshed delineations, with an inset 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 groundwater within a Florida artesian spring at the 
7

regional scale. The diagram includes the major forcing functions of sunlight, atmospheric inputs 

(wind, rain, and storms), with associated water 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 energy, with counter flows of money for those 

pathways mediated by human activities. Variables in the diagram include: natural lands and low 

intensity land uses that do not receive significant inputs of nitrogen from anthropogenic uses; 

developed lands that include all other land uses; the water and nitrogen associated with the 

natural and developed lands; the artesian groundwater within the basin; and the physical and 

biological structure of a spring. Key system-level exports in the diagram include 

evapotranspiration, outflows of goods (recreational services), and downstream exports of water 

and organic material. 

Figure 3a – Energy symbols in the “Energese” model language, from Odum et al. (1998). 
Figure 3b – Conceptual springs regional diagram (M.T. Brown 2008). Key forcing function at the regional scal

wind, precipitation, and associated nutrients, hurricanes, and human goods and services. Springsheds are in turn modified by their associated undeveloped and developed terrestrial environments with fertilizer and waste nutriecumulative synergistic impacts resulting from consumptive water uses, recreation, and various resource manage
9 

Documentation of the effects of regional land use 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. This analysis found that approximately 75% of the 

water discharging from Silver Springs originates within a two-year 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 predominantly natural landscape to a mostly 

urban/agricultural area. This change corresponds to increased nitrogen loading, principally as a 

result of increased groundwater nitrate concentrations. Figure 4 provides another example of the 

observed correlation between increases in human population and increases in nitrate 

concentration of the ground water discharging from Weeki Wachee Spring in Hernando County. 

Figure 4 – Population growth (diamonds) in Hernando County, Florida and nitrate nitrogen 

concentration (triangles) at Weeki Wachee Spring from 1923 to 2006 (from DEP 2006). 

Spring Scale 

Springs in Florida occur where 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 incised 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, artesian 
10 

groundwater inputs, biological exchanges, and human activities. In general, a spring ecosystem 

includes all the internal abiotic and biotic components typical of most aquatic ecosystems 

including: aquatic macrophytes (higher plants) that support a diverse assemblage of attached 

algae (periphyton, benthic algae, and pseudo-plankton), detritus (dead plant and animal material 

generally associated with benthic organic sediments), associated animals feeding on detritus, and 

a faunal food web of herbivores, omnivores, and carnivores. Because of their continuous 

discharge, springs commonly serve as headwaters for streams, rivers, and sometimes lakes and 

ultimately contribute a great deal of freshwater to downstream inland and coastal aquatic 

ecosystems. 

Figure 5 provides a systems diagram illustrating some of the immense complexity inherent in a 

relatively unaltered Florida artesian spring ecosystem. This model provides a summary of the 

major external forcing functions, internal energy storages and processes, and interactions 

between the spring ecology and the human economy. Functional groups within 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 aggregation illustrated forcing functions 

include: sunlight, atmospheric inputs of water and nutrients, groundwater inputs of water and 

nutrients, sediment derived inputs of nutrients, and anthropogenic inputs such as management 

actions, goods, services, and people. Aggregated outflows in the diagram include 

evapotranspiration, surface water discharges with 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. 
Figure 5 – Detailed conceptual diagram of a Florida artesian spring ecosystem (M.T. Brown 2008). 
Figure 6 – Aggregated spring ecosystem diagram that incorporates human recreation and management activitie
13 

ENVIRONMENTAL FORCING FUNCTIONS 

Solar Inputs 

Solar inputs comprise the largest portion of the total external energy delivered to natural 

communities, and, in turn, are converted by photosynthesis into organic matter and other forms 

of potential energy, with the entropy of heat losses equally balancing inputs. The conceptual 

model of energy flow through a natural community, described and quantified by Odum (1956, 

1957a, 1957b) in his work at Silver Springs, Florida, 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 conversion of solar energy to useful 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 30.8 and 27 degrees north 

latitude, and are subject to seasonal variation in incident light due to the orientation of the sun. 

Maximal solar inputs occur during the spring and summer and minimal solar inputs occur during 

winter, with corresponding patterns in daylight length. Solar inputs occur across 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 active radiation (PAR), is typically measured with a quantum sensor and 

expressed in units of μmoles of photons per square 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 Silver 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

1/30/043/20/04

5/9/04

6/28/04

8/17/04

10/6/04

11/25/04

1/14/05

3/5/05

PAR (uE/s/m2

0

50

100

150

200

250

300

350

1/30/04

3/20/04

5/9/04

6/28/04

8/17/04

10/6/04

11/25/04

1/14/05

3/5/05

Total Insolation (J/s/m2)

Figure 8 – Annual cycle of daily insolation and photosynthetically active radiation data 

recorded at Silver Springs, Marion County, Florida (from Munch et al. 2006). 

Atmospheric Forcing Functions 

Broadly, atmospheric inputs include 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 precipitation (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 receive between 52 and 60 inches of rainfall on an 

annual average basis (Miller 1997). The majority (~ 80%) of this 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 nutrient inputs directly on their water surface area and indirectly 

through their groundwater inputs and surface water 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 the 
16 

majority of their water and nutrient inputs from groundwater; direct atmospheric inputs are 

generally negligible. 

Atmospheric nitrogen can be delivered to the environment as inorganic forms dissolved in 

rainwater or as particulate organic and/or mineralized forms. Atmospheric nitrogen exists 

primarily in the form of inert nitrogen gas (N2), but inorganic nitrogen reaches the atmosphere 

through volatilized ammonia from industrial emissions, liquid ammonia fertilizers, and organic 

wastes. Emissions of oxides of nitrogen (NOX) to the atmosphere occur 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) atmospheric deposition are obtained 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 sinks for nutrients in springsheds are described in Chapter 2, 

and as demonstrated, the atmospheric deposition loading 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 their subsurface water source, springs frequently 

have low concentrations (and percent 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 supplied by primary production. Carbon dioxide is an 

essential nutrient for plant growth in springs 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 groundwater source (Duarte et al. 2007). 

Other important atmospheric forcing functions include the effect of wind on mixing of waters 

and the transport of seeds and reproductive propagules. 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 colonize spring heads, which are otherwise 

inaccessible to floating propagules. Atmospheric forcing functions may also be broadly 

interpreted to include invertebrate and vertebrate air borne inputs. The adult forms of aquatic 

invertebrates commonly occurring in springs (especially midges, stoneflies, and aquatic moths) 

depend upon flying skills to propagate in and re-colonize head spring areas. A large variety of 

aquatic and wetland dependent birds 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 hydrological, elemental, and mineral forcing functions for the 

artesian spring ecosystems of Florida. The groundwater supplies of Florida are derived from 
17 

precipitation, primarily within the state’s boundaries, although some groundwater is derived 

from Alabama and Georgia precipitation (Fernald and Purdum 1998). 

Florida has five principle aquifer systems. From largest to smallest they are: the Floridan, 

Biscayne, Surficial, Intermediate, and the Sand and Gravel aquifers (Miller 1997), each of which 

has corresponding importance as a source of freshwater for human usage. The largest aquifer in 

Florida and the one that is of primary importance to artesian springs is the Floridan (Figure 9). 

Key points of importance related to an understanding 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 practical purposes) of human control; 

• artesian aquifers are the primary source 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 chemical characteristics are largely determined by 

the natural and anthropogenic uses 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 Fernald 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 forcing function for spring discharge); the second is 

that cultural land use activities may result in increased quantities of nutrients and other chemical 

constituents in water delivered to the aquifers. 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 withdrawals by humans and spring discharge is 

illustrated by long-term data for Rock and Wekiwa Springs groups near Orlando (WSI 2007a). 

Figure 10 depicts measured and modeled discharge 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 combination of increased consumptive use (withdrawals) and 

declining rainfall totals over the corresponding time period (Intera 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 has 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 nitrogen is one of the essential 

elements for plant life. However, nitrate is readily utilized by aquatic plants and algae, and in 

excess can lead to the development of nuisance 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 nitrogen loads and loads appearing at the spring 

boil show fairly good agreement. Other studies have “fingerprinted” specific nitrogen sources 

(animal waste, inorganic fertilizer, septic tank, etc.) based on ratios of stable nitrogen isotopes 

appearing in the spring vents (Phelps 2004; Toth 2003). 
19 

y = -0.3124x + 688.31

R2

 = 0.4654

50

55

60

65

70

75

80

85

90

95

100

1950 1960 1970 1980 1990 2000 2010

Year

Annual 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 established 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 observed in springs contained 

in undeveloped springsheds (Champion 2001, Munch et al. 2006). This correlation between 

developed land use in a springshed 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 portions of nitrate nitrogen found in 

the Upper Floridan Aquifer and in spring discharges 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 nitrogen 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 spring 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 Springs (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 the contributory nitrogen sources for two spring 

systems, Wakulla Springs south of Tallahassee and Wekiwa Springs near Orlando. In the 

Wakulla springshed, human-derived wastewater sources (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 forest land and replace it with pasture and grass 

covered land uses (MACTEC 2007). For the Wekiwa springshed, residential fertilizer 

applications (20%), agricultural fertilizer applications (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 via atmospheric deposition and groundwater supply. 

Indirect anthropogenic impacts on springs may also result from changing downstream conditions 

through construction of dams and weirs, stream channelization and sedimentation, and 

harvesting or otherwise manipulating fish and other wildlife populations that normally include 

the spring and spring run in their home range. 
21 

Figure 11 – Top pie chart illustrates the relative contribution of total 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 total nitrogen load to the springshed of Wekiwa 

Springs for 2004 (MACTEC 2007). 

Direct anthropogenic inputs or forcing functions to Florida’s artesian springs include: 

recreational activities that trample or dislodge sediments and submersed aquatic vegetation; 

aquatic plant management activities (principally in the form of physical, chemical, or biological 
22 

control of non-indigenous or “undesirable” plant and animal species); modification of spring 

habitats by construction that reduces connectivity (bulkheads, shoreline modifications, damns, or 

weirs); and general modifications to food webs by introduction of non-indigenous species and 

reductions in top level consumers through hunting, fishing, and habitat alteration. 

Nutrient loads to the groundwater represent an indirect anthropogenic forcing function at the 

regional level while nutrient loads delivered via stormwater runoff to springs 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. Springs have been classified based upon geomorphic 

features, and can be further subdivided as springs, sinks, river rises, or karst windows (Wilson 

and Skiles 1989). As karst features, spring pools (or vents) may initially form through 

catastrophic collapse, but their overall genesis may require millennia. Exceptions include springs 

discharging from karst fissures which underlie surface 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 basins approaching 90 m (~ 300 ft) 

in diameter. The bathymetric profiles of spring basins are variable, but a typical spring basin has 

gradually sloping sides which descend to the vent/s like a funnel. 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 underwater structure of 

many of Florida’s largest spring systems have been explored and mapped, none more extensively 

than Wakulla Springs (Figure 12). 

Spring runs begin as a direct extension of a spring basin, drain according 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 their base-flow from groundwater sources (Scott et al.

2004). In the karst-rich portion of Florida, many streams receive substantial base flow from 

springs. The size of spring runs corresponds to the volume of water they discharge, ranging from 

a few meters to 100 m (~330 ft) in width, while lengths 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 Wakulla Cave and Spring System in Wakulla 

County, Florida (from Loper et al. 2005). 

Spring flows are created by a difference in potentiometric head between the groundwater and the 

receiving water. A portion of the potential 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 spring run corresponds to the quantity of water it 

discharges, water residence times are typically short, in the range of hours, and preclude the 

development of measurable plankton abundance (Hynes 1970). 

Spring flow can expel floating macrophytes 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 otherwise 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 growth, 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 spring 
24 

flow is reduced by natural or anthropogenic factors, 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 require considerable time if spring flows are 

restored. 

High water clarity often facilitates littoral zones that include the entire spring and spring run, 

although other factors including human recreation, herbivore grazing, stream velocity, 

sedimentation, and riparian shading commonly influence the distribution and abundance of plant 

life (see Chapter 4). 

Water and Dissolved Constituents 

Arguably the most important structural component 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 dissolved and particulate forms. This is beneficial to many forms 

of life, both autotrophic and heterotrophic, and provides the basis for the observation that some 

springs have been found to be among the most productive ecosystems on the earth (Odum 1971). 

Water discharged from artesian springs associated with the Floridan Aquifer typically has 

slightly basic pH, is commonly under-saturated in dissolved oxygen and over-saturated with 

dissolved carbon dioxide, and contains a variety of elements in trace amounts. Spring water is 

typically low in inorganic particulate matter as well 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 deposited by anthropogenic activities on surface lands 

and waters can be transported to groundwater and subsequently discharged to spring ecosystems. 

The first and most commonly utilized classification for Florida springs is based on magnitude of 

discharge as initially described by Meinzer (1927), subsequently 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 ft3 = 0.0283 m3; 1 gal = 3.785 liter; 1 pint = 0.473 liter 
25 

Because the groundwater discharged from springs is qualitatively influenced by surrounding 

geology, soils, and land surface activities, the resulting 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 cluster analysis of the six predominant ion 

(calcium, bicarbonate, magnesium, sulfate, sodium, and chloride) concentrations for 170 springs, 

Woodruff (1993) arrived at four distinct spring types and their 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 most predominant ions (calcium, bicarbonate, 

magnesium, sulfate, sodium, and chloride) from 170 Florida springs which were utilized to 

create four distinct springs groupings via cluster analysis (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 dominated by photoautotrophic life forms which use 

energy derived from sunlight to fix organic matter from inorganic compounds. 

Chemoautotrophic bacteria are also capable of generating organic material from inorganic 

sources, and may exist in anaerobic springs, but their ecology or relative importance is unknown. 

Additional sources of organic chemical energy to spring ecosystems 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 considered in regard to the trophic 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 Florida’s springs are the macrophytes, a grouping that includes vascular 

plants, bryophytes, and macroscopic and colonial algae. The vascular aquatic plants have 

internal transport structures (xylem and tracheids), are principally angiosperms (flowering plants 

with enclosed seeds) evolved from land plants, and are comprised of relatively few species (< 

1%) in comparison to the diversity of terrestrial vascular plants (Sculthorpe 1985). Bryophytes 

lack xylem and true roots and are represented by mosses and liverworts (Stream Bryophyte 

Group 1999), with native Florida spring genera including Brachelyma and Fontinalis. 
26 

Aquatic macrophytes can be further subdivided by their growth habits as submersed, floating, or 

emergent. Common native aquatic vascular plant species include 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 ecosystems 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 vascular system or true roots, and are composed of 

individual cells that are independent and undifferentiated. Macroalgae that occur in springs and 

spring runs include the genera Chara and Nitella which have vascular plant 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 had macroalgae growing in them, with an 

average of 50% of the spring bottoms covered 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 observed in the spring survey. 

Collectively, macrophytes provide a physical structure upon which smaller primary producers 

can attach, thereby preventing their downstream export and facilitating access to a better light 

environment. These epiphytic communities consist 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 diatoms. This epiphytic community has high 

rates of primary production which can eclipse the productivity of macrophytes. 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 quarters 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 applied to shoot density data to 

obtain estimates of production (Hauxwell et al. 2007). Areal leaf production ranged between 3.6 

and 23.0 g/m2

/d of dry biomass, with highest productivity observed in late spring/summer, with 

an annual total leaf production estimate of 2.7 kg/m2; 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 clear-water spring ecosystems. 

The lotic nature of spring ecosystems would appear to preclude the formation of a self-sustaining 

phytoplankton community. However, stream and spring run waters can contain autotrophic 
27 

material made up of protists, algae, and cyanobacteria which have originated from the sloughing 

of benthic or epiphytic growths (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 total standing crop of chlorophyll on the stream bottom, as well as 

with the bottom surface area upstream of the sampling point. A similar pattern was observed in 

three Florida spring-fed streams, with an increase in chlorophyll 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 and benthic autotrophic 

communities, and, as a result, the species composition is likely to be similar. Although perhaps 

not a true phytoplankton community, this autotrophic material represents a steady downstream 

export of chlorophyll and primary production available to primary consumers. 

Primary Consumers 

Primary consumers in spring ecosystems are represented by herbivores and omnivores. Vascular 

aquatic plants are directly consumed 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 previous section, the highest rates of primary 

productivity are derived from the epiphytic community, which, in turn, supports a diversity of 

aquatic macroinvertebrate primary consumers. Invertebrate 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 trichopterans. Observations made by Munch et al. (2006) 

found chironomids, trichopterans, and lepidopterans 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 mussels which are found in Florida springs, but not 

at the diversity or abundance observed state wide, perhaps 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 commonly occurring was the non-indigenous Asiatic 

clam, Corbicula fluminea (Walsh and Williams 2003). In that study, a general increase in the 

abundance of mussels downstream of the spring vent 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 suspended material may limit the abundance and 

distribution of filter feeders in these ecosystems. 
28 

Table 4. The mean number (and std. dev.) of organisms collected from vegetated substrate in 

Mill Pond Spring, a feeder spring of the Ichetucknee River (from Woodruff 1993). 

Subclass / Order Number of Organisms 

(average #/m2

) 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 connection to coastal waters, are utilized by 

manatees as a place to feed and as a thermal refuge 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 status. Manatees have been 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-evolved with herbivores and should be able to recover from 

cropping by herbivores, when consumption by herbivores continually exceeds vegetation 

production rates, overgrazing and loss of vegetation can result. Non-indigenous snails may 

cause overgrazing, with possible destructive avenues being diverse feeding 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 submersed vascular plants 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 the abundance of submersed vascular plants and a 

shift from vascular plants to non-grazed filamentous algae has been reported as a result of 

manatee feeding at Kings Bay, a spring group which forms the headwaters of Crystal River in 

Citrus County, Florida (Hauxwell et al. 2004). 

Obligate herbivorous fish are rare in Florida springs as most species which consume plant 

material are also general omnivores and are likely ingesting a mix of detritus 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 sucker 
29 

mouth catfish (Hypostomus sp.), and blue tilapia (Oreochromis aurea), which have become 

common in several spring systems, including Rock, Wekiwa, Volusia 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 accidental and purposeful events in state waters 

beginning in the 1960’s. While non-indigenous fish can also be found in stream and lake 

systems, the thermal refuge provided within spring 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 observed 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 fish 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, Walsh 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 its connectivity with other drainages. Silver Springs, which 

appears to have the best historic fishery records available, has relatively stable fish species 

diversity, as evidenced by the following: Hubbs and Allen (1943) reported thirty-five species in 

the 1940s, Walsh and Williams collected twenty-nine species in 2002, while Munch et al. (2006) 

observed thirty-three species in 2004-05. Munch et al. (2006) pointed out that while the number 

of fish species utilizing Silver Springs was similar to historical observations, the loss of channel 

catfish and mullet, presumably as a consequence 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 allochthonous material 

forms an important reduced chemical energy pool available for 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 detritus 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 Florida artesian springs, Odum noted that the biomass of spring 

trophic levels (i.e., producers, primary consumers, secondary consumers, etc.) forms a classic 

“pyramid”, with lower biomass and numbers of living organisms at the higher trophic levels 

(Figure 13). This figure illustrates the predominance of strap-leaf sagittaria (S. lorata = S. 

kurziana) and associated periphytic algae as the biological flywheel of the Silver Springs 

ecosystem. Average standing stock (dry weight biomass) of these plants was estimated by Odum 

as about 809 g/m2

. Herbivores, including turtles, midges, snails, and the herbivorous fish mullet 

(Mugil cephalus) are the next step up on the biomass pyramid with an estimated average biomass 

of only 37 g/m2

. Higher trophic levels include primary 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/m2

 while the top consumers had an 

estimated biomass of only 1.5 g/m2

. 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 researchers 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 changes to estimate whole ecosystem metabolism 

in flowing aquatic ecosystems has been implemented 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 sondes that collect and store in situ measurements of dissolved oxygen and 

temperature, as well as quantum sensors to do the same for photosynthetically 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 dissolved oxygen concentration, percent saturation, 

and temperature conditions. Diffusion is estimated from knowledge of percent saturation and 

average flow velocity. Oxygen mass transport and rate-of-change are calculated from 

continuous flow records, hydraulic residence time, and dissolved oxygen concentration changes. 

Five components of community metabolism can be estimated using this methodology: 

• Gross primary productivity (GPP), the total primary productivity of the entire submersed 

photosynthetic community, including all algae and macrophytes; 

• Net primary productivity (NPP), the net difference between gross primary production and 

community respiration equal to the organic (fixed) carbon available at the base of the aquatic 

food web; 

• Community respiration (CR), the dark respiration or oxygen consumption of the entire 

aquatic community, including microbial populations, algae, macrophytes and fauna; 

• P/R ratio, ratio between gross primary production and community respiration, indicating 

autotrophic vs. heterotrophic status of the aquatic ecosystem; 

• Photosynthetic (or GPP) efficiency, the conversion efficiency of useable PAR to gross 

primary production. 

Primary Productivity 

Primary productivity can be defined as the rate of formation of organic material from inorganic 

carbon by photosynthetic activity and through the photosynthetic generation of oxygen, can be 

measured as a rate value representing the conversion of solar energy to reduced chemical energy 

(Bott 2007). In the process of creating this organic material, some energy is lost through 

respiration (Ra), 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 + Ra (Eq. 1) 
32 

Gross primary productivity (GPP) is the total primary productivity of the entire submersed plant 

community, including all algae and macrophytes, and typically expressed in terms of the amount 

of oxygen produced per unit area and unit time, (i.e., g O2/m2/d). Net primary productivity 

(NPP) is the net difference between gross primary production and community respiration and is 

equal to the organic carbon available at the base of the aquatic food web. Daily dissolved 

oxygen rate of change measurements integrate a number of ecological functions and are readily 

convertible to units of energy (Reeder and Binion 2001). 

While there are primary producer bacteria which generate organic mater without solar inputs (i.e.

chemoautotrophs), they do not generate oxygen and their importance in Florida spring 

ecosystems is unknown but believed to be minor. Therefore the measurement of primary 

production in spring ecosystems is based on the equation describing photosynthesis: 

6 CO2 + 12 H2O → 6 O2 + C6H12O6 + 6 H2O (Eq. 2) 

The rate of photosynthesis can be measured via changes in dissolved oxygen (or alternatively 

carbon dioxide) typically in units of mass per area per time, such as g O2/m2

/d. The utilization of 

dissolved oxygen changes to measure community metabolism in streams was pioneered by 

Odum (1956) in his study of Florida springs. It was based on the theoretical considerations that 

oxygen is introduced into the water as a result of photosynthetic primary production during the 

daylight hours; oxygen is consumed by living organisms (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 through accrual of groundwater and surface drainage 

inputs (see Figure 14). Therefore, changes in dissolved oxygen concentrations (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)
0123456789

10

5/15/05 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

-0.5

0

0.5

1

1.5

5/15/05 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

-1.0-0.8-0.6-0.4-0.2

0.00.20.40.60.81.01.2

5/15/05 0:00 5/15/05 6:00 5/15/05 12:00 

Corrected DO Rate-of-Change (g/m2/hr)

0

500

1000150020002500

5/15/05 0:00 5/15/05 6:00 5/15/05 12:00 

PAR (umol/m

2

/s)

Air = 39.3

Plant 

CR = 8.40 g/m2/d

NPP = GPP- CR = +0.05 g/m

PAR Eff. = GPP / PAR = 0.30 g O

Figure 14 – Example determination of ecosystem metabolism based on upstream-downstream dissolved oxygen dSprings (from WSI 2007b). 
 34 

Community Respiration 

Community respiration (CR) is the oxygen consumption of the entire aquatic community 

(ecosystem), including microbial, floral and faunal populations. Respiration is reported as 

community respiration because there is no practical way to differentiate between the metabolism 

of heterotrophs, such as microbes, invertebrates, and fish; and that of autotrophs. Community 

respiration occurs throughout day and night periods 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 following inputs of organic 

matter such as riparian litterfall or aquatic plant detritus following herbicide applications (WSI 

2006). 

Because respiration can not practically be differentiated between autotrophs or heterotrophs, it is 

collectively termed community respiration and is commonly expressed on a daily basis as CR24, 

and which adds some uncertainty in the estimation of GPP and NPP (see equation 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 biological activity, thus: 

NDM = GPP - CR24 (Eq. 4) 

Productivity/Respiration Ratio 

The ratio of gross primary production to community respiration (P/R) was introduced by Odum 

(1956) as a simple index of the relative dominance (or balance) of autotrophic 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) depending 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 heterotrophic 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 nitrogen and phosphorus, while recognizing that 

stream food webs can be based on autotrophic 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 efficiency considers the light energy that primary producers can 

utilize, commonly referred to as photosynthetically active radiation (PAR), which has wave 

lengths between 400 and 700 nm (Kirk 1994) and expressed in units of light (quanta, moles, or 

Einsteins) per unit area and time. As gross primary production is typically expressed in terms of 
 35 

the amount of oxygen produced per unit area and unit time, (i.e., g O2/m2

/d), the resulting 

photosynthetic efficiency term can be expressed as g O2/mol of PAR. 

Community Export 

Community export refers to downstream losses of organic carbon from the spring system. Plant 

matter (algae and macrophyte biomass) is typically 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 microscopic plant and algal materials. 

Pollutant (Nutrient) Assimilation 

Flowing aquatic ecosystems are highly productive (Odum 1956), and hence have tremendous 

potential to productively assimilate nutrients and other pollutants within definable concentration 

ranges. Effectively assimilated nutrients are either 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 nutrient 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 beneficial uses. Three different but interrelated 

measures of the whole ecosystem function for spring (and stream) ecosystems are: community 

metabolism; community export; and nutrient assimilation rates. When quantified, they can be 

compared to reference sites and have been found to be sensitive to perturbations 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 efficiency 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 the biomass produced by a higher trophic level; on 

average only 10 percent of energy is transferred with the balance lost through respiration and 

heat loss (Ricklefs 1990). 

Based primarily on his research at Silver Springs and other Florida springs, H.T. Odum (1983) 

hypothesized that community metabolism as measured 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 importance becomes increasingly clear as researchers 

attempt to explain the complex cause and effect relationships of multiple stressors in spring 
 36 

ecosystems. Knight (1980) tested this hypothesis at Silver Springs and found that the in situ

periphyton productivity of mesocosms 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 Florida springs, Odum (1957a) found that most were 

operating at similar ecological efficiencies, or in other words, their abilities to effectively utilize 

solar inputs to fix carbon (i.e., primary producer biomass) were approximately equal. Until 

proven otherwise, it may reasonably be assumed that spring ecosystems cannot consistently 

exceed these observed ecological efficiencies. This assumption is borne out by observation that 

the gross primary productivity values measured in relatively unaltered springs are at the same 

approximate level as gross primary productivity 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 Principle for assessing nutrient effects in spring 

ecosystems cannot be overstated. All significant biological activity in a typical spring ecosystem 

is dependent upon autochthonous productivity. Ultimately the number and secondary 

productivity of macroinvertebrates, amphibians, reptiles, fish, birds, 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, springs 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 flora and thereby lower the life-support capacity 

of the spring. Work by Knight (1980) confirmed that 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 consumers (snails) and secondary consumers (small fish) 

were altered from their normal, adapted levels. Chapter 6 pursues this line of reasoning further in 

evaluation of the proximal and ultimate effects of nutrient increases 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 attracted prey resources. Based on these 

attributes, some early human settlements were associated with springs. The use of springs for 

drinking water supply still continues, commonly 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 usage include: the total number of visitors, the 

number of visitors actually entering the springs and spring runs for water dependent activities 

(wading, tubing, kayaking, canoeing, swimming, scuba diving, or snorkeling), and number of 

visitors for other non-water dependent activities (picnicking, hiking, and bird watching). The 

human component includes the inputs of money spent to visit and utilize springs for recreation 

and aesthetic purposes. In one perspective, maintaining and protecting spring ecosystems from 

degradation ensures that monetary 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 macrophytes resulting from wading, power boat propellers, and boat 

anchors. Indirect human effects include changes in the quality and quantity of groundwater 

supplying spring ecosystems from land use changes and population increases around springs. 

Ironically, it may be the aesthetic attraction that springs hold for human beings which leads to 

their subsequent recreational over exploitation. 

Recreational Uses 

During the last century, Florida’s springs have been utilized for recreational activities and, 

consequently, many springs have been physically modified for this purpose. Physical 

modifications found at a number of springs include the following: seawalls, rip rap, dams and 

weirs, deposition of beach sand, boardwalks and docks, concrete-lined swimming pools, and 

large covered facilities for boats and people. Broadly speaking, springs attract users who will be 

engaged in two types of outdoor recreation, 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 aesthetics; 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 Orange County, Florida, found that there was an 

approximately even distribution between water contact and non-contact activities, as well as a 

demonstration that recreational activities vary by spring due to site-specific characteristics such 

as water depth and flow velocity. Unique recreational or tourism features exist at many springs 

as well; Silver Springs is noted for its glass bottomed boat tours, Homosassa Springs its 

underwater “fish bowl” viewing chamber, Ichetucknee Springs is a noted tubing destination, and 

Volusia Blue Springs is noted for winter time viewing of manatees. 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 Relax 40.22%

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 11th and 12th, 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 springs 
 39 

experienced peak use in spring (Apr-Jun) and summer periods (Jul-Sep) (Bonn 2004). 

Comparisons of human use at two springs with very 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 for these springs occurred from spring through 

summer for Ichetucknee and Wakulla, where in-water activities are the dominant forms of 

recreation and from winter through spring for Homosassa and Volusia Blue (which had a July 

peak also), reflecting the influence 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 activities 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 time periods for manatee observation. Patterns in 

the weekly usage were reported in a comprehensive study of Rainbow Springs river recreation 

(Holland and Cichra 1994), in which weekend use levels 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 during summer holiday weekends (Memorial 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 water 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 the park’s recreation area and boats to gain access to the spring, but all 

fishing was done from boats ranging in types and sizes from small runabouts less than 5 m to 

houseboats approaching 20 m in length, and many of the larger vessels stayed in the spring run 

for multiple days at a time. During summer holiday events, as many as 200 boats could be found 

in the Silver Glenn Spring run, an area of approximately 6.1 ha (15.2 ac). 
 40 

0

10,000

20,000

30,000

40,000

50,000

60,000

70,000

Jan-01Jan-15Jan-29

Feb-12Feb-26Mar-12Mar-26

Apr-09Apr-23

May-07May-21

Jun-04Jun-18

Jul-02Jul-16Jul-30

Aug-13Aug-27Sep-10Sep-24

Oct-08Oct-22Nov-05Nov-19

Calendar Day

Monthly 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)])

R2

 = 0.8372

Paid Attendance

Y = 12799 (1- 1.04 cos [(2 pi() / 365)(t+159.2)])

R2

 = 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 Day

Average 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 Springs 
 41 

Visitation numbers are influenced by the unique features of the spring, as well as other factors 

such as adequate parking and regulations (i.e., no alcoholic beverages allowed at state parks). 

The number of people utilizing a representative sample 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 four 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 annual visitor attendance increased from over 200,000 

to nearly 266,000 (33% increase), and Volusia Blue Spring State Park annual 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 ecosystems 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 Springs excluded) was approximately 1,092,468 persons, 

representing an average annual 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 springs combined, which would have skewed the 

calculated average visitation statistics. 

Excluding river rises and karst windows, Florida has 24 first magnitude springs/ spring groups, 

the majority of which provide 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 visitors 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 (2nd and 3rd magnitude) also attract numerous visitors, sometimes as many or more than 

larger springs. For example, Ponce de Leon Spring (Holmes County), attracted an estimated 

259,000 visitors in 2002, although visitation numbers for smaller springs are probably in the 

thousands to tens of thousands per year (Bonn 2004). 

The estimate used above of applying the average attendance figures from six first magnitude 

springs to total number of statewide first magnitude springs is simplified and certainly an 

approximation. As the detailed studies by Bonn and Bell (2003) and Bonn (2004) illustrate, 

there is much variation from spring to spring, both in visitation numbers (and economic impact) 

due to the unique characteristics of each spring and the recreational opportunities it presents. 

Nevertheless, this estimate of the potential number of visitors 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 impacts 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 Spring. Individually, the Ichetucknee Springs and 

Wakulla Springs had similar estimates of overall spending for the two parks (approximately $22 

million). Although Ichetucknee Springs attracts more visitors from outside the area, Wakulla 

Springs’ visitors spend more than those visiting Ichetucknee Springs; on average $89 a person 

per day is spent at Wakulla Springs and $34 at Ichetucknee Springs (2003 dollars). At the lower 

end of economic impact were Homosassa Springs and Volusia Blue Spring with total spending 

estimates of $13.6 million and $10 million respectively in 2002. These estimates were the result 

of these parks being heavily attended 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 annual direct spending by 

visitors accounted for over $65 million, generated 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% for shopping, 13.5% for lodging, 6.4% for evening 

entertainment, 4.3% for transportation, and 4.3% for all other spending. Silver Springs, which 

has a large theme park associated with the natural 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 impacts generated by Silver Springs has more to 

do with its development as a tourist site (with 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 applied to other springs. 

This may be especially true considering springs- focused parks operated by governmental 

agencies that do not have a profit-focused business plan. Nevertheless, 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 observed the majority of the spring run (6.1 ha) showed impacts 

from recreation, with direct or indirect mechanical 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 recreational use, the quantity of uprooted and cut drifting 

plants did not increase as recreation 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 to 
 43 

recreational activity, instead appearing 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) occurred 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. The 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 potential to cause harm. Personal 

observations indicate that people wading in shallow areas result in the near total absence of 

submersed vascular plants, while filamentous and attached macroalgae can remain and appear 

substantially more resistant to damage from trampling. 

Perhaps the most influential study to gauge the impact of recreational use (largely tubers) on the 

submersed aquatic vegetation and benthic habitats was conducted on the Ichetucknee River 

(Dutoit 1979). This study demonstrated a linkage between recreational activities and damage to 

the ecology of the spring run, primarily through physical disturbance in the form of trampling, 

uprooting, breakage, and generalized destruction of submersed and emergent aquatic vegetation. 

Recognizing that the aquatic vegetation 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 management changes to the number and frequency of 

recreational tubers using the river were implemented to minimize damage to the submersed 

aquatic vegetation. This entailed setting the daily number of tubers to 750 in the upper river and 

2,250 at the midpoint tube launch, as well a closing the shallow upper section from September 

through May following the peak summer use period; such limitations 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 outdoor recreation as a technique to 

manage the impact of human use on natural resources. Various facets of carrying capacity as 

applied to springs in public use areas include: the physical amount of space needed to 

accommodate recreation; visitors needs such as parking or restrooms; the 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 would be diminished (Billington 

1995). Thus recreational carrying capacity has two components, ecological and social, with 

ecological relating to environmental damage and social relating to the visitor (Shelby and 

Herblein 1986). 

Published case studies of the carrying capacity of Florida’s spring ecosystems have been 

conducted on the Rainbow and Ichetucknee Rivers and relied on the visitor surveys to 

characterize maximum usage levels. Generally, perceptions of increased crowding resulted in 

decreased satisfaction among users. According to Holland and Cichra (1994) visitors’ 

perceptions of crowding on the Rainbow River indicated 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 all” to “slightly” crowded and 2% reported 

“extremely” crowded. Perceptions of crowding by visitors were correlated with self-reported 
 44 

levels of satisfaction of their springs experience by Holland and Cichra (1994), with 52% of 

weekend users and 49% of weekday users rating their trip was "perfect” or “excellent”, and only 

5% of weekend users and 4% of weekday users rated 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 recreational 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 boaters. 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 among tubers. The relation between crowding and 

satisfaction suggests that the social carrying capacity for tubers may be approached during heavy 

use days, but overall, insufficient evidence exists to conclude that social carrying capacity has 

been exceeded for the majority of river users at the Rainbow River (Billington 1995). 

A final study examining the social carrying capacity was provided by Paulauskas’ (2001) 

research examining the experience of recreationalists using the Ichetucknee River in relation to 

perceptions of crowding. Results were derived from 320 respondent surveys conducted on site 

between May and June 2000. The author reported that if respondents encountered fewer than 

expected recreationists, they rated their experience as “highly satisfying” and that respondents 

with no expectations or who encountered as many as expected, also rated their experience as 

“highly satisfying”. However, perceptions of crowding did appear to influence overall 

satisfaction levels of river recreationists and the study revealed an inverse relationship with 

crowding on overall satisfaction ratings, in that as perceptions of crowding increased, satisfaction 

levels decreased (Paulauskas 2001). The results from this study indicate that despite heavy 

weekend use, when between 50% and 65% of the respondents rated the river as crowded, 

perceptions of social carrying capacity did not appear to be exceeded. 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 ecosystems 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 receive from local, regional, and 

even global systems in which they are embedded. Far too often, the study of these spring 

ecosystems has focused on select forcing functions or individual structural components with 

inadequate consideration of the whole. While these species or process-level observations 

contribute towards our knowledge of spring ecosystems, 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 trophic 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 comparison of multiple forcing functions, entire 
 45 

trophic level structure, and ecosystem-level processes such as gross primary productivity and 

community respiration provides a more holistic view of springs and holds promise for 

developing commonalities among the hundreds of different springs that occur in Florida. 

Chapter 6 of this report provides an in-depth review 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 must 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 quantification or estimation of all of the principal forcing functions 

and key details of spring structure and function should be considered in future spring studies. 
 46 

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 53 

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 springs biota and ecosystems of necessity requires an 

understanding of the manner and mechanism of nutrient delivery. Moreover, as evidence 

accumulates for particular consequences of nutrient 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 and hydrogeology of Florida’s karst systems, with 

an emphasis on intrinsic vulnerabilities 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 and across springs is informative. Attention is also set on the 

future of nutrient loading, including the potential 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 solutes 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 springshed management, but these physical attributes 

are poorly constrained in karst systems. The presence of large 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 pollutants from anthropogenic activities at the 

surface. The implications of this are profound for 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 layers (e.g., Hawthorn Formation) limit the rate 

of interaction between surface and Floridan Aquifer waters. Nitrogen (in nitrate form) is 

particularly problematic because it is effectively 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 documented. 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 concentrations in some springs; overall, spring 

nitrate concentrations appear to have increased 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 since the 1970’s, perhaps a signal of changing 

subsurface flowpaths that deliver increasingly old groundwater. Similarly, dissolved 

oxygen concentrations have declined since the 1970’s, in many cases crossing 

ecologically significant thresholds (~ 1.0 mg/L); this cause for this decline is unknown, 

but may also be related to the age of the water emerging at spring vents. Finally, regular 

observation of trace quantities of pharmaceuticals and other chemical by-products of 

modern society (e.g., DEET, caffeine) illustrate the degree of connectivity between the 

groundwater and nearby human activities. 

• The preponderance of N pollution appears 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 computations and isotopic tracer studies 

suggests that mineral fertilizers, and therefore not septic tanks and wastewater 

sprayfields, are the principal sources of N pollution. One key caveat discussed, is that the 

isotope method is relatively poorly constrained for subsurface flowpaths, which may lead 

to interpretive errors. More importantly, the source determination for each spring needs 

to be considered unique; 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 mixture of water discharging from spring vents. 

Given significant discord between observed dye tracer residence times (days to weeks) 

and nominal ages (decades), a research program to better understand water age over time 

seems critical. Decomposition, auto-correlation and cross-correlation analyses of 

resulting time series should enable improved understanding of flow regimes and also 

links with other water quality variables that are strongly relevant to springs ecology 

(nutrients, salts, dissolved oxygen). Specifically, if binary mixing models are useful 

simplifications of spring vent hydrology, time series will allow decomposition of flows 

into two components, with insight into the relative importance of each, and changes in 

mixing with regional climatic trends. These methods need to be applied cognizant that 

different springs are likely to have different hydrogeologic properties, and inference from 

one spring may not apply in others. This uncertainty has key implications on pollutant 

delivery in karst systems, which can be highly variable. Pulses of nutrients and other 

water quality parameters (dissolved oxygen, salts, toxins) have been invoked for 
 55 

catastrophic shifts in other systems, particularly 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 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 what are generally 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 systems, 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 appropriate 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 mineral fertilizer, there remains significant 

uncertainty about the interpretation of bulk stable isotope measurements in complex karst 

hydrologic systems. First, there 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 questions beyond nitrogen delivery, since there 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 and pollutant sources. 

• 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 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 salinity, 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. There 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 activities. For example, significant 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 concentrations in the 

Upper Floridan, but these remain critically understudied. Septic tanks, animal waste 

disposal systems and sprayfields generate enormous quantities of biological and 

nitrogenous oxygen demand if discharged unoxygenated into the groundwater. Since 

dissolved oxygen concentrations are strongly spatially variable, have significantly 

declined across a broad spectrum of springs (Scott et al. 2004), and were predictive of 

algal cover, understanding the roles of hydrogeologic and anthropogenic processes in low 

DO at spring vents is a first-order research unknown. 
 56 

• What is the assimilative capacity of the landscape and aquifer for nutrients and other 

contaminants? Efforts similar to the springshed nutrient budgeting for watersheds in the 

northeastern US yield an estimate of 20% for the fraction of terrestrial N load that is 

transported by rivers to the sea. In a karst environment where N movement is essentially 

unimpeded, conservative estimates of landscape 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), 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 research 

priority. Issues that will need 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, there may be loading thresholds that exceed 

the assimilative capacity of the landscape and lead to elevated load breakthrough. The 

potential for observing this effect is low 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 activities and declining water quality in spring 

ecosystems, evaluating the extent to which 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 loads, transformations and 

sinks in springsheds, emphasizing processes occurring 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 nutrient source tracing, temporal dynamics, spatial 

dynamics) and then transformations and sinks (abiotic attenuation, biotic attenuation, 

management-inducible attenuation). The chapter 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 effects of nutrients on the 

primary productivity of freshwater (lotic and lentic), estuarine and marine environments has been 

widely discussed and demonstrated (Hutchinson 1957, Schindler 1977, Carpenter et al. 1998a, 

Rabalais 2002). Moreover, broad scale nutrient enrichment - principally resulting from the use 

of modern inorganic fertilizers but also from fossil fuel combustion and land use change - has 

dramatically altered global biogeochemical cycles (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), anthropogenic influence is particularly pronounced; 

commercial fertilizer use worldwide now accounts for 140 Tg (1012 g) of N yr-1, a figure that 

exceeds the upper estimate of the quantity fixed naturally in terrestrial ecosystems (Vitousek et 

al. 1997). N delivery to freshwater and marine ecosystems is principally via riverine transport 

(Boyer et al. 2002) accounting for ca. 25% of watershed load in Northeastern rivers (Figure 1). 

The addition of nutrients is responsible in many cases for changes in the rate of primary 

production, and therefore indirectly on the composition of primary producers as competition 

constraints are altered (Smith et al. 1999). For example, non-linear shifts in lake primary 

producers moving from rooted aquatic 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 enrichment typically include changes in oxygen and light dynamics of aquatic 

systems, with deleterious effects on animals adapted to pre-enrichment conditions (Hynes 1970), 

and on whole system metabolic processes (Dodds 2006). Generally, nitrogen (N), phosphorus 

(P) or both (Elser et al. 2007) limit ecosystem productivity, and as such, the focus of nutrient 

attenuation efforts. 

While the discussion of eutrophication in lake systems is relatively mature (e.g., Schindler 1977), 

the hydraulic circumstances in flowing water systems, such as constant advection to maintain 

diffusion gradients and turbulent mixing, make generalization about eutrophication in all aquatic 

systems from lake models alone problematic. A new synthesis of eutrophication 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 eutrophication, particularly with regard to alternative 

mechanisms for changes in primary productivity (Caraco et al. 2006). As such, while 

phosphorus may be the limiting nutrient in lakes (Carpenter et al. 1998a), and nitrogen most 

frequently the limiting nutrient in marine systems (Smith 2006), the influence of nutrients on 

flowing water ecosystems 
 58 

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 since 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 applications, enhance or restore 

the landscape assimilative capacity, reduce the load from fossil 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 managing the problem is maximized by understanding 

stores, fluxes and transformations intrinsic to the system being managed. 

Addressing the observed eutrophication (principally overgrowth of filamentous algae) in spring 

ecosystems in Florida requires attention to the myriad mechanisms and potentially complex 

dynamics of nutrient loading and processing. As such, a synthesis of nutrient effects in Florida’s 

springs starts with a discussion of the source and chronology of nutrient loading. The 

mechanisms and temporal dynamics of nutrient loading to spring vents are 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 opportunities for future management to be linked. 

For example, a role of phosphorus in the decline in spring biological condition has been 

suggested. Because phosphorus concentration in spring discharges has, in general, not increased 

through time, it has been essentially neglected in the literature on springshed management (a 

point to which this review will return in depth). 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 attenuate anthropogenic loads. This particular assumption is being tested as 

part of the loading literature (Nair et al. 2004), but has direct relevance 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 quality parameters has occurred within the 

scientific and regulatory communities due to strong evidence for anthropogenic N enrichment in 

the karst environment of North Florida; evidence for contemporaneous P enrichment is absent. 

Karst is vulnerable to nitrate-N transport; indeed Katz et al. (2005) and Phelps (2004), among 

others, use N2 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 loads 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 flows), b) time series nitrate concentrations 

at 5 1st 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 include intensive human land uses 

 (Scott et al. 2004), are the principal motivating force behind this chapter. Clear upward trends 

in load are evident, as is spatial identification of sources (i.e., sub-basins shaded grey, 

Georgia 

Florida 
 60 

representing 25% of the area and 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 concentrations of springs, which discharge 

directly from the upper Floridan Aquifer, have increased substantially from near background 

concentrations (≤0.1 mg/L) (Katz 1992, Maddox et al. 1992) to > 5 mg/L at present (Katz et al.

1999). 

In general, consideration of nutrient loading requires an accounting effort. The goal is to 

tabulate accurately the inputs and outputs of a key 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 importance 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, are an important component of this report. 

Landscape loading and subsurface transport/transformations 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 understanding of nutrient loading, temporal nutrient delivery trends, travel times of 

water and nutrients, principal sources of loads, and subsurface transformations is relatively 

mature. By way of example, the current synthesis 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 the number of references compiled, illustrating the significant 

focus on sources, and more limited focus on downstream effects. 

Figure 3 – Number of publication on each of a variety of topics germane to nutrient effects in 

springs identified for this study. This result highlights important knowledge 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, leading to anastomosing flow paths, long residence times, closed surface 

basins (isolated wetlands, swallets), and dominance of inter-flow (i.e., instead of surface runoff). 

Second, the geologic basement of Florida is limestone, leading to a karst landscape (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 characterize the development of karst lead to subsurface flow conduits, 

massive groundwater flow anisotropy, 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 subsurface flowpaths means attribution of hydrologic 

and pollutant loads to particular areas is less 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 Carolina, Georgia, Alabama and Mississippi. This 

carbonate aquifer is among the most transmissive karst 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 pollution to the aquifer is one of a constellation of stressors that 

are already deleteriously affecting it. While affects on the Floridan 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 deeply 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-term weather, a condition quite different from the 

karst regions of Kentucky, Missouri, 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 basalt aquifer springs, not karst. The reservoir from which spring 

flow derives is extremely large, and as such nominal 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 peculiarity of the Floridan Aquifer karst is that flow through the matrix 

and conduits is a dynamic interplay driven by hydrologic conditions. Martin 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 high flow, major conduits dominate. This has 

implications for nutrient delivery to springs; during periods of high flow, primary short-term 
 62 

Figure 4 – Extent and confinement of the Floridan Aquifer. The Floridan 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 County) 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 important, and dispersed nutrient loading across the 

unconfined region is likely to be more predictive of what emerges from the spring vent. 

The matrix that contains the Floridan is not uniform, leading to significant anisotropy in flows. 

Vertically, limestone deposits of different ages have different properties, and harbor water with 

different chemistry. For instance, in the Silver 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 Upper 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. Laterally, anisotropy of aquifer properties is extreme, 

with high fracture porosity (Martin and Dean 2001), and complex anastamosing networks of 

large and small conduits that move water through the subsurface in predictable (Werner 2000) 

Figure 6 - Hydrogeologic units in the Silver Springs basin (after Phelps 2004, Scott et al. 2004). 
 64 

but still massively complex ways (Kincaid et al. 2004, Florea and Vascher 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 the connectivity between surface and groundwater 

(Scott et al. 2004). Where that confining 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 the 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 hydrogeologically 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 features 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 where the limestone is exposed, prevalence of 

surface water features (wetlands, stream channels, lakes) drops dramatically and are replaced by 

spring vents and karst windows where the potentiometric surface of the Floridan intersects the 

land surface. Springs and spring runs are therefore surface expressions of the level and quality 

of the aquifer. 

Similarly, the importance of the Cody Escarpment in delineating zones with distinct 

biogeochemical processes is also difficult to overstate. As observed by Hornsby (2007), above 

the scarp where the Miocene Hawthorn Formation precludes surface water percolation to the 

Florida, the surface waters and the 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 have high water tables, leads to rapid rates 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 phosphatic clays of the Hawthorn, but then decline due 

to sorptive processes occurring between the phosphate and the CaCO3 aquifer matrix. However, 

it should be noted that some springs have naturally high levels of P (to be discussed at length 

later) suggesting that P sorption may not be uniformly high. Nitrate 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 hydrologic 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 has 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 intersects 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 into magnitudes partially obscures the dramatic 

differences in flow between for example a 1st magnitude (> 65 million gallons per day) and 3rd

magnitude (< 6.5 million gallons per day) spring. Comparison of hydraulic and ecological 

processes between springs of such differing size is potentially problematic, and scaling 

relationships for generalizing process-level measurements made at a spring of a particular size 

are clearly required. Examples of factors expected to be different for varying spring magnitude 

are light regime, depth, flow velocity, 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 (Wakulla, 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 1st, 191, 2nd, 151 3rd magnitude). 

b) Nomenclature for spring types based on where and how water is discharged. 

Figure 9 – Locations of Florida’s > 700 springs, including those visited by Scott et al. (2004). 
 67 

the unconfined karst region. Figure 9 shows the 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, Chassahowitzka, Weeki Wachee springs among many 

others are formed in this manner. Tidal effects (Yobbi 1992) can complicate 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 after sinkhole capture and passage 

through subsurface conduits (e.g., Santa Fe River 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) demonstrate that there is indeed 

substantial flow-dependent mixing with groundwater. 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 subsurface) will collect at a particular spring vent (Upchurch et al. 2004b). 

More often than not, the contributing 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 aggregate 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 elevation of the groundwater potentiometic 

surface (Shoemaker et al. 2004). This presents several challenges. First and 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. Second, the potentiometric 

surface that is the groundwater terrain is dynamic, depending on rainfall and recharge rates. As 

such, the boundaries of springsheds tend to be fuzzier to accommodate the interannual 

variability. Third, the measurement 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 various focal points of springshed delineation 

and protection. Loper et al. (2005) use similar terminology (primary, secondary and tertiary 

protection zones) for the level of nutrient management recommended to reduce loads to Wakulla 

Springs. The combined inputs of surface water and subsurface drainage are extremely important 

to parse for purposes of understanding nutrient loads 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 flow course of the karst conduit (“trace”), and the 

spring. Priority areas for water quality protection are shown (Upchurch and Champion 2004). 

By way of example, Figure 11 shows the delineated springshed for the Ichetucknee complex in 

the Suwannee River Water Management District (SRWMD) (SRWMD 2002). The dashed line 

delineates areas where, over some time frame, 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 (Hawthorn Group) in the 

northern watershed leads to surface 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 Creek (which drains via surface channels to the east, and out of 

the springshed) and Falling Creek (a tributary of the Suwannee, draining via surface channels to 

the north) are in the springshed, but dominant flowpaths for the latter two are not to the 

Ichetucknee. This complicates the process of developing a hydrologic budget, which also 

complicates the ability to construct a nutrient budget. Moreover, much of the characterization of 

subsurface flow boundaries is predicated on model assumptions; the size and geography of 

springsheds appears to be strongly sensitive to these assumptions (Shoemaker et al. 2004), with 

as little as 40% spatial concordance between different models 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 pools of water in the subsurface have major ion 

chemistry profiles, and since most of these ions are 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 discriminant and principal components analyses on 

well water samples is then used to identify regions with generally similar and generally distinct 

hydrochemistry. Hydrochemical facies have been used to delineate flow paths in the 
 69 

 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 tributaries to the north (SRWMD 2002). 

Ichetucknee springshed (Upchurch and Champion 2004 – Figure 12), springs typology in the 

Southwest Florida Water Management District (Champion 2001), and broad groundwater basins 

in the Suwannee River Water Management District (Hornsby 2007). The two principal benefits 

of this is method are 1) confirmatory (i.e., vis-à-vis the potentiometric 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 Ichetucknee complex, and up-gradient wells; concordance 

between springs and certain wells suggests a hydrologic connection. The inference is that there 

are two distinct catchments (1: Ichetucknee Head, Cedar Head, Blue Hole; 2: Mission, Mill 

Pond, Devils Eye). Similar research from the St. Marks system (Barrios and DeFosset 2006) 

suggest that four groundwater flow components comprise the St. Marks Rise. 

Rose Creek 

Falling Creek 

Cannon Creek 

Olustee Creek

Olustee Creek 
 70 

Figure 12 – Use of major ion chemistry to delineate the flowpaths and source areas for 

particular springs in the Ichetucknee complex (A, B, C, D, E, F). The shape of the “star” 

represents the relative concentration of various ions, and concordance between a spring vent 

star and well samples upstream (blue stars) suggests a hydrologic connection. Clear groupings 

appear to be A/B/C and D/E/F. These springs are notably different in their nitrate 

concentration, among other water quality parameters. 

Delineation of springsheds has utility for developing budgets of water and nutrients for springs. 

However, several important caveats about delieated boundaries should be made clear. First, as 

already mentioned, certainty about flowpaths of rainfall in the geographic bounds of a delineated 

springshed is complicated by surface water drainage that may transport water elsewhere. 

Second, where there is a predominance of conduit flow for moving water through the aquifer, 

some of the water (and therefore nutrients) delivered 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 conduits bypass the spring completely, 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 comes 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 Hole, while a more local conduit 
 71 

system loads to Mission, Mill Pond and Devils Eye 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 such, delineation of the Ichetucknee 

springshed is actually a composite of smaller springsheds, each with different hydrochemical 

Figure 13 - Delineation of springshed (red dashed line) 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 nominal 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 direction, they do not imply flow velocity. Shallow 

potentiometric gradients, flow anisotropy and extremely variable recharge rates lead to the 

condition where the time of concentration (i.e., time required for a molecule of water to move 

from precipitation to a particular gaged location, in this case, the spring vent) can vary from 

hours to millennia. This leads to strongly incongruous evidence about the residence times and 

transit velocities of water and nutrients in the subsurface. For 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, however, is that the water at the spring vent is a binary mixture 

of young (days to years) and old (decades to millennia); 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 near the St. Johns River 

(Knowles et al. 2002), which may explain spatial variability in nitrates, dissolved oxygen, 

salinity and calcite saturation (Toth and Katz 2006). 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 incomplete pictures of groundwater movement, they are critical. 

Given this uncertainty, a key refinement in springshed delineation is the definition of timespecific capture zones. This technique relies on modeling and makes potentially problematic 

assumptions about subsurface flow, but the resulting 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 underscores 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 Protection Area) has highest priority due to short 

capture times and highly pervious soils. Surface Water Protection Area delineates areas that 

drain to surface features captured by sinkholes (e.g., Rose/Cannon Creek sinks) feeding spring 

conduit systems. Lowest priority is the Springshed Protection Area, where aquifer confinement 

and surface drainage minimize risk to spring water quality (Upchurch and Champion 2004). 

Hydrogeology and Nutrient Loading 

The importance of hydrogeology on nitrate loading to surface water is exemplified in the Santa 

Fe River basin, which has the contrasting hydrogeology of confined and unconfined regions. 

Specifically, the aquifer in an 820 mi2

 area of the upper basin is confined and surface water 

dominates transport, while the aquifer in a 546 mi2

 area in the lower basin is unconfined and the 

primary transport is via the Floridan. Loads by reach (Table 1) indicate that although the lower 

basin occupies only 41% of the watershed, it generated roughly 98% of the load over the last 5 

years (see Figure 2 for spatial partitioning). While 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/km2

/yr) generated by the two reaches of the Santa Fe 

River watershed during the years 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 1302 
 74 

Assessing Aquifer Vulnerability 

In a setting like the karst landscape of North Florida, water quality vulnerability is far from 

uniform; the example from the Santa Fe River basin underscores the role of geology in 

regulating contamination risk from nitrogen. Across all contaminants, 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 management 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-spots in Florida conform with areas 

where the aquifer is unconfined and land use is intensive. It is important to note that the model 

was not developed with sufficient density of karstic calibration sites to be of predictive utility. 

However, this map underscores the sorts of methods used to understand 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 Geological 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 resulting map product (Figure 15) conforms closely to environmental geology. 

Similar local efforts (e.g., Wekiva Aquifer 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). Areas in white are outside the study bounds. 
 75 

Timescales of Water and Nutrient Fluxes 

Times of flux between recharge and emergence at a spring vent represent a source of significant 

and important uncertainty, with direct relevance to understanding current and predicting future 

nutrient concentrations observed in springs. Because the drainage network consists of subsurface pathways through stores with unknown effective volume, it is generally not possible to 

use conventional hydrograph techniques 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 springshed. Phelps (2004) further reports two estimates of water age 

from the same spring made 3 months apart that are 1-5 years and >50 years associated with large 

observed shifts in spring discharge. Katz et al. (2004) suggested that recharge of recent origin 

coming from shorter groundwater flow paths and/or from features that are hydraulically 

connected to the upper Floridan Aquifer (e.g., sinkhole) could contribute 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 landscape and the outflow of spring 

vents. The first is an aggregated method based 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 are unreactive, and that no gas exchange occurs 

once water has recharged, concentrations of these 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 represent different time scales (e.g., 14C has been used for longterm age dating – Toth and Katz 2006), or checks in event of contamination. Probably the most 

promising technique is the combined measurement of 3

He and 3H (the daughter product), both 

based on nuclear bomb-testing that peaked in the early 1960’s. The concentrations 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 based on whether piston flow (limited mixing, water 

moves through the system with other water recharged at the same time), binary mixing (two 

water sources at various fractions) or exponential mixing (mixing of multiple water ages) is 

assumed (Katz et al. 2001). Evidence varies for which of these flow models is most plausible. 

Katz et al. (2001) show evidence supporting a piston flow model, while Toth and Katz (2006) 

working in some of the same systems, argue for 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 intrinsic reason to expect springs to behave uniformly with regard to 

hydrogeology, but site-specific uncertainties arising from unknown flowpaths, residence times 

and mixing dynamics significantly affect management strategies and expectations. 

The general picture that emerges from the aging measurements is water 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 provide 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 (decay 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 poorly 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 fluorescent dye (Karst Environmental Services 1997, 

Butt and Murphy 2003, DeFosset 2003, Kincaid et al. 2004, Moore and Martin 2005). In 

general, while some dye injections are never recovered over the period of observation, the travel 
 77 

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). This discordance in time domains for subsurface 

pathways is strongly suggestive that a piston flow model is not appropriate, calling into question 

the utility of the nominal water age for management purposes. Indeed, if, as demonstrated in 

Toth and Katz (2006), there is strong negative covariance between nutrient, dissolved oxygen 

(Katz et al. 1999), and ionic saturation deficits of water and its age, then it is plausible that shortresidence time flowpaths dominate nutrient delivery. 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 roughly 1 cfs of flow per square mile of delineated 

springshed area (Upchurch and Champion 2004). Strong correlations in spring discharge with 

climatic forcing are observed (Knockenmus and Yobbi 2001), but the time 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 variability (5-28 cm/s) in flow velocities over 6 months 

of observations in Wakulla Springs; the responsiveness of spring hydraulics to antecedent 

rainfall conditions is an area of needed reseerch. Interestingly, only a small fraction (6-23%) of 

the incident rainfall on the entire springshed (both confined and unconfined areas) can be 

accounted for in spring flow; since evapotranspiration (ET) accounts for only 67% of rainfall on 

a landscape basis, a significant fraction of flow (as much as 25%) leaves the 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 springshed water budgets highly problematic. 

Flow matters in springs for a variety of demonstrated and hypothetical reasons. 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 response to nutrient enrichment so well described for lakes 

(Wetzel 1983) is not as clear in running waters (Hynes 1969, Dodds 2006) where biological 

access to nutrients is constrained in part by residence times. Residence time clearly plays a role 

in the degree to which lakes respond to P enrichment (Vollenweider 1976, Schindler 1977), and 

Hilton et al. (2006) invoke residence time as a critical 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 major 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 indirectly affect biota by changing the 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 primary 
 78 

production is increases in filamentous algal growth, and algal attachment has been shown to 

respond to shear stresses induced 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 affected by such changes, with unknown ecological 

consequences. 

Evidence for flow declines are subject to some uncertainty 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 Springs (~350 cfs) are markedly lower than historic values, which 

were over 600 cfs. Champion and Starks (2001) report 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 decline at Rainbow Springs, among the largest springs in the state, 

is also evident. Garman and Garey (2005) report 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 Peace River basin, historically a 2nd magnitude spring, is 

another example where discharge ceased 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 consumptive 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 water resources is not made a high 

priority. While it is clear that such a fate has 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 continued vigilance to protect flow rates 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. Excellent 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 the global N deposited on land annually, and that 

nearly 50% of the N transported in the world’s rivers (~ 36 x 1012 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 uptake) and dissimilatory (denitrification) potential in regions with 

high water table (e.g., confined regions), and 2) negligible sorption 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 concordance 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 organic N spends very little time in the root zone 

(where plant uptake can occur) and encounters conditions unfavorable for dissimilatory removal 

in the shallow UFA, which is generally oxic (limiting the potential to use nitrate (NO3) as a 

terminal electron acceptor) and low in available carbon (the electron donor for microbial 

respiration). 

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

Nitrogen is, by far, the most well studied nutrient in springsheds. Of the 174 references on 

nutrient loading to springs synthesized for this chapter over 50% dealt principally with sources 
 80 

and fate of N. Most of the rest address flowpaths (dye trace studies, springshed delineation). 

With the exception of reports that provide reconnaissance data on numerous water quality 

parameters (e.g., Scott et al. 2004, Stevenson et al. 2004), only 5 reports address possible effects 

of other nutrients, so the discussion that follows focuses predominantly on N load (and the nitrate 

species of N in particular). 

Nitrogen has several forms in an aqueous environment; these include two oxidized inorganic 

ions (nitrate – NO3

-

, and nitrite – NO2

-

), reduced inorganic species (ammonium - NH4

+

), 

dissolved gases (nitrogen gas - N2 and nitrous oxide – N2O), and myriad organic species of 

varying lability (collectively referred to as dissolved organic N – DON and particulate organic N 

- PON). While all species can be found in spring rivers, nitrate dominates delivery 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 dominated by ammonium and DON. 

However, most springs, particularly those that are severely enriched, are dominated by nitrate. 

Temporal trends in other forms of N are unknown. Several properties 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, moving rapidly through soil profiles, 

due to low anion exchange capacity and low affinity 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 nitrate reactivity; in fact, some authors have shown that nitrate 

reactivity in groundwater is so slowed (0.013 μmol 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 μmol N L-1 d-1) more 

than compensated for this attenuation rate. Others have documented enhanced nitrification of 

dissolved organic nitrogen (DON) and ammonia in shallow 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 observed trends in nitrate-N are nearly ubiquitous 

in areas where anthropogenic loading to the springshed 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 significant increasing trend in the concentrations of nitrate-N (Ham and 

Hatzell 1996), which is primarily attributed to groundwater discharges in the Middle and Lower 

Suwannee reaches (Pittman et al. 1997). Similar trends have been observed for the Ocklawaha 

River, particularly where that system interacts most closely with groundwater (Mytyk and 

Delfino 2004). Figure 20 shows nitrate concentration trends for springs in the Suwannee Valley. 

Figure 21 shows a similar synthesis across a total of 69 springs (Strong 2004). The x-axis is N 

concentration in the 1970s and the y-axis is contemporary concentration. Note that these data are 

principally 2nd and 3rd magnitude springs; upward trends for 1st magnitude are well established in 

Katz et al. (1999), Champion and Starks (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 greater, 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 concentrations (e.g., Hornsby springs – 

Figure 22). 
 81 

Figure 20 – Nitrogen loading to major springs of Florida; shown are both nitrate and total N load. Note that foof the N (>75%) arrives as nitrate, but ammonia is important for some that have low nitrate concentrations.
 82 

Figure 21 – Relationship between early time period (1907-1979) nitrate-nitrogen and recent 

time period (1990-2003) nitrate-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 1.13
 83 

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 significant trends occur in both directions (positive = 55% and negative = 

25%), with some springs exhibiting no trend (20%); these are reported spatially in Figure 23. 

The picture that emerges is complex, and suggests that recent control efforts may be working in 

some places and not others. The prevalence of negative trends may also be driven hydrologically 

(see Figure 22) with recent drought and flood conditions clearly manifest in the signal at 

Hornsby Springs. Toth and Katz (2006) observed a negative trend between nitrate 

concentrations and age, suggesting that as spring vent flow became older (i.e., deeper source 

water during low flow conditions) nitrate concentrations declined; Einsiedel and Meyer (2006) 

observed the identical inverse relationship for karst springs in Germany. This formal link 

between discharge and age is an important research need because of its implications for 

regulatory minimum flows and levels. 

Analysis of the concordance between nutrient loading and flow may be confounded by hysteretic 

(i.e., path dependent) behavior in the delivery 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 arrive during the recessional phase of the storm 

hydrograph. Current monitoring data are insufficient to determine 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 effect 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 current sampling frequency. 
 84 

What is clear from the time series record, at least 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 expected to respond to up-gradient climatically-driven changes in 

potentiometric surfaces due to nearly instantaneous transmission of potentiometric 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 residence 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 water carries the bulk of the nitrate, an observation partially 

confirmed by negative covariance between age and nitrate concentrations (Katz et al. 1999). It is 

also plausible that the positive cross-correlation between flow and nitrate concentration is due to 

proximate sources only; if this is true, the amplitude 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 Ichetucknee was entrained by the sinkholes that capture confined region 

flow. Chelette et al. (2002) report concentrations in the Wakulla System, where nitrate 

concentrations increase dramatically between Middle River Sink (one of several key conduits to 

the vent) and the spring. Indeed, upward trends clearly observed in the concentrations at 

Wakulla are not matched by those in the sink, suggesting that either the nitrate 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 (of which some may be attenuated), compared 

with 270,000 kg N at the spring vent; this supports the contention that most of the N is 

introduced in the karstic short-time capture zones. 

Figure 24 - Flow vs. nitrate hysteresis (van Herpe and Troch 2000). 
 85 

Figure 25 - Time series measurements of NO3-N and discharge in the Wekiva Springs. 

Given the potential for disproportionate importance of local sources, the observation (Upchurch 

and Champion 2003) in the Ichetucknee springshed of nearly monotonic increases in NO3

concentrations with proximity to the trace and springs is important. In that study, distinct areas, 

mostly near the spring and tracer, were geostatistically characterized as having UFA nitrate 

concentrations in excess of 2 mg/L, which is markedly 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, ecologists 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) found 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 present in flowing water systems like springs. The 

same conclusion regarding eutrophication in flowing waters was arrived at regarding nitrates 

(Odum 1957, Duarte and Canfield 1980), suggesting that nitrogen may not have significant 

ecological effects in springs; this represents probably the most 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 deposits 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 Floridan Aquifer water are loaded heavily with 

P. Indeed the nominal concentrations of P in springs are between 50 and 100 ppb, which 
 86 

suggests that historic N:P ratios (molar basis) were less than 1:1. Note, however, that N:P ratios 

in advective (flowing water systems) are only informative regarding nutrient limitation if some 

other input (e.g., light) is not limiting. 

Figure 26 – Generic P cycle. The P cycle in Florida is characterized by low sorption in the 

Plio-Pleistocene sediments of the confined region, 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 (CaPO4) 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 the 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 confined region of the springshed (Frydenbourg 2006). This suggests 

that most of the P is sorbed during passage through 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) observed a significant negative relationship between 

water age and phosphate concentration, consistent with P enrichment by surface waters 

interacting with the Hawthorn and then following shallow flow paths to the spring vent. 

A second factor regulating P loading to the springs is the strong chemical sorption potential of 

the aquifer matrix for mineral P. While P breakthrough 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 for P loading trends that are starkly 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 concentrations in 23 springs comparing 

observations in the 1907-1979 period and observations in 1990-2003. Nominal concentrations 

are low (~50 ppb) and no significant trend is observed (Strong 2004). 

Figure 28 – Summary of TP concentrations in some of Florida’s 1st magnitude springs, many of 

which have seen significant recent increases 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 outside the scope of this review, but several key 

observations should be made. First, as in surface streams, microbial processing and trophic 
 88 

concentration of energy occur; Simon (2000) showed interesting similarities between surface 

stream organic matter processing and the processing that takes place in karst conduits, noting 

among the several differences the relative absence of nutrient limitation, presumably because of 

the ratio of flux to production. In addition to subsurface heterotrophic respiration which 

mineralizes both particulate and dissolved organic carbon, there are geochemical processes that 

result in abiotic attenuation. It is worth pointing out that the self-organizing character of the 

karst solution conduits occurs in part because of interactions between surface sources of organic 

acidity and carbonate dissolution. 

The carbon emerging in spring waters is, with few exceptions, almost exclusively inorganic. 

Nominal organic carbon concentrations 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 breakthrough 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 than 1.2 mg/L, with most less than 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 been observed. If BOD loads to the aquifer are 

significant, it will be manifest as DO changes. 

The result of organic mineralization (biotic or abiotic) is increases in solution of carbonates. 

Consequently, inorganic carbon levels are uniformly high; dissolved CO2 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 hardness (a proxy for dissolved inorganic carbon in 

a carbonate system) has increased significantly (p < 0.05) since the 1970’s (Strong 2004), with 

nominal increases of 25% (Figure 29). One speculative 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 residence times would affect the degree of calcite or 

dolomite saturation, and might also affect other water quality parameters (e.g., salts, dissolved 

oxygen and nitrate concentrations) that are ecologically relevant. Declines 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., oxygen 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 strong 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 and ultimately favors the process of nitrification 

of organic and ammonia N. Barrios and DeFosset (2006) also report high levels of DO (> 3.5 

mg/L) in the Upper Floridan, and a significant positive 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 the vent is the only variable 

significantly associated with algal 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 temporal variability of DO is implicated in ecological change, 

and, if so, whether it responds to anthropogenic stressors and management is a topic of future 

inquiry of immediate relevance to springs. Moreover, 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 system within an algal steady state even after the 

stressor that led to algal dominance in the first place is removed (Dodds 2003). It is worth 

pointing out also that high rates of P mineralization are observed in anoxic environments 
 90 

(Chambers and Odum 1990); consequently, declining DO in spring vents could make more P 

available for fueling plant or algal productivity. 

 

Figure 30 – Map showing the significant variability in Upper Floridan Aquifer DO 

concentrations, with trend towards anoxic conditions closer to the springs. Underlying the 

points are the Silver Springs springshed 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 significant declines (p < 0.001) (synthesized from Scott et al. 2004). 

Other water quality metrics have also been discussed as possibly playing a role in ecological 

changes in springs. Among the numerous confounders of the nitrate-algae relationship in springs 

is the presence of abundant algae at Silver Glen and Alexander springs, both of which have 

nitrate concentrations at background levels (~30 ppb). However, both have comparatively high 

specific conductance (> 1000 μS/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 that conductivity and alkalinity were the best 

predictors of algal abundance, while nitrate provided no predictive power. Formal survey and 

experimental consideration of this possibility should be among research 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 temperature 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, personal 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-traditional contaminants 

(e.g., benzene, bisphenol A, caffeine, DEET, naphthalene, atrazine, many others). The results 

are important, though not suggestive at this time, of a strong toxicity 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 frequently 
 92 

include bisphenol-A (polycarbonate 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 attention that has been paid to nitrate 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 long, and includes atmospheric deposition (wetfall and dryfall), 

imported fertilizer, imported food (for people and livestock) which result in domestic wastewater 

discharges (centralized systems and septic tanks), 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 terrestrial N budget, and is estimated 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 various N inputs to 16 watersheds 

in the northeastern US (van Breeman et al. 2002) 
 93 

Figure 33 – Concordance between springshed development and observed NO3-N concentrations 

in Silver Springs (Munch et al. 2006). The strongest signal appears to be from urbanization 

since agricultural lands decline 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 between 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 which 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 based on proximity to the springshed in question. Interestingly, 

trends vary quite substantially among sites; Bradford Forest has been relatively steady or slightly 

upwards for both ammonia and nitrate over the last 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 alone is used to represent deposition because 

dryfall measurement (i.e., dust) is often confounded 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 studies 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 availability of N in soils, and the prevalence of epiphytic cyanobacterial 

colonization of soil surfaces, tree 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 some basins, agricultural 

fixation (~2300-4500 kg/km2

/yr) is larger than fertilization. These values are 5-fold the expected 

total atmospheric deposition (~500 kg/km2/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 (NH4) and nitrate 

(NO3) 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 should be included because of nitrification. 

Fixation of nitrogen is an energetically expensive process that requires 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 are capable of fixing N from the atmosphere, in fact it is certain 

prokaryotes, such as microbial 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 reason that so few organisms can perform this 

process is that conditions under which it can occur are highly reducing, a setting created 

symbiotically in root nodules of N-fixing plants and in the cells of blue-green algae. Diazotrophs 

(microbes that fix N) use the enzyme nitrogenase, which is strongly sensitive to oxygen, and 8 

moles of ATP (cellular energy carrier) per mole of ammonia fixed to overcome the strong triplebonds that make atmospheric N biologically unavailable. The ammonia resulting 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 source 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 fertilizer inputs to springsheds (e.g., Katz et al. 1999) are forced to 

make assumptions about loads when county boundaries (for which sales data are compiled) do 

not correspond with springshed boundaries. Some authors (MACTEC 2007) use recommended 
 96 

fertilizer application rates (e.g., from UF/IFAS: www.edis.ifas.ufl.edu – Table 2) coupled with a 

detailed land use map to estimate fertilizer applied to the land. One problem with both method is 

that attenuation is unaccounted for; that is, as N is absorbed by plants or denitrified within the 

field boundary, it will limit the loading rate to the aquifer. The rate of uptake is not well known, 

and is expected to be quite variable 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 land 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 loading; in Obreza (2002), N loading rates by land use 

in Florida are synthesized (Tables 3 and 4). Moreover, the effects of BMPs on attenuation of N 

pollution is weak; Currens (2002) reports an increase in the median N load after implementation 

in the deeper karst aquifers in Kentucky. The data in Table 4 are more specific to loading rates 

 

Table 2. Recommended fertilizer application rates 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 applied load only (Table 2), which is that some 

of that applied load is harvested and only a fraction of the load reaches waterbodies (surface or 

subsurface). Several observations can be made regarding 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 reality is that these rates are unverifiable on a case 

by case basis, so it is hard to improve accuracy. 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 suggestive of massive 

denitrification (e.g., recreational areas, corn, improved pasture). Further disagreement can be 
 97 

observed between estimates of row-crop loads and observed concentrations in Table 4 and in 

Albert (2002), where UFA nitrate levels were up to 130 mg/L. 

Both methods (fertilizer application rates and generic land use loading parameters) are 

questionable for high-recharge low-reactivity karst landscapes and may dramatically 

underestimate the loads. By way of example, Table 5 summarizes the concentrations in wells in 

the Silver Springs springshed (Phelps 2004), which indicate strong impacts 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 event-mean concentrations nationally and in 

Florida. Data are for total N load, total N concentrations 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 Omeric Reckhow Reckhow - Reckhow

Florida 11.6 16.6 32.1 11.2 6.9 7.2 3.75

Total 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 1.31

Total 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† 0.08

Nitrate Conc. 

 (mg/L) 

Reference McConnell McConnell McConnell 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) reports 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 NO3-

N, no evidence of phosphorus enrichment was observed. Such high levels of NO3 contamination 

are emblematic of the intrinsic vulnerability of karst areas to over-fertilization. The presence of 

abundant dissolved oxygen and essentially no organic carbon suggest that opportunities for 

denitrification along the flow paths to the regional springs (Blue, Telford, Convict springs, all 

with dramatically elevated concentrations) are minimal. 

Clearly, as nutrient budgeting matures as a science, actual observations of vadose zone fluxes 

within different land uses are required to constrain estimates of load. Studies of residential 

fertilization (e.g., Cisar et al. 2004) have shown strong covariance between applied loads and 
 99 

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 differences 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 concentrations predicted by the modeling output in Table 4. Tree crops 

(peaches, citrus, pecans) have been studied for groundwater 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 declines in concentration with depth. 

Table 5. Nitrate-N concentration (mg/L) in ground-water samples grouped by land use at 

well locations in the Silver Springs 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) are 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 recharged directly to the Floridan 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 loading derived from confined animal operations, a 

source of significant nitrogen, BOD and other water quality concerns. Andrews (1994) reports 

on concentrations in the groundwater downgradient of dairy operations, and observes nitrate 

concentrations averaging 26 mg/L and exceeding 100 mg/L in some cases. Woodard et al.

(2002) also observe massive enrichment 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 concentration is unknown, but in karst areas with limited 

denitrification and rapid porewater turnover, a strong correspondence is likely. Poultry 

operations (Hatzell 1995) are similarly significant sources 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, most of that N will be as ammonium or organic 

N. That form of N is far less mobile in the environment than nitrate and thus is not observed in 

spring vents at concentrations of concern. However, the microbially mediated nitrification 

process converts ammonium to nitrate, using oxygen as the terminal electron acceptor. As such, 
 100 

high loads of dairy waste can be converted to nitrate 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 loads from septic tanks. The typical number for a 

standard On Site Treatment and Disposal System (OSTDS) is that it yields 3-4 kg N per person 

per year; assuming that 70% (Anderson and Otis 2000) of that N reaches the groundwater, a 

typical household is expected to discharge 5-7 kg NO3-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 are 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 experience (Figure 35). In particular, that work shows that drainfield 

concentrations of both N and P are extremely high (3-4 orders of magnitude higher than ambient 

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

environment conditions) and that P and N attenuation 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 sites where groundwater discharges to the surface there is typically a 2-

3 log (2 to 3 orders of magnitude) decline in dissolved inorganic P (DIP), and typically much 
 101 

more; the majority of samples along the flow path are within a factor of 10 of ambient 

concentrations (i.e., 3-log reduction). In contrast, 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 wastewater 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 for 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 unconfined 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 comparison of bulk loads, but the substantial uncertainty 

about the fraction of load that actually reaches the aquifer has led researchers to use the ratio of 

stable nitrogen isotopes (14N and 15N) in a sample to deduce the source (Hornsby 1994, Panno et 

al. 2001). This method has a long history of being used in surface flow tracing, and has been 

carefully explicated, including several important caveats, in Kendall (1998). The central premise 

is that animals preferentially excrete the lighter isotope in urine, resulting in enrichment in 

organic wastes; nitrification followed by denitrification during storage (e.g., pipes, septic tanks 

and waste lagoons) and transport (through soil and aquifer) further fractionates organic N with 

resulting waste stream nitrate-enriched with 15N (10-20‰). Since fertilizer 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 loading of N to Alexander Springs was 

from on-site sewage disposal; while high 15N enrichment in the vegetation (~8-12‰; note that 

vegetation incorporation is typically a lightening process) is suggestive of unusual isotope 

loading, differences in flow between 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 alone (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 tracing N sources in springs (e.g., Toth 1999, Champion 

and Starks 2001), and the results suggest that, in the 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 sources: 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 δ15N values measured in springs and groundwater basin 

Location N δ15N 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 δ15N 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 and therefore indicative of inorganic sources. This neglects 

fractionation of N isotopes during nitrification of organic and ammonia N to nitrate; during this 

process, the heavy isotope is preferentially excluded, 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 inference relates to the omission of biological nitrogen fixation 

(BNF) among the sources. Generally speaking, the fixation of N2 gas (defined as 0‰) from the 

atmosphere induces some kinetic fractionation, preferentially uptaking the lighter (14N) 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 represent mixture sites for multiple nutrient 

sources, our mechanistic understanding of flowpath changes in 15N and subsequent mixing needs 

to be improved before isotopic information is used with authority. Moreover, there is a 

significant need to understand the temporal variability of isotopic signals; Phelps (2004) reports 

values from Silver Springs that vary dramatically (~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 discrepancies in the tracing of sources that are likely fruitful areas 

for further inquiry. Principal among these is the observation based on loading estimates that the 
 104 

Tallahassee sprayfield is responsible for over 70% of the total anthropogenic N enrichment in 

Wakulla Springs, while the nitrate isotopic measurements are consistent with fertilizer sources 

(6.3‰ – Chelette 2002). For reasons related to the complexity of nitrogen fractionation, even in 

the subsurface where denitrification is presumed to be negligible, it is relevant to ensure that 

isotopic measurements are being interpreted correctly. 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 potential that our use of this binary 

inference about sources is incorrect. Careful tracing of isotopic fractionation from source to vent 

is necessary to establish the appropriate manner in which to interpret δ15N 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 recent example of a detailed springshed N budget 

(MACTEC 2007), wherein they conclude that a) only a small fraction (~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 catchment yielded a similar conclusion about the 

principal source of N. Fertilizer, principally from improved pastures, was estimated to contribute 

almost 50% of the total load (Table 7 and Figure 39); horse farms and confined cattle operations 
 105 

were responsible for an additional 30% of the load. Overall, atmospheric deposition was also 

important as a source (over 15% of the total load) while septic tanks and sewage discharges were 

considered negligible. This is somewhat consistent 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 loads 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 attenuation 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 region 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 deposition 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 all the effluent loads occurring in the unconfined 

region, despite most of the fertilizer application occuring in the less vulnerable confined and 

semi-confined regions; part of this may be explained by sinkhole-conduit systems that link large 

regional loads (e.g., Tallahassee’s sprayfield) to the 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 the state where surface water and 

high-carbon soils and sediments co-exist. 

Foster (2008) concludes that atmospheric deposition is a large portion of the load in Ichetucknee 

(Figure 41); the fate of atmospheric N is unknown, 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 about the magnitude of P sources in springsheds. While there 

have been numerous studies of P loading from biosolid application (Graetz and Nair 1995), 

septic tanks (Reay 2004), wastewater sprayfields (Loper et al. 2005), and fertilizers, there is 

essentially no consolidated information on the relative magnitude of these sources in the fluxes 

observed at springs. Part of that is almost certainly because no significant changes in P loading 

have been observed (e.g., Figure 28, which shows springs with marked N enrichment exhibiting 
 108 

P concentrations unchanged from inferred background levels). Moreover, most 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 attention on N is warranted by two observations: 1) P loading 

is not increasing in any systematic way, so using it to explain recent ecological changes in 

springs is problematic, and 2) geology is the principal regulator of P loading, so loading is not 

management sensitive. Stevenson et al. (2007) echo this observation, suggesting that while 

modest correlations between algal 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 groundwater 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 mobility; this area is sandy and wet, so the 

generality of these results to springs more broadly may be confounded by characteristics of the 

flowpath. This finding is particularly surprising because surface water 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 between submarine groundwater discharge flowpaths 

and the karstic flowpaths to freshwater springs is unknown. 

On the latter point regarding the geological controls 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 because of the presence of the Miocene Hawthorn 

formation, previously described. Among the many important features of this interbedded 

clay/sand layer is the abundance of apatite minerals (CaPO4) 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 Hawthorn is expected to dissolve significant 

quantities of P. If this weathering/erosion process 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 used for mining phosphate, illustrating the potential significance of 

proximate P sources. No systematic effort to understand the links between geologic setting and 

P concentrations has been undertaken, but it is clear that substantial and unexplained variability 

in P concentrations exists. Among the hypotheses that might be tested are that P levels are high 

in springs with flowpaths that are sufficiently shallow to interact with the Hawthorn, or in close 

enough proximity to the Cody Escarpment where 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 principal filamentous 

algal species of concern) (Stevenson et al. 2007), additional attention to P as a driver of springs 

ecology is warranted. As such, efforts to understand 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 negligible 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 creates a backwater event, also 

during extreme floods. The delivery of dissolved inorganic carbon, a necessary condition for 

driving high levels of submerged aquatic production, is clearly sufficient. Calcite saturation 

indices and alkalinity, both of which measure the degree of cation (Ca, Mg) saturation, but also 

reflect the carbonate concentrations 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 δ15N (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 leachate, 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. Modern systems (drainfields and 

sprayfields) allow greater oxygenation 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 and surface water basin of the Wekiva 

(MACTEC 2007). By extrapolation, 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 unimportance 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, and high conductivity water delivered more evenly. 

Extrapolating that relationship suggests 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, for 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 that the source is human waste waters, the time 

domain for delivery, the hydrogeochemical controls 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 monitoring for temporal trends, but probably not yet 
 110 

for detailed accounting of sources. Analysis of ecological impacts, on the other hand, is 

extremely important because of evidence of ecotoxicological consequences of even relatively 

low concentrations. 

SINKS AND TRANSFORMATIONS OF NUTRIENTS 

Typical nutrient sinks in watershed budgets include assimilation by biota (including harvest and 

export for crops, wood and animals), dissimilatory 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 for loss of N (plant uptake, crop removal, 

denitrification, riverine export) and conclude that denitrification 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 which 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 denitrification. 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 removal 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 

Denitrification (potential) Little River, Georgia 31.5 

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 NH4+ adsorption Maryland 0.8 

Tar River, NC 16.2 Microbial Immobilization 

Coastal Plain, Georgia 87.0 

Tar River, NC 15.5 

Plant Uptake Little River, Georgia 51.8 

Maryland 15.0 

Denitrification, the most important long-term sink for N, is principally constrained by the 

availability of carbon (denitrifying bacteria are heterotrophic) and the absence of oxygen (when 

O2 is present it provides a thermodynamically favorable electron acceptor); the process may also 

be nitrate limited (in low concentration settings), but is rarely constrained by the presence of the 

microorganisms. Indeed, Andrews (2000) observed that, while all other conditions for 

denitrification were unfavorable, counts of denitrifying bacteria were frequently greater than 

10,000 per 100 mL. 
 111 

Figure 42 – Summary of measured N loss rates from selected wetland and riparian zone studies 

(from Mitsch et al. 2001). To convert g N/m2

/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 uptake (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 drive heterotrophic microbial metabolism. 
 112 

Evidence for denitrification once nitrate is in the groundwater is absent. Several lines of 

evidence appear to confirm that nitrates are effectively conservative once in the UFA. First, as 

reported in Katz et al. (2004) and Phelps (2004), the partial pressure of N2 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 N2 concentrations; this condition was observed at 

only one of 14 locations in Suwannee and Lafeyette counties (Katz and Bohlke 2000). Second, 

the Rayleigh distillation model for isotopic kinetics (Kendall 1998) in a closed system predicts 

that processes in which fractionation occurs (i.e., preferential use of 14N vs. 15N during 

denitrification) would tend to increase the “heaviness” of the remaining nitrogen (Einsiedl and 

Meyer 2006). As such, with denitrification, we would expect to see δ 15N of the groundwater 

steadily increasing with age. While there is no study of that particular relationship, the large 

fractionation due to denitrification (~30‰ – Kendall 1998) and generally low values of δ15N in 

spring discharge indicates that this process is small in magnitude if it occurs at all. Finally, the 

UFA is characterized by oxic conditions (Barrios and DeFosset 2006) and low levels of DOC 

(Katz et al. 2001, Phelps 2004) which suggests an environment unconducive to denitrification. 

 

Figure 43 – Chart comparing concentration of N2 and Ar gas dissolved in spring water (Katz et 

al. 2004). The fact that observations cluster near the air equilibrium line and at a position that 

indicates a 20-25 degree C recharge temperature suggests the absence of denitrification. 

Given that N is likely to be nearly conservative in the subsurface, efforts to attenuate loads prior 

to transport through the soil to the aquifer are critical. Clearly, land use planning figures 

prominently in reducing loads (see sections above). Best management practices can also, in 

some cases, provide attenuation, but the degree of efficacy is likely to be a function of the extent 

to which water remains at or near the surface where BMPs can operate. In areas with rapid 

recharge (i.e., karst plains), opportunities to store and treat stormwater are limited. Some of the 

conventional treatments for managing diffuse loads of nitrogen (denitrification walls, lagoons, 
 113 

treatment wetlands) are limited by hydrogeologic constraints on surface water storage and 

residence times; as such, managing/reducing sources is likely to be far more effective than 

managing/enhancing sinks. Some nitrate removal in the riparian/hyporheic zone after spring 

emergence is likely (Mattson et al. 2006) and sensitive to land use practices (Hearne and 

Howard-Williams 1988), but also largely unquantified, and generally irrelevant for protecting 

headwater areas. 

Phosphorus 

P sinks in carbonate landscapes are overwhelmingly dominated by mineral adsorption. While in 

other parts of Florida, movement of P through the sandy aquifer is rapid and 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 saturation 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 stronger association of ecological change with P 

rather than N, and b) the time domain of water quality recovery after 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 objective to determine the degree of sorption saturation 

along principal flowpaths, and the timing and implications of future breakthrough. 

In a carbonate aquifer system in the Florida Keys, Dillon et al. (2003) used radio-labeled 32PO4

to demonstrate rapid adsorption onto the Key Largo limestone. That research established an 

equilibrium phosphate concentration in solution after extended interaction with phosphate-rich 

water of ~ 26 μM, which is markedly higher than concentrations observed in springs. Desorption 

isotherms, during which the sample that had been exposed to high levels was exposed to 

phosphate-free water, showed convergence to the same concentration, suggesting that this is the 

equilibrium P concentration of that limestone substrate. While inference from this distant 

experiment is perilous, it does suggest that the capacity for limestone to sorb P is extremely 

large. The kinetics of this uptake is not well-known 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 calcareous soils in South Florida, Jaber et al. (2006) report extremely 

low P levels (0.04 mg/L) in groundwater beneath lands to which organic amendments were 

added. These concentrations support the contention that 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 under 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 population density and under reducing conditions; the results of that 

study , however, are not applicable to most of the Floridan Aquifer 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 strongly limited in the Floridan Aquifer, most aquifer water is 

saturated to super-saturated with respect to CO2, a condition that leads to considerable degassing 

of CO2 at spring vents. One of the consequences of this is proximate declines in pH, and the 

super-saturation of groundwater with respect to calcite in spring river water (with important 

implications for karst channel development). 

Aquifer flowpath attenuation of organic carbon is nearly complete, based on typical low DOC 

conditions observed in springs. While episodic pulses of DOC rich water stains the vent flow at 

Wakulla Springs, among others, DOC levels in spring vents are ~ 2 mg/L or less. Numerous 

pathways for organic carbon attenuation 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 porosity (see Simon 2000 for karst aquifers 

generally). Other possible sinks include abiotic processing (carbonate adsorption – Zimmerman 

et al. 2006), though the relative importance of this pathways is, as yet, poorly quantified; areas 

where surface water interacts most with groundwater are dominated by microbial processes, 

while areas with relatively lower levels of connectivity exhibit DOC breakdown products 

indicative of abiotic attenuation pathways. The relative importance of the different pathways is 

relevant to the karstification of the carbonate matrix. 

Other Water Quality Parameters 

Much of the literature on other water quality parameters focuses on the presence of emerging 

contaminants (e.g., DEET). Surprisingly little is published about sinks for DO, despite the 

obvious merits of understanding behavior in the subsurface. What can be inferred from existing 

data is that natural attenuation is small (high DO is observed 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 conservative. Overall, however, there is limited 

information on the extent of attenuation, nor on the factors that might lead to variability in 

attenuation rates. 

Springshed nutrient budget – Synthesis of Sinks 

Numerous springshed nitrogen budgets exist and, despite subtle differences in assumptions and 

data quality, offer an opportunity 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 sources (previously presented) and sinks are quantified 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 sink/transformation pathways across 16 watersheds, 

and illustrates a level of detail not matched in Florida’s springshed examples. It is important to 
 115 

note that Boyer et al.’s study sites are all surface water systems, 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 cases 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 explicitly account for N sources and sinks would aid 

in understanding some of the curious results obtained, particularly with regard to the degree to 

which N loads are attenuated in springsheds despite 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, these may be less important when 

allocation of load reduction responsibility is the principal objective. 

One striking thing about the N fate pie-chart (Figure 44) for surface water systems is the 

magnitude of riverine export. On average, rivers 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 contribution 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 storage potentials are higher. 

Figure 44 – Summary of the nominal fractional importance of various 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 Florida reveals the opposite result (Figure 45). 

For 4 of 5 springsheds (Ichetucknee is the exception), 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, and observe large attenuation (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 that spring is of organic origin (e.g., 

wastewater sources) despite low δ15N levels. 

There are several plausible explanations for the observed attenuation in springsheds. First is that 

intrinsic time lags (Katz et al. 2001, Toth and Katz 2006) make comparison of contemporary 

spring fluxes with contemporary loadings 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) would be expected to have concentrations near 

background, confounding budgeting efforts. While this is certainly possible, it is likely, as 

previously discussed, that most of the N is delivered via flowpaths with short residence times. 

Figure 45 - Estimated N attenuation from 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 between the surface and subsurface; levels of total 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 springshed; 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 flowpaths 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 occurring 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 observed attenuation is the result of biological 

processes. While it would be surprising to observe accelerated denitrification in the karst 

regions, denitrification in the confined and semi-confined regions is clearly 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 suggest, however, that these levels of 

attenuation are confounded. 

Regardless of the mechanism for the observed attenuation, it is clear that 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 springshed 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 water quality parameter of choice, understanding the 

springshed and flowpaths is a crucial baseline. Further work refining these concepts should 

remain a priority. 
 118 

UNKNOWNS AND RESEARCH PRIORITIES 

The preceding synthesis of nutrient loading, and, perhaps 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 dynamics 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 mixture of water discharging from spring vents. 

Given the discordance between observed dye tracer residence times and nominal ages, a 

research program to better understand water age over time seems relevant. 

Decomposition, auto-correlation and cross-correlation analyses of resulting time series 

should enable improved understanding of flow regimes and also links with other water 

quality variables that are strongly relevant to springs ecology (nutrients, salts, dissolved 

oxygen). Specifically, if binary mixing models are useful simplifications of spring vent 

hydrology, time series will allow decomposition of flows into two components, with 

insight into the relative importance of each, and changes in mixing with regional climatic 

trends. These methods need to be applied cognizant that different springs are likely to 

have different hydrogeologic properties, 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 flow in what are generally 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 systems, 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 

springs ecosystems, and linking that to how water 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 water quality parameters (dissolved oxygen, salts, 

toxins) have been invoked for catastrophic shifts in other systems, particularly 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. 

2. Improved N isotope tracing – There is significant uncertainty about the interpretation of 

bulk stable isotope measurements in complex karst hydrologic systems. First, there is a 
 119 

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 (lightening) 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 questions beyond nitrogen delivery, since there 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 and pollutant sources. 

3. Water Quality beyond nitrogen. 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 salinity, 

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. There 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 activities. 

4. BOD Loading. Septic tanks, animal waste disposal systems and sprayfields generate 

enormous quantities of biological and nitrogenous oxygen demand if discharged 

unoxygenated into the groundwater. Indeed, 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 broad spectrum of springs (Scott et al. 2004), and were the 

only parameter significantly associated with algal cover in Stevenson et al. (2004), 

understanding the roles of hydrogeologic and anthropogenic processes in low DO at 

spring vents is a first-order research unknown. 

Subsurface Nutrient Processing 

1. Explanation of observed assimilative capacity. Efforts analogous to the springshed 

nutrient budgeting for watersheds in the northeastern 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 unimpeded, conservative estimates of landscape 

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), 

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 priority. Among the issues that will need to be 

dealt with are the intrinsic lags due to aquifer 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 load 
 120 

breakthrough. The potential for observing this effect is low for nitrogen pollution, but 

may be high for P. Ensuring that P saturation fronts are not advancing towards spring 

vents is important for forecasting.

3. Management responsiveness – As we begin to understand the links between springshed 

activities and responses in spring ecosystems, evaluating the extent to which parameters 

are management-sensitive is relevant. 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. 
 121 

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3 

Biogeochemical Processes and Implications 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 determined by an interaction of physical, chemical, 

and biological processes in an ecosystem. This interaction of processes, collectively known as 

biogeochemistry, is important as it determines the forms, transformations, and ultimate fate of 

nutrients in a given system. This chapter focuses 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 throughout 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 display indicators of major processes and nutrient effects (e.g., 

through enzyme activities). 

• A variety of processes are known to occur in springs including organic matter 

decomposition, nitrogen fixation, nitrification, denitrification, iron reduction, sulfate 

reduction, and associated processes involving phosphorus. The relative importance of 

each process depends greatly on the conditions of individual springs. 

• Current data have increased our understanding of nutrient sources, and have provided 

indications of some nutrient transformations such as overall system productivity, 

potential biotic uptake and transformation of inorganic nutrients, and potential losses of 

nutrients from the system (e.g., gaseous 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 nitrate levels in many springs; however, 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 biogeochemistry in other aquatic systems, there is little 

information regarding biogeochemical processes in spring systems where the focus has 

mainly centered on water column nutrients/water quality indicators. Consequently, our 

current biogeochemical models of Florida spring systems are more of a black-box with 

emphasis on water column nutrient levels and little understanding of individual processes 

or differences between systems of various types. 

• There is little information to ascertain the involvement of processes/cycles controlling 

nitrate levels (e.g., plant uptake versus denitrification) or alternative pathways affecting 

nitrate conversion and loss (e.g., iron and sulfur involvement in lithotrophic 

denitrification). 

• 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 explain 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 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 spring run characteristics contributing to 

nutrient uptake and transformation. 

o Characterization of sediments (e.g., nutrient 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 deposition of organic matter, however, 

as yet there are no studies to document the potential for these activities to alter 

biogeochemical conditions (i.e., creating anoxic zones) and cycles (e.g., 

increasing denitrification loss of nitrate or increasing release 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, nutrient 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 growing concern, particularly in Florida where a 

significant portion of the groundwater interacts with karst geology. 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 interaction of biological systems with geochemical 

reactions which determines the overall cycling of elements within ecosystems. Both abiotic and 

biotically-controlled geochemical reactions can affect nutrient and mineral solubility, as well as 

reduction/oxidation (redox) chemistry, which can affect the composition and activity of the 

biotic system which in turn affects the water chemistry. Whether a spring system is a source or 

sink for nutrients and pollutants in the landscape is determined by the coupled biogeochemical 

processes occurring in its water column, sediments, biota, and adjacent floodplains. 

The potential biogeochemical reactions and pathways occurring in springs 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 articles 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 forced 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 similar processes occurring in lotic (flowing) systems. 

SPRING ENVIRONMENTS/ORGANISM GROUPS 

Spring water chemistry

Water chemistry is one of the most important characteristics 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 anion) and, 4) calcium sulfate. 

This review will focus on the dominant spring water type in Florida - the hard water type, which 

has high calcium-magnesium-bicarbonate concentrations 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 spring vent or boil and is directly introduced into conditions at the 

earth surface. Groundwater (at more or less equilibrium with subsurface temperature/ pressure/ 

mineral compositions) changes dramatically when 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 CO2 in ground waters, which in limestone systems is 

responsible for the dissolution of the aquifer bedrock to form karst features. The equilibrium 

reaction for this process is given by the equation: 

CaCO3 + CO2 + H2O Æ Ca2+ + 2HCO3

- 

Dissolution of limestone and dolostone results in high levels of alkalinity (carbonate buffering) 

and dissolved constituents (e.g., calcium, magnesium, chloride, sulfate) which give rise to the 

chemical composition of most spring waters in Florida. 

Figure 1 – Depiction of degassing and oxidative conversion of chemically reduced species 

(Red.) to oxidized (Ox.) forms following emergence of groundwater in a spring vent/boil. 

Emergence of groundwater at a boil or vent results in an immediate disequilibrium of the water 

with respect to CO2, and as a result, degassing occurs. Loss of excess CO2 shifts the dissolution 

equation toward precipitation of CaCO3 which results in supersaturation of groundwater with 

respect to CaCO3 (Dreybrodt et al. 1992). Other gaseous constituents of groundwater have the 

potential to degas upon emergence including N2 (produced through aquifer (denitrification), the 

greenhouse gas N2O from (from incomplete nitrification/denitrification processes), and caustic 

H2S from sulfate reduction (evidenced by a rotten egg smell). 

Supersaturation of CaCO3 may further result in CaCO3 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 CaCO3 precipitation can result in mineral encrustation in localized 
 139 

areas of some springs, especially where processes such as photosynthesis further enhance 

supersaturation through uptake of CO2. This process can result in mild calcifications on surfaces 

of submersed aquatic vegetation (SAV) or algae filaments (Berrie 1992). 

If sufficient amounts of organic materials are present in the groundwater, respiration processes 

within confined aquifers result in a depletion of available oxygen potentially creating anaerobic 

conditions. Many springs report dissolved O2 levels <1 mg L-1 (Scott et al. 2004). Under 

anaerobic conditions, other oxidized species (e.g., NO3

-

, Fe(III), and SO4

2-) can become 

chemically reduced (forming e.g., NH4

+

, Fe(II), S2-). In extreme cases, the highly reduced gas, 

methane (CH4) 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 reduced spring waters encounter atmospheric levels of O2 leading to a 

suite of chemical oxidation reactions: 

S2- + O2 + 2H2O Æ SO4

2- + 4H+

NH4

+

 + O2 + H2O Æ NO3

-

 + 6H+

4Fe(II)2+ + O2 + 10H2O Æ 4Fe(III)(OH)3 + 8H+ 

CH4 + O2 Æ CO2 + 4H+ 

Two important consequences of each of these oxidation reactions include: 1) consumption of O2

and 2) the production of H+ ions. Depending on the redox status of the groundwater, 

consumption of oxygen may continue even after springwater emergence. As groundwater 

species are oxidized, the resulting acidification also increases the potential for dissolution of 

CaCO3 (in systems supplied with highly reducing groundwater) while degassing and 

photosynthesis (reducing CO2 levels) results in precipitation of CaCO3 (Guasch et al. 1998). 

The balance between these two extremes determines the extent of calcification processes which 

is an important determinant of vegetative species composition (particularly algal assemblages) in 

spring systems (Freytet and Verrecchia 1998). 

Reactive Zones 

The high energy, lotic characteristics of springs create a variety of different zones where 

biogeochemical reactions can occur. The most important zones are the interfaces of the water 

column with sediments, epilithic/periphytic biofilms, 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 chemical parameters (e.g., pH, O2, redox). In addition, solutes are actively 

produced and consumed creating concentration 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 may be present. In a stream or a spring run, the 

presence of rocks and stones, litter, detrital material, logs, plants 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 occurring 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 that 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 both 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 dominated by algae, diatoms, bacteria, and fungi. 

Exuded C provides the basis for other heterotrophic organisms to colonize, and as the biofilm 

develops, respiration can consume available oxygen to create anaerobic zones within the biofilm. 

The creation of anaerobic microsites and layers allow a myriad of other bacterial groups to 

become established. These groups represent important classes of biogeochemical processes 
 141 

including denitrifiers, other anaerobic (iron reducers, sulfate reducers), and other aerobic 

organisms (nitrifiers, iron/sulfide oxidizers). 

Benthic Algal Mats 

Conspicuous mats of benthic algae are becoming a common feature of many spring systems 

(Stevenson et al. 2007). These macroscopic algal growths, similar to biofilms, are actually 

communities of microorganisms including: algae, bacteria, and fungi. The initial formation of 

mats is usually based on the structure provided 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 common algal taxa forming these mats in Florida 

systems are Lyngbya and Vaucheria, however a variety of other filamentous 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 characteristics. Physical gradients (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 become structured (or in some 

species, migrate vertically) to maximize their location in the presence of light and temperature. 

(Stal 2000). With the localization of C fixation, nutrient gradients develop in response to 

increased autotrophic and subsequent heterotrophic growth. Gradients of reduced and oxidized 

compounds also form and drive mat structure and 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 oxygen, pH and C production. As a 

result, the character (i.e., species composition and physical characteristics) and associated 

functions of an algal mat can be drastically different both vertically and temporally within the 

mat. 
 142 

Figure 3 – Illustration of the sampling device and example of the measured profiles of nutrients 

from the study of benthic macroalgal mats by Stevenson, et al. (2007). 

Sediments 

In addition to a dissolved load, springs from carbonate 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 derived from the direct surface connections with 

underground conduits, or from particles 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 discharge. For example some springs, normally clear 

under low flow conditions, become turbid after storm events. 

Not much is known about the composition and distribution of sediment types in Florida springs 

and spring runs, including information regarding particle size, mineralogy, or organic content. 

To our knowledge, only one study has attempted to characterize sediments/subaqueous soils 

from a spring system (Saunders 2007). In other stream environments, flow velocity is the most 

important determinant of sediment distribution patterns, 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 decrease 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 Chassahowitzka River where the areas near the boil were 

dominated by sands low in organic matter, and areas further downstream consisted of finer sands 

with higher organic content. 
 143 

The main biogeochemical feature of spring run sediments is the presence or absence of redox 

gradients. In sediments, microbial respiration frequently consumes available O2 leading to the 

formation of anaerobic zones at depth. Under these conditions O2 contained in the water will 

diffuse downward to the anoxic sites, creating a gradient of O2 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 acceptor, while in anaerobic layers, facultative 

and obligate anaerobes function using alternate electron acceptors (e.g., NO3

-

, Fe(III), and SO4

2-) 

in their respiration. 

One feature of sediments in high energy lotic systems is the high percentage of coarse particle 

size classes (e.g., sands) which greatly enhances advective processes (Figure 4). Sediments of 

high energy spring runs are also more likely to include varying stratigraphies with layers of 

different composition (e.g., organic 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 reactants and products are more easily exchanged between reactive 

zones to greatly increase reaction rates (Harvey and Wagner 2000). 

Characteristics of each of the sediment habitats in spring systems give rise to a variety of 

functional groups of organisms. Varying redox and nutrient conditions in sediments of springs 

likely result in the coexistence of both aerobic and anaerobic genera. This increased diversity of 

organisms and metabolisms should enhance the potential biogeochemical cycling of nutrients 

such as carbon, nitrogen, and phosphorus. Bacteria, in particular, are constantly responding and 

adjusting to carbon and nutrient availability, and thus, their activities can indicate conditions of 

availability. One of the ways microorganisms react to environmental conditions is through 

expression of enzymes which play 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 environmental conditions (e.g., nutrient/contaminant impacts) of aquatic 

systems (Chrost 1991; Sinsabaugh et al. 1997). 

Figure 4 – Diagram of characteristics and hyporheic flow processes related to biogeochemical 

functions in a hypothetical spring run sediment profile. H, hyporheic 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 involves 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 autotrophic production within the spring run would be 

considered autochthonous (within-system) sources while the remainder are C sources external to 

(or allochthonous to) the spring run. 

In aquatic systems, carbon is predominantly present as dissolved inorganic carbon (DIC), 

dissolved organic carbon (DOC), and as particulate organic (POC) 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 photosynthesis. 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 sources of C to spring systems is through the connections of many 

spring runs with extensive floodplains with forested, 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., variations in rainfall and spring flow/discharge) can also lead to the 

introduction of variable amounts of high DOC surface water into groundwater systems (Pronk et 

al. 2006). Under high flow conditions there are also greater inputs 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 are likely important as 

autochthonous C sources in springs. These exuded 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 composition and production of extracellular C is 
 145 

specific to both organism type and growth conditions including 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 fixation in macrophytes have been observed (Wetzel and 

Penhale 1979), while in some species of cyanobacteria, 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 Vincenzini 2003; Renstrom and 

Bergman 1989). For this reason, light availability, and rates and patterns of algal photosynthesis/ 

photorespiration, can be a significant determinant 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 and Wetzel 1995; Espeland et al. 2001). Consequently, microbial 

associations have been shown to alter the composition 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 must hydrolyze complex polymeric OM into 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 compounds, secondary production is based on C derived 

through decomposition. In most aquatic systems, this is the dominant pathway of C flow, and is 

the primary influence on the metabolic 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 inhibiting 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 large pools. In a slow moving spring run (upper 

Chassahowitzka River), Saunders (2007) observed large (>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 velocities and rates of sediment and organic matter 

accumulation (Cotton et al., 2006). For this reason, seasonality 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 complex process beginning with the physical 

breakdown of macroscopic plant/algae tissues. The 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, lignin and associated proteins), other water-
 146 

insoluble components (including fats, waxes and resins), and mineral matter (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 contain more soluble and less 

structural C compounds than vegetation in the riparian zone, algae contain even fewer structural 

compounds. Approximate ranges in values for these constituents in plant material, peat, and soil 

organic matter are shown in Table 1 (Reddy and DeLaune 2008). 

The initial step in the decomposition process is leaching of water soluble organic and inorganic 

compounds as a result of cell autolysis. Water soluble components include soluble 

carbohydrates, nucleotide bases, fatty acids, and amino acids. These components are readily 

bioavailable and serve as energy and nutrient sources to bacterial and fungal communities 

colonized on detrital plant tissue. Physical leaching of soluble components is generally complete 

within a few weeks of plant senescence, depending on structural components, temperature, and 

water flow. Leaching can be a significant factor 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 immersion in the water column (Benner et al. 1985). 

Table 1. Major carbon components of plant tissue, 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 particulate in nature and must be broken down. In other 

aquatic systems, macroinvertebrate colonization is a key factor affecting observed litter 

decomposition rates; however, studies of their effect in springs are few and inconclusive. For 

example, in a study of four springs in Minnesota, Bortodziej and Perry (1990) found similar rates 

of decomposition of watercress (Nasturtium officianale) in different spring types despite 

significant differences in macroinvertebrate colonization. In the same study, they also found that 

macroinvertebrate colonization was correlated with decay rates of Boxelder (Acer negundo). 

Results such as these may indicate that macroinvertebrate shredder and grazer function may 
 147 

depend on plant biomass type/quality (e.g., high versus low contents of structural compounds) in 

high energy spring environments. 

Decomposition of particulate matter into smaller particles is an important process in determining 

C availability for secondary production and biogeochemical metabolism in spring systems. Plant 

detritus (either attached or detached from the plant) undergoes physical fragmentation through 

the action of waves and currents, UV exposure, and through grazing 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 activities). 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 macroinvertebrate shredders results in a transport of 

C particulates from the headwater reaches and a steady accumulation of fine particulate organic 

matter (FPOM) at lower stream zones (Webster et al. 1999). As it travels downstream, this 

FPOM material becomes increasingly more bioavailable through continued decomposition and 

the incorporation of added nutrients 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 molecular weight organic compounds associated 

with plant detritus and soil organic matter, which in their original form are too large for transport 

into the periplasmic space of the cell (Weiss et al. 1991). Microorganisms including bacteria and 

fungi are the dominant producers of extracellular 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 require a multicomponent enzyme system including 

a wide range of microbial species. Rarely can a single microorganism 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 limitations to enzymatic breakdown include slowing 

diffusion of enzymes and substrates, adsorption of enzymes on solid phases, and complexation of 

enzymes to humic substances (Sinsabaugh 1993; Wetzel 1991). Complexation of humic 

substances with enzymes may not be a significant issue in hard water springs as major cations 

(e.g., Ca2+) react with humic acids and reduce the precipitation 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 Ca2+ and Mg2+ 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 = Enzyme 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 measured as indicators of organic 

matter decomposition in wetlands and aquatic systems. 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 carbon degrading enzyme expression in Florida 

springs. 

As decomposition progresses, soluble or easily hydrolyzed compounds are selectively degraded 

yielding increased recalcitrance with age of detritus. Decomposer biomass is also incorporated 

as populations wax and wane. In this way, organic matter in waters and sediments can be 

viewed as a complex of plant, microbial, and animal products in various stages of 

decomposition. Continued chemical alteration of the C mixture ultimately results in the 

formation of highly recalcitrant 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 systems. Nitrogen levels in many springs statewide have shown 

steady and dramatic increases over the past half century. In many cases, this excess N could 

result in significant changes in productivity; however, 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 within-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 (N2), nitrate (NO3

-

), nitrite (NO2

-

), ammonium (NH4

+

), 

and trace gases such as N2O. Organic N is derived primarily from amino-bonded C (-NH2) such 

as amino acids and proteins, and nucleic acids, amino sugars. Generally, the sum of NO3

-

/NO2

-

, 

and NH4

+

 measured after filtration is taken to represent the dissolved 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 cycle processes. From Reddy and DeLaune 

(2008). 

Process Enzymes EC number Substrates 

N2 fixation Nitrogenase 1.18.6.1 N2 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 reductase 1.7.99 nitrogenous substrates 

Protein degradation Peptidase 3.4.21 Amides 

 

Biological Nitrogen Fixation 

Nitrogen is unavailable in the gaseous N2 form and must be first converted (or fixed) to become 

bioavailable. One of the most studied enzymes (nitrogenase) is responsible for this conversion 

of N2 gas to NH4

+

, and is possessed by only select prokaryotic organisms including bacteria and 

cyanobacteria. Both of these groups are present 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 expression of nitrogenase enzymes including species 

composition, light availability, availability of P and metallic co-factors (Fe, Mo), oxygen levels, 

and NH4

+

 concentrations. In the case of the springs, NH4

+

 levels are generally not high, and 

should not be sufficient to suppress nitrogenase. Metal availability (in particular trace elements 

required for nitrogenase function) may be limiting in some springs; however, there are little data 

to support this conclusion. The presence of cyanobacteria in spring systems indicates that Nfixers may have a competitive advantage for N acquisition. It is true that many springs are Nlimited, therefore, it is also likely that active N2 fixation may be occurring. This has been 

documented in a number of stream systems; however, no data exist to ascertain the significance 

of this process for Florida springs. Furthermore, it is also uncertain what the fate of this fixed N 

would be in the spring environment. 
 150 

Nitrogen uptake 

In streams and springs with high biomass and productivity, nitrogen required for growth that is 

not provided by nitrogen fixation must be supplied by N uptake and assimilation from the water 

column. High production of algae and SAV in Florida spring runs should also contribute to high 

N assimilation rates especially in areas where growth is N limited (Notestein et al., 2003; Inglett 

et al., 2007). Of the two inorganic dissolved N forms, both NO3

-

 and NH4

+

 are bioavailable and 

have been shown to be rapidly taken up by biota in headwater streams. Much of this headwater 

uptake is likely due to the high abundance of organic matter (leaf litter) in these systems. In one 

study of a low organic matter, brackish portion of the Chassahowitzka River, Saunders (2007) 

calculated rates of N loss (as NO3

-

) from the water column as high as 5 mg m-2 d-1, however; it 

was not determined what proportion of this N loss was due to biotic uptake, denitrification, or 

dilution. 

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

In springs it is unclear whether 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 may be equally important for productivity in SAV 

(compare Madsden and Cedergreen 2002 and Xie et al. 2005). Both NO3

-

 and NH4

+

 are 

considered bioavailable; however, NO3

-

 must first be reduced to NH4

+

 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 NO3

-

 in vegetation (Melzer and Kaiser 1986). 

At least one study has used this enzyme activity to assess potential NO3

-

 use by biota in springs 
 151 

with the findings that 1) higher water column NO3

-

 levels resulted in higher levels of NO3

-

reductase activity and 2) enzyme expression in biofilms differed depending 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 significant 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 presence of organic forms of N (in particular, urea) 

has been shown to effect algal species composition (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 (μmol NO2

-

 produced g-1dw h-1) in sediment and algal 

biofilms from various substrates (S, sediment; FB, filamentous benthic algae; BF, brown 

filamentous; E, epiphytic algae; FG, floating green algae; RW, rock wall biofilm; D, detrital 

material; SB, submerged bark biofilm) at four locations along the Wekiva River. (from Inglett 

et al. 2007). 

Nitrogen mineralization 

The biological transformation of organically combined nitrogen 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 process 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 enzymatic processes (e.g. cellulose, cellulobioase, 
 152 

glucosidase, aminopeptidase, protease etc.). Therefore, similar to C decomposition, the 

regulators of organic nitrogen mineralization include substrate quality, microbial biomass and 

enzyme activity, temperature, redox potential and hydrology. In a spring run, sites of 

ammonification would coexist with sites of organic 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, mineralized ammonium may be released, rapidly 

utilized by the microorganisms (immobilized), or taken up by plants. During decomposition, the 

C:N ratio reflects the mineralization-immobilization pattern of nitrogen, where a high C:N ratio 

promotes immobilization, while a lower ratio favors 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 decomposition rates. Thus, organic matter with a 

low C:N ratio (such as algae or SAV) is considered 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 NH4

+

 in the water column in a spring can be highly 

variable. 

Ammonium in the water column is derived from (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 biofilms; and (4) diffusion and mass flow of ammonium 

from anaerobic soil layer to aerobic soil/sediment layer and water column. Saunders (2007) 

found that sediments in the Chassahowitzka River were reducing at depth, while Stevenson et al. 

(2007) measured porewater NH4

+

 levels >10 mg L-1 under mats of Vaucheria sp. Both studies 

indicate that sediments in a spring run system can 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 mixing at and near the soil-floodwater interface. 

Once NH4

+

 enters the water column, a variety of processes can affect its fate in a spring 

run/system. Ammonium is bioavailable, and therefore can be rapidly assimilated by 

phytoplankton, biofilm organisms, or SAV. Alternatively, if pH is high (>8) a significant 

proportion of NH4

+

 will be in the form of gaseous NH3 which can volatilize (flux out of the water 

column). Volatilization of NH3 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 NH4

+

 in the well-oxygenated water column of a spring run will be 

nitrification where microorganisms oxidize NH4

+

 to NO3

-

. Autotrophic bacteria are dominantly 

responsible for the process of nitrification, and therefore these organisms do not require organic 

matter to function. Nitrifying populations of bacteria are likely to be found in all components of 

spring systems, and should be particularly concentrated at the aerobic/anaerobic interfaces where 

NH4

+

 and O2 will both be available. 

Denitrification 

Nitrate has a high oxidation potential (similar to oxygen) and can be used by a variety of 

organism groups as an electron acceptor (becoming chemically reduced). One of the most 
 153 

significant and highly studied processes reducing nitrate is the process of denitrification where 

organisms couple organic C oxidation to nitrate reduction. This processes is almost exclusively 

anaerobic (a few aerobic examples 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 NO3

-

 production (nitrification). 

Denitrifying organisms are predominantly heterotrophic. Therefore high organic matter C 

deposits (sediments, debris dams, biofilms, and algal 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 NO3

-

 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 biofilm growths possess enough C and can sufficiently 

reduce water flow to induce anaerobic conditions either spatially (e.g., in microsites) and/or 

temporally (short-term transient events or diel cycles). This would be particularly important at 

night in the absence of photosynthesis. 

To date, the only reported measurements of denitrification potential 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 N2O 

produced gdw-1 h-1 in surface sediments (0-10 cm) to 170 nmol N2O produced g-1dw h-1 in 

biofilms collected from bark samples from the headspring. The rates measured in the 

Wekiwa/Wekiva system are similar to DEA rates reported in wetlands (0.5 to 20 mg N2O kg-1

soil h-1; White and Reddy 2003), estuarine sediments (15-130 nmol N cm-3d-1; Joye et al. 1996), 

and in periphyton mats (1.9 μmol N2O m-2 h-1; Triska and Oremland, 1981). Based on their 

results, Inglett et al. (2007) concluded that despite a conclusive trend in DEA downstream (along 

a depletion gradient of surface water NO3

-

) denitrification appeared to be a significant pathway 

for N removal from the system. 

Alternative N pathways 

Conversion of NO3

-

 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 processes to account for significant reduction of NO3

- in 

aquatic systems. For this reason, we will discuss them briefly here. 

The high energetic potential of using NO3

-

 as an electron acceptor is not solely possessed by 

heterotrophic bacteria. Several pathways have also been identified that can reduce NO3

-

 through 

the oxidation of other reduced compounds. Identification of these new mechanisms has 

challenged the classical view of denitrification in aquatic systems. Among these, NO3

-

 reduction 

can be coupled to the oxidation of either reduced iron (Fe(II)) or sulfur (S0

, S2-). The importance 
 154 

of these lithotrophic pathways to overall rates of denitrification has not yet been determined. 

Likewise, in springs and lotic systems, Fe(II) and S2--based denitrification has received little 

attention. It is not unlikely that these pathways could be significant given the presence of iron 

and sulfur in many spring systems and the myriad of reactions occurring at the oxic/anoxic 

interfaces of microbial mats, biofilms, or sediments. 

Another pathway for NO3

-

 reduction can convert NO3

-

 into NH4

+

. This process, called 

dissimilatory nitrate reduction to ammonium (DNRA), is different from the assimilatory process 

in that NH4

+

 produced is not directly incorporated into biomass. Fermentative DNRA couples 

organic matter oxidation to NO3

-

 reduction by fermentation reactions (Tiedje 1988). 

Figure 7 – Denitrifying potential in sediments, and algal biofilms collected from various 

substrata. S, sediment; FB, filamentous benthic 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 NO3

-

 reduction to oxidation of 

reduced S (e.g., S2-). 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 activity in springs; however, these areas would 

be highly localized to accumulations of organic matter and/or slow or no flow conditions. 

Most recently, a novel pathway for N conversion to N2 has been identified which adds even more 

uncertainty to the aquatic N cycle. The process known as anaerobic ammonium oxidation 

(Anammox) has now been identified in a number of freshwater systems including lakes and 
 155 

wetlands. This process occurs in strictly anaerobic conditions, and couples the reduction of NO2

-

with the oxidation of NH4

+

 to produce N2 gas. Annamox bacteria appear to be autotrophic, and 

the presence of organic matter slows the annamox process. Another important requirement for 

the anammox process is the presence of NO2

-

 which is rarely observed in most systems. 

Figure 8 – Hypothesized controls on predominant dissimilatory pathways of NO3

-

 removal. C 

inputs refer to labile organic carbon available to microbes. Sulfidic refers to the presence of 

either free sulfide (S2–), elemental S (S0). 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 = denitrification; DNRA = dissimilatory nitrate reduction 

to ammonium; ferment = fermentative. (From Burgin and Hamilton 2007). 

Nitrite accumulation is possible however, and several conditions of spring systems may be 

adequate for NO2

-

 accumulation to occur. The formation of NO2

-

 is largely a byproduct of 

incomplete denitrification or partial ammonium oxidation. For NO2

-

 accumulation to occur in 

the denitrification and DNRA processes, NO3

-

 reduction must occur at a greater rate than that of 

NO2

-

 reduction. One way in which incomplete denitrification can occur is the lack of suitable 

organic substrates, or high concentrations of NO3

-

. Similarly, in the nitrification pathway, 

elevation of NO2

-

 concentrations can occur only if the oxidation of NH4

+

 is greater than that of 

NO2

-

. Both of these scenarios have been reported to occur in the aquatic environment (Herbert 

1982; Smith et al. 1997). Nitrobacter spp., which are responsible for the oxidation of NO2

-

 in 
 156 

the nitrification process, may be partially inhibited by free ammonia (NH3) (Anthonsieu et al. 

1976; Smith et al. 1997), leading to NO2

-

 accumulation. Likewise, NO2

-

 accumulation is a 

common trait of DNRA (Cole 1988) resulting from either inhibitory effects of NO3

-

 on NO2

-

reductase (Smith, 1982) or repression of this enzyme (Paul and Beauchamp 1989). 

Higher than expected NO2

-

 levels (100-200 μg N l-1) have been reported in some European 

rivers, particularly in watersheds receiving large inputs of nitrogenous pollution (Smith et al. 

1995; Kelso et al. 1997). In one study, Kelso et al. (1997) found that NO2

-

 accumulation in six 

Irish rivers corresponded with NO3

-

 declines via DNRA activity. This process was occurring in 

areas with high organic matter contents (Kelso et al. 1999), leading these authors to conclude 

that high NO3

-

 concentrations, free ammonia, or high flow velocity created conditions suitable 

for NO2

-

 accumulation to occur. All of these studies suggest NO2

-

 may be much more abundant 

than previously thought. Therefore, it is likely that NO2

-

 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 declining inorganic N concentrations (primarily NO3

-

) 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 result in a transformation of N, including conversion 

of DIN to DON through biotic uptake and senescence. 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 that DON production in spring runs is likely the 

result of excess N, where Wekiva and Rock Springs Run (both NO3

-

-impacted) show much 

higher production of DON in specific river segments, while Alexander and Juniper spring runs 

(reference conditions) show smaller production (or even net reduction) of DON. 

It is also quite likely that springs are active zones of denitrification. Evidence of denitrification 

in springs comes from at least one study (Inglett et al. 2007), and the conditions exist for other 

unexplored areas where denitrification could occur (Saunders 2007). Areas with direct inputs of 

reduced groundwater into sediments with low organic matter content, or water column NO3

-

being advected into anaerobic sediments, are also key conditions for alternative denitrification 

pathways to occur in springs (Grimaldi and Chaplot 2000). In systems with significant quantities 

of iron or sulfur, additional lithotrophic N loss mechanisms may 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 examples of declining NO3

-

 to either the process of biotic uptake or 

denitrification; however, this question is of critical importance for management of springs as 

nutrient sources. 

Phosphorus 

Phosphorus (P) is one of the major 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 landscape 
 157 

level, streams and wetlands form a critical interface between uplands and adjacent water bodies, 

as all of these ecosystems are hydrologically linked. Water and associated nutrients (such as P) 

are transported from uplands either by subsurface or surface flow. An idealized P cycle for 

spring systems is shown in Figure 9. 

Both biotic and abiotic mechanisms regulate relative 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 depend on the physical, chemical, and biological 

characteristics of the systems (Meyer and Likens 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, biofilm/periphyton and sediment microorganisms. 

Abiotic processes include sedimentation, adsorption by sediments and organic/inorganic 

substrates, precipitation, and exchange processes between sediment and the overlying water 

column. 

Figure 9 – Idealized diagram of the phosphorus cycle occurring in a typical spring run system. 

PIP=particulate inorganic P, POP=particulate 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 use 
 158 

characteristics of the aquifer, or springshed. To trace the transport and transformations of P, it is 

convenient to classify forms of P entering these systems 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 inorganic 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 based 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- and 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 sediments of south Florida (Reddy et al. 

1995a). In this example, the inorganic P extracted with neutral salt such as KCl represents 

loosely absorbed P, which usually accounts for less than 2% of the total P. This pool of P is 

considered bioavailable as it is sufficiently labile to enter porewater for uptake of plants or 

potential flux into the overlying water column. 

The NaOH-Pi represents inorganic P associated with Fe and Al, and represents P not readily 

available. The HCl-Pi represents P associated with Ca/Mg, is also relatively stable and not 

readily available. Alkali extractable organic P includes both readily 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 redox-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 yield insights on instantaneous bioavailability, but 

clearly sorption and desorption reactions can shift the solubilities and availabilities rapidly. 

These exchange reactions are markedly influenced by microbially-induced 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 proportion of organic P in sediments 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, and as monoesters of a variety of compounds, 

particularly those involved in biochemical pathways. In microorganisms and algae, more than 

half of the organic P is in nucleic acids, whereas phospholipids and monoesters constitute the 

remainder in varying proportions (Magid et al. 1996). Many organisms can store orthophosphate 
 159 

or polyphosphates (Cembella et al. 1983, 1984). In plants, inositol 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. Dissolution of inorganic substrates such as minerals 

can potentially provide some essential nutrients to algae and microbes colonizing on them. 

Although relatively little of the total P pool within actively growing macrophytes is released, this 

released P can be important to certain epiphytic 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 demands (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 water 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 microbial biomass and the organismal dynamics, 

particularly the bacteria, as they influence organic P dynamics. Microbial metabolism influences 

rates of mineralization of organic P in two primary ways: (1) by direct biochemical mobilization 

using extracellular or periplasmic 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 matter, 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 requires enzymatic cleavage of the ester linkage 

joining the inorganic P group to the organic moiety. Such hydrolysis is achieved by 

phosphomonoesterases (phosphatases) bound to or within the cell membrane or by dissolved 

enzymes released extracellularly within the interstitial solution adjacent to the cells. There is a 

general lack of specificity for the organic moiety that include, 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 inorganic P-irrepressible 

enzymes, which tend to be intracellular and function in intermediary metabolism, or are 

phosphate ester-induced inorganic 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 exogenous organic P compounds through enzymatic 

hydrolysis of terminal phosphate groups. A broad 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 solubilized 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 correlation has been found between bacterial biomass and phosphatase 

activity in surficial lake sediments (Reichardt 1978). Correlations between enzyme activities and 

bacterial biomass were less robust several centimeters below the sediment-water interface. A 

higher affinity of phosphatases was found in sediments where a low mobility of sediment P 

occurred (Pettersson unpublished; cited in Boström et al. 1982). The lowest substrate affinity 

(highest half-saturation constants) was found in organic sediments with low chemical adsorption 

capacities that release P under anaerobic conditions. Low availability of P in sediments often 

induces production of phosphatase activity with high affinity for substrates. 

Biotic and Abiotic Ca Interactions 

In calcium dominated streams, the presence of high Ca2+ can result in formation of complex 

calcium phosphate compounds of varying solubilities such as calcium phosphate, dicalcium 

phosphate, beta-tricalcium phosphate, octacalcium phosphate, and hydroxyapatite. Under these 

conditions, the phosphorus concentration in the pore water of calcareous sediments is a function 

of Ca2+ activity. Solubility of these compounds decreases with an increase in Ca content. 

Insoluble beta-tricalcium phosphate is more likely to found at a high pH. Thermodynamically, 

apatite is the most stable compound, while at relatively 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 phosphorus 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 major role in regulating P concentrations of the water column. 

Algae can assimilate both organic and inorganic forms of P (Bentzen et al. 1992), and can induce 

marked changes in pH and dissolved oxygen concentration 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, especially in streams. Some studies (e.g., Hartley et al. 

1997) have demonstrated that this process can result in significant co-precipitation of phosphorus 

with calcite in the presence of photosynthesizing algae, like Chlorococcum sp. 

Retention of inorganic phosphorus by precipitation will be significant in spring waters with high 

Ca2+ and alkalinity. In calcium-dominated streams (e.g., chalk streams) a significant portion of 

water column phosphorus can co-precipitate with CaCO3 where dissolved Ca2+ is available. 

House (1990) attributed only 6% of the overall phosphorus 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 phosphorus transported in the River 

Jordan. Several studies showed inhibition of calcite growth in the presence of high phosphorus 

concentrations (Avnimelech 1980; Dove and Hochella 1993) while at low phosphorus 

concentrations, coprecipitation 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 (Fe(II)), and adsorbed phosphate to be mobilized 

into soil pore water (Figure 9). Such conditions are of particular importance in areas where 

organic matter (and reducing conditions) can be transient, or when ground water varies 

temporally between aerobic and anaerobic states. Under strongly reducing conditions where 

sulfate is reduced to hydrogen sulfide, some S2- can react with iron compounds (both Fe(III) and 

Fe(II)) resulting in the precipitation of Fe-S compounds. In this manner, sulfide formation 

coupled to sulfate reduction can displace iron-bound P to enhance 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 anthropogenic sources, particularly in 

oligotrophic, softwater lakes and rivers. In highly reduced soils with low sulfate content, 

microbial reduction generates mostly Fe(II) compounds from microbial reduction of Fe(III) 

oxide. PO4

-3 can be retained with Fe(II) compounds, but Fe-associated PO4-3 is released when 

amorphous Fe(III) oxide is converted to iron sulfides (Roden and Edmonds 1997). Thus, the 

conversion of reactive soil Fe compounds to iron sulfides by sulfate-reducing bacteria can lead to 

a more efficient release of Fe-associated PO4

-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 the 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 solubilization 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 (particularly ion exchange), oxygen 

and other electron acceptor-dependent redox interactions, and the physiological and behavioral 

activities of many biota from bacteria, algae to other living communities. Exchange rates 

between various deposits of P and the interstitial water of the sediments depend on local 

adsorption and diffusion coefficients and their alteration by enzyme-mediated reactions of the 

microbiota. The sink/source function of springs could be transient, 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 mostly as protein/AA-S), sulfide (S2-), and sulfate (SO4

2-). 

Sulfide Oxidation 

Emerging groundwater potentially contains quantities of S2- that become oxidized when 

contacting atmospheric and photosynthetically-produced O2 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 (lithoautotrophy), or respire organic C (heterotrophy). 

Data from several springs indicate this process may be occurring. Sulfide is rarely determined in 

routine monitoring, but some springs show post-boil increases in SO4

2- (Wekiva, e.g.; Wetlands 

Solutions, Inc. (2005)) indicating that S2- oxidation could be occurring. 

 

Springs with high SO4

2- content may also host this process, but it likely occurs 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 surfaces of plants, interspersed with other 

filamentous algae, or on decomposing organic matter. Gradient culture techniques revealed that 

these S-oxidizing genera were culturable even at the lowest S levels tested (55 μmol 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 substrates (dissimilatory sulfate reduction). This 

process where SO4

2- becomes reduced to S2- only occurs under highly reducing conditions with 

redox potentials less than 200 mV. For this reason, measurable sulfate reduction is likely only 

possible in slow-moving spring runs or areas with high organic matter sediments (algal mats, 

SAV beds, litter accumulations). 

Sulfate reducers are obligate anaerobes that couple oxidation of organic substrates to carbon 

dioxide with the reduction of the terminal electron acceptor sulfate to sulfides (Widdel 1998). 

Sulfate reducing bacteria cannot synthesize hydrolytic enzymes, 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 monomers to simple readily utilizable organic 

compounds. 

Figure 11 – Idealized diagram of the sulfur cycle occurring 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 (S2- oxidized post boil). With 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 H2S. 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 predict that addition of nutrients to a nutrient-limited system 

will enhance productivity and lead to greater overall 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 dominance and growth (even if not 

resulting in an overall increased productivity) would still reflect a change in the quality of 

organic matter in the system and could have cascading effects through other heterotrophic 

communities both locally and further downstream. 

Other aspects of the C cycle in springs could be altered through the addition of C. Increased C 

inputs could occur through increased sewage/pollutant discharges which would act to raise the 

concentration of solids C or DOC. Management of exotics/invasives could also have profound 

impacts on nutrient dynamics through increased and episodic deposition of C following herbicide 

application. The effects of large discharges of C into streams are well known and most likely the 

same effects apply to springs, spring runs and springsheds. In particular, 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 denitrification), or P (increased release of iron-bound P). 

Nitrogen Impacts 

In the natural state, most Florida springs contain high levels of inorganic P, and based on N:P 

ratios, are N-limited. In the N-limited state, springs would be considered highly susceptible to N 

pollution. One of the key responses to N in these cases would be increased biomass or shifts in 

dominant species of all groups, including algae, macrophytes, fungi and bacteria. One example 

of this would be that under N-limiting conditions there should be a dominance of particular Nfixing microbiota. With the shifts in species composition, changes in C abundance and quality 

can also be expected. For example, increased algal production would result 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 increased dramatically, in 

Florida’s major springs during the last 30 years (Scott et al, 2004). With this trend, many springs 
 165 

are increasingly becoming P limited (Figure 12) (Inglett et al. 2007; Frazer et al. 2001). Aside 

from effects of incomplete denitrification (producing NO2

-

 and potentially N2O), it is currently 

uncertain if high levels of NO3

-

 can have any other adverse impacts to biogeochemical processes. 

Phosphorus Impacts 

In general, many springs already have high amounts of bioavailable 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 NO3

-

 pollution) (Figure 12). As in other aquatic systems, 

elevated P levels should result in higher overall productivity, or if limited by some other nutrient 

(or light) a higher potential productivity level. For 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 increased productivity of certain spring species; however, in this 

manner, high-P springs may demonstrate a greater 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 levels of N. Fluctuations in 

water column P levels will determine if the sediments act as a sink or source, as low water 
 166 

column P levels will favor flux of sedimentary P. The response of some systems to additions of 

P has been observed as a decrease in macrophytes chlorophyll levels. 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 the balance of aerobic/anaerobic conditions, and thus, would also 

effect the cycling of redox-sensitive elements such as N, P, Fe, and S. 

Sulfur Impacts 

It is uncertain what role S may be playing in the biogeochemistry of Florida springs. Sulfur can 

be significant as an electron acceptor (as SO4

2-) with a toxic byproduct of S2-. Also, S2- can 

become a powerful electron donor for particular organism groups and can diversify the pathways 

for NO3

-

 removal through denitrification. 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 groundwater 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 contrasting water high in S (e.g., that observed when tributaries 

meet). 

 SUMMARY AND CONCLUSIONS 

Based on the above discussion, springs are clearly dynamic systems with a myriad of potential 

biogeochemical processes/reactions and interactions among elemental cycles. Although the 

majority of these basic biogeochemical processes have been identified in other aquatic systems, 

many may not be universally applied to spring systems because of their unique features including 

karst-derived water chemistry, and relatively constant temperature and flow conditions. A 

diversity of conditions within spring systems also gives rise to numerous microbially active 

zones where conditions may exist for as yet unexplored processes and pathways. 

Nutrient cycles in spring systems are dominated by the conversion of inorganic to organic forms. 

Organic matter (dissolved and particulate 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 biogeochemical reactions, and is thus an important 

determinant in the cycling of other nutrients including nitrogen, phosphorus, and metals. Despite 

this importance, organic matter distributions, concentrations, 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 management practices can be a contribution of organic 

matter deposition in spring runs (affecting internal nutrient cycling); 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 additional to external loads of N through NO3

-

, nitrogen fixation is an 

important processes contributing to N, but is also currently unstudied in Florida springs. There 

are several pathways which can act to convert NO3

-

 (reducing its concentration), with the 

dominant most likely being heterotrophic denitrification or plant assimilation. The importance 
 167 

of these two pathways will largely determine if a spring is a sink or source of N; however, 

current studies have not attempted to partition N flow through each pathway. High flow rates 

and patchy distributions of organic matter also indicate that conditions in many springs could 

support several alternative (novel) pathways of NO3

-

 reduction, but again, these have received no 

attention. 

Phosphorus may also play a key role in regulating species composition or productivity in a 

spring system. A variety of biotic and abiotic processes act to determine the biogeochemical 

cycling of phosphorus in spring systems. The generally high calcium content 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 regulating 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 determine both its form and 

solubility/mobility and its particular interaction with PO4

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 processes in alternating 

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 NO3

-

 levels in some aquifers and in some spring runs. 

Management and Policy Implications 

Nutrient enrichment of Florida springs is major concern. The steady increase in nutrient 

concentrations observed in many Florida springs 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 watershed 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 for management of spring systems. Nutrient 

limitation and biogeochemical processes are a very interactive 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 particular, microbial communities will likely 

demonstrate responses to changes in nutrient conditions immediately through changes in species 

or functional groups and through activities of enzymes 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 long 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 their 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 strategies 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 supply from sediments, it is important to document 

storages and fluxes of nutrients between sediment, vegetation, and water column. At this time 

little or no information is available on internal cycling of nutrients and its overall role in 

regulating eutrophication process. When considering internal cycling of nutrients in springs, 

several issues need be addressed. These issues are: (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 water column to maintain eutrophic 

conditions, once external loads are reduced? Thus, monitoring internal 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 aspects of spring systems. The complexity of 

biogeochemical pathways, the complexity of the high energy/lotic spring environment, and the 

diversity of spring water types makes it difficult to form any wide-ranging generalizations. For 

this reason, biogeochemistry is also one of the 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 characterization of spring systems for biogeochemical purposes. 

Characterization data and long term measurements of water chemistry are available; however, 

the addition of some parameters to this monitoring are warranted, including sulfide (S2

) and 

reduced iron (Fe(II)). Also, there 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 matter contents and qualities, and stratigraphic 

morphology. For this reason, data should also be collected to characterize the biogeochemical 

properties of spring run sediments including basic nutrient chemistry, mineralogy, and potential 
 169 

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 function 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 (NH4

+

, NO3

-

, DON), and the occurrence, rates and importance of biological N2-fixation. 

Whole spring run N budgets should be performed to better isolate the processes responsible for 

N decline, and these data put in the context of nutrient limitation of systems of differing geology, 

channel morphology, water chemistry, and biotic productivity. Given the potential for 

structuring aquatic microbial communities, the presence and effects of potential DON sources 

(including urea) should be investigated in relation to algal composition, microbial activity, and 

community nutrient limitation. Potential pathways of NO3

-

 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)/S2- 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. Determination of system N uptake can be done in small batch 

or whole stream segments (Saunders 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 N2 fixation assays 

(Inglett et al. 2004). It is well known that natural abundance level stable N isotopic methods can 

be used to assess sources of N in spring systems; however, the analysis of δ15N can also provide 

information on rates of N uptake or denitrification (De Brabandere et al. 2007). Experiments 

utilizing tracer levels of 15N can also provide useful identification of N pathways (e.g., novel 

NO3

-

 reduction pathways) as well as quantitative information of pathways and flows (e.g., N2

fixation, N uptake, and denitrification). 

Carbon Dynamics 

 Carbon sources and distribution play a key role in structuring the biogeochemistry of the spring 

system, therefore better understanding is needed regarding the sources and processing of C in 

spring systems. This would include identification and quantification of autochthonous and 

allochthonous C sources (quantities and qualities), availability and cycling of different C types, 

and investigation of the importance of different C sources for respiration processes and 

denitrification. The role of aquatic plant management in the distribution and fate of C within 

particular systems is also of significant 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 where smell or taste issues 

are concerned; however, S may be a key element regulating several biogeochemical cycles 

including N and P. For this reason, more emphasis should be given to S as it relates to 

biogeochemical processes including 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 regulating N and P biogeochemistry, 

therefore, studies linking redox processes with denitrification or P flux from sediments are 

potentially useful for elucidating the functioning of S in some nutrient-impacted spring systems. 

Coupled Biogeochemical Cycles 

Identifying and understanding biogeochemical pathways 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 matter and biotic productivity, it can be seen that 

elements (and thus elemental cycles) are tied together (or coupled) so that the pathways of one 

cycle inherently affect the pathways of another element. 

 Currently, there have been no attempts to understand 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 biota (e.g., questions of N versus 

P limitation), but it could also have application to modeling the 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 assessing the potential for 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 sets (both current and 

future) should incorporate the concepts of coupled biogeochemical cycles when collecting data 

to assess spatial and temporal scales. 
 171 

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 179 

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 are enriched in nitrate due to broad-scale 

contamination of groundwater supplies. This observation 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 toxicity. In addition, high nitrate concentrations 

can promote eutrophication or an increase in organic matter, which drives a diverse array of 

responses and feedbacks in aquatic 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 delivery to aquatic systems 

favors production of fast-growing algae that ultimately outcompete and displace 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 consistent with such a 

progression and a cue that management 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) changes 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 systems. However, the potential for 

sublethal effects should be considered during discussions of management 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 and 

magnitude of any changes are determined by complex interactions among bottom-up and topdown processes, such as grazing pressure, habitat use and predation pressure. The outcomes of 

such interactions clearly relate to sustainable management of spring systems. 

The Literature: What is Not Known 

There are few quantitative data that definitively 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 from 
 180 

contemporary concentrations of nutrients in the water column. Currently, patterns in the 

distribution and abundance of plants 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 potential 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 floral 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 would benefit from compilation of inventories, 

collection of baseline data, and completion of diagnostic studies focused on bottom-up and topdown interactions. Ultimately, unforeseeable, emergent properties that arise when management 

is implemented at real-world, spatiotemporal scales will create a need to manage adaptively. 

Efficient and effective research to improve management of Florida’s springs and spring-fed 

systems requires input from multiple stakeholders. 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 drivers and valued components of spring systems 

chosen with the best available 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 temporal scales that support rigorous analyses 

of interactions (note that these inventories and baselines form the platform for 

adaptive management) 

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

• diagnostic studies that determine assimilation of nutrients by flora as key elements in 

nutrient budgets for spring systems, including contributions by epiphytes, macroalgae, 

vascular plants and microphytobenthos or microalgae found in sediments 

• diagnostic studies that determine how nutrients and other drivers affect overgrowth, 

shading and other relationships among periphyton, 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 
 181 

− extend beyond nitrogen and phosphorus to include 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, including uptake and other physiological 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 relationships form the core of an adaptive 

management approach) 

− elucidate the direction, magnitude and form of limitations (e.g., nonlinear and 

nonreversible) to identify suitable targets and goals for management actions 

• diagnostic studies to identify and elucidate changes in composition and function of faunal 

assemblages related to changes in habitats in an effort to identify habitats 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 consumers in an effort to identify flora, 

grazers and trophic links to be protected or restored by management actions 

• diagnostic studies to identify and elucidate the toxic effects 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 
 182 

INTRODUCTION 

Anthropogenic activities have increased the delivery of nutrients in aquatic systems of all types 

resulting in concerns about the biotic integrity of these systems. In fact, excessive growth of 

primary producers in water bodies due to high inputs of nutrients, especially 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 reported to threaten wadeable streams, 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 key 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 ecosystems 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 management responses. Nutrient standards and criteria 

have been set in response to: 1) adverse effects on humans and domestic animals, 2) aesthetic 

impairment, 3) interference with human use, 4) negative impacts on aquatic ecology, and 

5) excessive nutrient transport to downstream systems (Dodds and Welch 2000). These five 

concerns interact. For example, systems exhibiting negative ecological impacts probably fail to 

be aesthetically pleasing or support a full range 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 resulted in broad-scale nutrient contamination of 

groundwater (Cohen et al. Chapter 2). This nutrient-laden groundwater exists within a very 
 183 

permeable, karst setting, with myriad pathways between aquifers and surface 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 nutrients, such as nitrogen and 

phosphorus, are mixed into spring-fed, surface waters, they attain the greatest potential to 

negatively alter the ecology of aquatic ecosystems. The effects of increased nitrogen in the form 

of nitrate raise particular concerns for flora and fauna associated 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 contributes 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 algal growth creating undesirable aesthetic and 

ecological consequences. Such an increase in plant or algal growth is commonly referred to as 

eutrophication (c.f. Nixon 1995). The potential eutrophication of Florida’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 managers is that plant and algal 

populations in Florida’s spring-dominated ecosystems are undergoing major structural and 

functional changes, due, in large part, to increases in anthropogenic enrichment of nutrient levels 

in groundwater and the consequent nutrification of spring discharges. 

This widely held view is not without merit. There is, in fact, a growing body of literature on the 

chemistry of spring discharges that supports the conclusion that nutrient 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 anecdotal accounts and qualitative 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 nitrate delivery to spring systems or demonstrates 

that the purported changes have compromised the ecological health 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 relationships among nutrients and spring flora, the 

habitat associations of spring fauna, and the potential toxic effects of some nitrogen compounds. 

From this synthesis, we aim to draw guidance or recommendations for policy and management. 

Where sufficient information is lacking, our goal is to suggest potential remedies. 

PRIMARY PRODUCERS IN SPRING ECOSYSTEMS AND FACTORS AFFECTING 

THEIR ABUNDANCE AND DISTRIBUTION 

Florida’s springs often are perceived as stable-state ecosystems in which major chemical 

constituents, temperature and discharge characteristics of individual springs are essentially 

constant (Odum 1957b). In fact, there is substantial 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 within spring systems (Odum 1957b). Patterns in 
 184 

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 heterogeneous environments in which 

various chemical, physical and biological processes interact to determine ecological structure and 

function. Thus, the inherent complexities in spring-dominated ecosystems must be recognized, 

and any broad generalizations related 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 essential 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 other, naturally variable, chemical properties of groundwaters and 

surface waters influence the 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 influence 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 spring systems have been shown to vary with specific conductance and 

salinity (Whitford 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-scale or fine-scale patterns in plant and algal 

abundance. For example, water depth, light availability, stream velocity, flow regimes, and 

characteristics of substrates all represent important considerations (Butcher 1933, Hynes 1972). 

Beyond chemical and physical influences, biological 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 the 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 species 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 focus of this chapter and a logical complement to the discussion 

of direct impacts on spring-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 dissolved organic substances (Kirk 1994). Suspended solids 

include algal cells (Canfield and Hodgson 1983), nonvolatile suspended solids (Canfield and 

Bachmann 1981; Hoyer and Jones 1983) and detrital material (Buiteveld 1995). 

All of the “in-water” constituents that reduce light availability are generally in low 

concentrations in groundwater serving as the origin of flow for Florida’s springs and spring runs. 

For example, Florida springs with discharges of 5 m3

 s

-1 or higher (1st magnitude springs) had the 
 185 

lowest concentrations of dissolved organic matter yet reported (13 ± 1.6 μM C L-1), and springs 

with lower discharges exhibited only 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 limit 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 spring runs in interior central and northern Florida 

and concluded that shading by riparian vegetation was a primary determinant of primary 

productivity and macrophyte standing crops in these systems. It is interesting to note that these 

investigators reported negative rates for daily maximum production in several spring runs devoid 

of “vegetation,” and it was assumed that biochemical oxygen demand exceeded any production 

by benthic and/or pseudoplanktonic algae in these systems. It was further suggested that oxygen 

concentrations in groundwater emanating as spring discharge may, in some cases, be too low to 

support the respiratory demands of aquatic macrophytes (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 generally sufficient to support net photosynthesis 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 distribution of aquatic plants and 

algae. Correlations with the percentage of shoreline canopy cover 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 increased, the total vegetative biomass decreased, which led these authors 

to suggest that canopy cover in the Weeki Wachee 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 Odum suggested in his seminal work (Odum 1957b) 

that primary production in Silver Springs (and other spring systems 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 crucial element of the argument that nutrification of 

Florida’s spring systems has led to a proliferation 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 Duarte 1995 and also Valiela et al. 1997). At this time, there is no direct 

evidence that accumulations of periphyton on macrophytes 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 plants (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 (CO2) or active 

uptake of bicarbonate (HCO3

-

) is hindered by increased boundary layer effects created by a layer 
 186 

of algae. Thus, the effects of changes in light climate will be mediated by stream velocity and the 

supply of carbon, especially dissolved inorganic carbon (DIC). There is a need to quantify the 

relationships among light climate, DIC availability, 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 are naturally high due to microbially mediated 

remineralization of organic matter in the groundwater system, this situation typically applies 

(Inglett et al. Chapter 3). 

Many springs (particularly 1st magnitude springs) are supersaturated with DIC and there is, in 

fact, a net flux of gaseous CO2 from the springs to the atmosphere 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 primary 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 velocity) processes, however, may result 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 key chemical constituent of all systems that 

support biological communities. Any factor or combination of factors that reduces primary 

production in spring systems will necessarily reduce the availability of dissolved oxygen. The 

potential consequences of reduced 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 dissolved oxygen to support aerobic respiration, especially in 

belowground tissues. Reduced photosynthetic capacity compromises the supply of oxygen, and 

such a change may negatively affect plants by disrupting their carbon balance (see Hemminga 

and Duarte 2000). Oxygen deficits might be expected to occur where macrophytes are highly 

epiphytized or covered by extensive macroalgal growth because of increased respiration and/or 

boundary layer effects on diffusion. We reiterate, however, that there are no experimental studies 

clearly demonstrating a negative effect of epiphytes or macroalgae on vascular plants in 

Florida’s spring systems, although there are a number of empirical 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 hypoxia, one might expect rooted macrophytes to be 

more susceptible to the potential effects of algal overgrowth. It is interesting to note that 

Stevenson et al. (2004) reported a significant negative relationship between dissolved oxygen 

concentrations and algal cover based on a broad survey of springs in Florida. The nature of this 

relationship merits further investigation, especially in light of recent information that suggests a 

broad-scale decline in dissolved oxygen concentrations in Florida’s springs (Heffernan et al. 

unpublished data). 
 187 

Stream Velocity and Substrate 

Water movement can affect the gas exchange, physiology and growth form of plants and algae, 

particularly periphyton. In addition, stream velocity can determine 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 that are continually being 

rolled or scoured will have little submersed aquatic vegetation, while a river bed that is 

comprised largely of mud, silt and sand has the potential 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 systems 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 support all types of submersed aquatic vegetation. 

For example, exposed limestone is prevalent in the immediate vicinity of spring vents, and 

outcrops of limestone are distributed 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 algae 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, stream velocity will be a concern primarily because of its effects on 

the periphyton assemblages associated with macrophytes. Although submersed vascular plants 

and large macroalgae in springs 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 producer 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 consequence of climate variability or anthropogenic 

demands on source water (Odum 1957b). 

In systems with relatively high flow rates, periphyton associated with macrophytes or other 

substrates are comprised mostly of low-lying mucilaginous diatoms that are resistant to 

sloughing even at high velocities (Biggs et al. 1998). Long, filamentous algal species are much 

less resistant to high velocity environments. 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 experimental studies in Florida spring systems that 

quantify the effects of flow on periphyton biomass 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 periphyton associated with macrophytes in any of the five spring-fed, 

coastal rivers they investigated. In a subsequent study of these same systems, however, 

statistically significant decreases in periphyton associated with macrophytes in the Weeki 
 188 

Wachee River coincided with an increase in flow (Frazer et al. 2006b). Kurz et al. (2004) found 

no significant relationship between periphyton abundance on macrophytes and flow in the main 

stem of the Ichetucknee River. It should be noted, however, that periphyton abundance on 

macrophytes was significantly higher in several feeder springs where flow rates were 

significantly less (Kurz et al. 2004; see also Dormsjo 2007; Politano unpublished data). 

These findings are counter-intuitive because reduced 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 dissolved inorganic carbon) to the periphyton. 

When boundary layers are thick, as in areas with slow-moving water, nutrients and other 

chemicals must diffuse or be actively transported across large distances. 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 limited 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 reducing the concentration of metabolic by-products 

that might impair photosynthesis. In general, periphyton 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 availability 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 abundances and the particular physiologies of the taxa present in the 

system. In some situations, plants and algae may be limited by the availability of dissolved 

inorganic carbon, although there is little indication that this might be the case in any of Florida’s 

springs, except perhaps in algal mats. Due to their import, our focus here is on nitrogen and 

phosphorus and their potential to limit the production of submersed 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 largely from sediment porewaters (Chambers 

et al. 1989), although some research suggests that many plants (some of which are present in 

Florida’s spring systems) are capable of nutrient 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 vascular plant in a Florida, spring-fed system is 

limited by either nitrogen or phosphorus. Moreover, empirical data generated in a large number 

of spring systems suggests that there is no significant 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 rooted vegetation and macroalgae) was related 

to nutrient enrichment. Standing crops of macrophytes in these systems were not correlated with 

in-stream concentrations of total phosphorus or total nitrogen. Furthermore, in a subsequent 

investigation, Duarte and Canfield (1990b) sampled 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 relationship between nutrient concentrations in surface 

waters and characteristics of submersed aquatic vegetation in five spring-fed coastal rivers 

subjected to increasing nitrate delivery. They found that statistically significant, positive 

relationships between nitrate and standing crops of either macrophytes or filamentous 

macroalgae were evident only in the Chassahowitzka and Homosassa Rivers (Table 1). 

Mean nitrate concentrations declined markedly at downstream sites in both rivers, which is 

indicative of rapid nitrate removal. Subsequent research in the Chassahowitzka River suggests, 

however, that the vast majority of the nitrate removal occurs as a consequence 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 markedly 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 rivers, 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 substrates relative to nitrate amended treatments or 

controls (no nutrient addition). Frazer et al. (2001a) reported that soluble reactive phosphorus 

was correlated with macroalgal biomass in the Weeki 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 log10–transformed 

annual mean nutrient concentrations and log10–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 Correlation coefficient Probability

CHASSAHOWITZKA 

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

CRYSTAL 

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

 

CRYSTAL (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 0.11
 191 

Nutrient Biomass Correlation coefficient Probability

Total phosphorus Periphyton 0.01 0.97

HOMOSASSA 

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

WEEKI WACHEE 

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 statistically significant, positive relationships between nutrients 

and coverage or biomass of submersed aquatic vegetation 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 complementary analysis, these investigators related a 

suite of chemical and physical parameters to periphyton abundance on macrophytes in several 

small spring runs adjoining the main river. Periphyton abundance was not correlated with either 

nitrate or total nitrogen concentrations, but it was significantly correlated with soluble reactive 
 192 

phosphorus concentrations suggesting phosphorus limitation. 

The issue of nutrient limitation of algal production can be viewed at several levels of 

organization, from individual algal cells or filaments 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 concentrations 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 (Grimm 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 the growth rates of natural, benthic, algal mats suggests that nitrate 

concentrations currently observed in the water 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 impacted 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 (Bothwell 1985). Half saturation 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 increase in nitrogen levels in many Florida springs, 

phosphorus concentrations are reported 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 systems 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 exhibit 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 sediments may provide sufficient nutrients to sustain 

this biomass. This hypothesis is supported by observations of increasing phosphorus levels in the 

pore water of some benthic algal mats (Stevenson et al. 2007). The absence of sufficient 

information on nutrient dynamics within algal mat communities in spring ecosystems hampers 

our ability to predict the outcomes of changes in nutrient loads. 

Community Responses to Increased Nutrients 

Plant and algal communities appear to exhibit 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 the root of observed changes 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 Minimum). At low nutrient concentrations in the 

water column and under favorable light conditions, slow-growing, rooted vascular plants are 

superior competitors owing to their relatively 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 nutrient 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 overgrow 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-lived 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. The 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 continuous, gradual process that runs parallel to the rate of 

increased nutrient delivery. Rather, changes occur 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 concentrations of nutrients in the water column. 

Thus, indirect effects of nutrient 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. Increased algal respiration during blooms may create 

hypoxic or anoxic conditions that kill grazers, which decreases grazing pressure on the fastgrowing algae and creates a positive feedback in the eutrophication progression. 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 nutrient over-enrichment for vegetative communities in 

Florida’s various spring systems. 

One potentially key shift in the floral communities 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 hypothesized that heavy infestations of 

Lyngbya on the skin of manatees in springs may be related to recent observations of health 

problems (Bledsoe et al. 2006). From a human health perspective, 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 submerged aquatic vegetation communities following 

nutrient remediation and reduction. Recovery is likely to be a very slow process, in large part, 

because it involves the replacement of fast-growing species with relatively slow-growing 

species. Duarte (1995) employed a simulation model that predicted recovery times of centuries 
 194 

for systems originally occupied by slow-growing seagrasses. It is worth noting here that many of 

the historically dominant vascular plants in Florida’s spring systems (e.g., Vallisneria americana

and Sagittaria kurziana) share morphologies and life histories 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 habitats for fauna that are subject to stress by 

increased nutrients. Key elements of spring systems 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 conducted 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, the focus falls on benthic organisms 

and trophic links. The base of trophic webs in flowing systems includes allocthonous material, or 

external inputs such as leaves from terrestrial vegetation, and autocthonous material that is 

produced in the system. Autocthonous material becomes more important as shading by riparian 

vegetation decreases and in-stream primary producers photosynthesize more effectively. Key 

primary producers in flowing systems include macroalgae, vascular macrophytes and periphyton, 

with these algal species growing on almost all biotic and abiotic surfaces. Primary producers 

supply food to primary consumers dominated numerically by a variety of invertebrates, 

especially snails and larvae of aquatic insects. 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, invertebrates 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 collect fine particles of organic detritus, filterers 

of suspended particulate organic matter, and predators feeding on live animal prey. Similarly, 

fish in the Wekiva River spanned the expected trophic roles, including detritivores, herbivores 

that often feed on algae, planktivores, invertivores 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 (Walsh 2001; Shelton 2005). Nearly one third of 

stygobites have recorded distributions limited to single cave systems, and 67% of these taxa have 

been found in 10 systems or less. Therefore, a high percentage of these species are listed as of 

special concern, rare, threatened or endangered. Obligate fauna of springs include hydrobiid and 

pleurocerid snails with highly restricted 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 single 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 primary 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 flora found in spring systems. In turn, changes in 

vascular plants, attached macroalgae and mats of filamentous algae can alter the food and refuge 

available to invertebrates in Florida springs. However, few studies have assessed invertebrate 

assemblages in different types of vegetation to obtain a synoptic view. 

A survey of the Homossassa and Weeki Wachee Rivers found differences in the number of 

species and relative densities of certain insects along the length of the waterways (Sloan 1956). 

Habitats were classified according to the areal coverage of dominant 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 Weeki Wachee River. In 

general, the insect assemblages appeared to vary primarily according to concentrations of 

dissolved oxygen near spring boils and salinity near the estuarine endpoints of the rivers. 

In Sulphur Springs Run, quantitative samples taken in bare sediment and algal mats with a petite 

Ponar dredge showed no significant differences 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 and 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 methods were not assessed. Therefore, the data support general 

comparisons of faunal assemblages among habitats, but detailed quantitative comparisons may 

prove unreliable. 

The relative abundance of different functional types varied among habitats (Figure 3; Warren 

et al. 2000). Gatherers were found in all habitats, and they dominated bare 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 abundances. 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 systems, represented less than 1% of the organisms 

sampled. Overall, this survey indicated that altering the areal coverage and species compositions 

of habitats might lead to changes 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 lotic systems are used as indicators of ecological health (e.g., 

USEPA 2006). For example, Florida has two biotic indices, the Stream Condition Index (SCI) 

that relies on laboratory processing of samples and the more rapid Biological Reconnaissance 

(BioRecon) that uses sorting 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 relative abundances or 

percentage dominance of selected taxa. The concept that invertebrate taxa exhibit some 

consistency in their response to disturbances underpins biotic indices. In Florida and elsewhere, 

biotic indices were developed empirically using samples of invertebrates collected from locales 

chosen to represent different levels 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 differences or changes typically are derived from 

correlations with auxiliary data on water quality, habitat quality, or other putative forcing factors. 
 198 

Interpreting correlations between biotic indices and forcing factors may be less than 

straightforward. For example, over a 14-year period, invertebrate abundance 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 were negatively correlated with 

nitrate concentrations (Hornsby et al. 2004). These results were interpreted 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 Florida 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 Condition Index (SCI) showed that SCI values 

were not significantly correlated with nitrate–nitrite concentrations or habitat assessment scores 

(Table 2; Figure 4). In nine springs with very poor to poor SCI values, 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 concentrations and habitat assessment scores accounted for less than 1% and 16% 

of the variation in the data, respectively. In contrast, SCI values were significantly correlated 

with dissolved oxygen concentrations, with a positive correlation that accounted for 36% of the 

variation in the data (Table 2; Figure 4). Such results illustrate the complexities associated with 

interpreting biotic indices, and they provide little 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 NOx NH3 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 F 
 199 

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 Index

Nitrate-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 Index

Habitat 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 Index

Dissolved oxygen (mg L-1)

Figure 4 – Correlations between Stream Condition Indices (SCI) and (a) nitrate–nitrite 

concentrations, (b) habitat assessment scores and (c) concentrations 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 = fair SCI values; lines provided 

to illustrate the correlations
 200 

Relationships Between Nutrients and Fish in Florida Springs 

By affecting the flora of springs and other aquatic systems, nutrients also can affect fish that rely 

on this flora for food and shelter, including fish 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 adjacent 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 examined 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 reduced numbers of mullet and catfish ascribed to 

disrupted migratory patterns following 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 with 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. Regardless, higher standing stocks and biomasses were 

found in reaches with an open forest canopy, which coincided with reaches where macrophytes 

were more abundant. 

Detailed data on fish assemblages in specific habitats 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 vegetated 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 large numbers of small minnows in 

the family Cyprinidae and killifishes in the family 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, broad comparisons between the two studies of the 

Wekiva River indicated that the fish assemblage had changed little 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 concentrations 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 for blocknet samples (Warren et al. 2000).
 207 

Summary of Relationships Between Nutrients and Fauna in Florida Springs 

In summary, surveys of faunal assemblages in Florida springs have suggested the possibility of 

indirect influences from nutrients mediated through differences or changes in habitats. However, 

results of repeated surveys indicate faunal assemblages have not changed even if nitrate levels 

have risen. Unfortunately, these studies were not designed to identify the form and magnitude of 

interactions among nutrient inputs, primary producers and fauna. 

Studies conducted in streams outside Florida and in lakes can provide additional insights into 

how nutrients affect habitat use and trophic relationships. 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 Kornijόw 1998). For example, 

fish may rely heavily on macrophytes for refuge from predation in the early portion of their life 

histories even if it reduces their growth rate (Figure 12). A nonlinear change in growth rate may 

be associated with changing habitat use after achieving a size refuge (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 optimal habitat once it has 

achieved a size refuge (adapted from Diehl and Kornijόw 1998). Solid line represents growth 

rate exhibited by a fish that switches from one habitat to another once it achieves a size refuge 

from piscivores at body size S. Dotted line represents the growth rate of a fish able to use all 

available habitats and food resources due to the 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 productivity of macrophytes, which 

provides more food for herbivores and may lead to increases in secondary 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 and macroalgae) have been regarded as a 

little used food source due to high C:N ratios (e.g., 20–1,340), large quantities of cellulose and 

lignin, decreased digestibility of their proteins and 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 reduce macrophyte biomass through consumption or 

damage during grazing on periphyton (Cummins and Klug 1979; Mann 1988; Lodge 1991; 

Lodge et al. 1998; Newman 1991). 

Regardless of their contribution as a source of food, macrophytes 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 has the potential to lead to loss of macrophytes, with subsequent 

effects on refuge and trophic interactions. 

Grazers can mediate the growth and effects of periphyton. Experiments using elevated tiles as a 

mean of excluding the herbivorous larvae of the caddisfly, Helicopsyche borealis, demonstrated 

effects of grazing on biomass and turnover of microalgae (Lamberti and Resh 1983). In grazed 

treatments, biomass, measured 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 experiments comparing grazers at ambient densities 

to treatments without grazers indicated that grazers significantly reduced periphyton biomass 

in 70% of experiments and measurably altered 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 the 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 polysaccharide sheath 

produced by Lyngbya wollei deterred grazing by the amphipod Hyallela azteca (Camacho and 

Thacker 2006). In streamside experimental troughs, herbivorous protozoa were shown to graze 

on benthic diatoms, with estimated consumption 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 systems 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 growing on them. Addition of phosphorus and phosphorus 

plus nitrogen to a stream led to a greater biomass of periphyton and larger larvae (Hershey 
 209 

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 periphyton, 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 grazing. Periphyton in streamside, flow-through channels exhibited 

statistically significant increases above controls only when nitrogen and phosphorus were added, 

light was increased, and grazing was eliminated (Rosemond 1993). In addition, manipulating 

single factors led to different responses. For example, increased light elicited both a 

physiological response resulting in decreased chlorophyll a concentrations in cells and a 

change in the periphyton assemblage when algae adapted to low light conditions were lost 

(Rosemond 1993). 

All the bottom-up influences of nutrients and light and top-down influences of grazing can be 

mediated through actual or potential trophic interactions. Manipulations of pools in a stream 

demonstrated that grazing minnows controlled periphyton growth unless predatory fish were 

present, in which case predation and avoidance of predation led to increased standing crops of 

periphyton in deeper sections of the stream (Power et al. 1985). 

In some cases, spring systems may exhibit very complex trophic cascades, which involve 

reciprocal interactions among predators 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 phytoplankton 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 productivity. 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 did not disrupt the trophic cascade. Piscivorous 

fishes reduced numbers of zooplanktivorous fishes to an extent that allowed zooplankton 

numbers to increase. The resulting increased grazing by zooplankton controlled the 

phytoplankton assemblage. 

Over time, the interactions among bottom-up and 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 lagging 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 phosphorus levels in streams led to the conclusion that increases in 

fish biomass with increased phosphorus were greater at lower initial phosphorus 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 loops, characterize one of the most 

cogent descriptions of eutrophication (Duarte 1995). In the eutrophication progression, 

interactions among nutrients, plants and algae, and grazers combine to produce changes that are 

not linearly correlated with nutrient 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 systems dominated by fast-growing 

macroalgae and microalgae and potentially to systems dominated by phytoplankton 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 producers, grazers and consumers, 

nutrient enrichment also may alter interactions between parasites and 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 physiology, 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 individuals. 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 LC50. Florida law contains a definition of a 96-hour LC50, which is a 

concentration that kills 50% of the test organisms over a 4-day period (Florida Administrative 

Code Chapter 62–302.200). Sublethal endpoints can also be used to yield concentrations that 

cause a defined effect in a defined time interval. Terms for such endpoints include least 

observable effect level (LOEL), which 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 observed. In addition, experiments can examine chronic toxicity or the 

effect of longer exposures to lower concentrations of toxicant; however, the time needed to 

conduct such experiments has led to various approaches to extrapolate results of acute tests. The 

results of laboratory experiments documenting effects on individuals typically are extrapolated to 

predict ecotoxicological effects, or effects at the population, assemblage or ecosystem levels. 

Effects at higher levels of organization may arise from density-mediated changes resulting from 

the death of organisms or trait-mediated changes resulting from sublethal 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 ecotoxicological effects may employ uncertainty factors to account for limitations 

imposed by testing relatively few individuals of relatively few species 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. The USEPA recommended an intraspecies uncertainty factor 
 211 

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 toxicokinetics (processes determining uptake rates, 

distribution among tissues, detoxification and elimination) and toxicodynamics (processes 

determining effects such as enzyme inhibition, enzyme facilitation and 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, known 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 equilibrium levels only after long-term exposure; 

physiological compensations to toxins may fail after longer periods of exposure; latent, adverse 

effects may appear after exposure ceases; and longer exposures may affect 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 concentrations in the tissues of the target organism. An 

uncertainty factor also can be used to predict NOELs from observed 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 the 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 systems, 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 more toxic than nitrate (Armstrong 1979). 

Toxicity of Ammonia 

The USEPA has invested significant effort in identifying water quality criteria for ammonia (e.g., 

USEPA 1999). Their results provide important guidance for establishing criteria related to 

concentrations of ammonia considered safe in aquatic environments. Some key findings from 

their analysis of data related to ammonia toxicity are: 

• toxicity of ammonia can depend on the ionic composition, pH and temperature of water 

due to poorly understood mechanisms, with pH dependence strongly suggesting joint 

toxicity of unionized ammonia and ammonium ions 

• ionic composition generates small effects on ammonia speciation in freshwater, and its 

other effects on toxicity are poorly established; 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 strictly by an empirical approach 
 212 

• pH effects were described by equations that include 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 acute toxicity data yielded criteria expected to protect aquatic 

life, except for unusually sensitive species. Criterion maximum concentrations (CMCs) drawn 

from acute toxicity tests address short-term exposure, and analysis of chronic toxicity data yields 

criterion continuous concentrations (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 Zealand would be protected 

minimally by the proposed criteria (Hickey et al. 1999). In Florida springs, the effects on 

obligate residents in springs and stygobiota represent special concerns. For example, a survey of 

stygobitic crayfish linked their presence or absence to ammonia concentrations in source water 

(Spanjer and Cipollini 2006). Regardless of the associated uncertainties, the proposed criteria are 

based on detailed interpretation of the best available 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 + 107.204 – pH + 1 + 10pH – 7.204 where salmonid fish are present 

0.411 58.4 CMC = 1 + 107.204 – pH + 1 + 10pH – 7.204 where salmonid fish are not present 

2A. The thirty-day average concentration 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 + 107.688– pH + 1 + 10pH – 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 + 107.688– pH + 1 + 10pH – 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 average 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 (NO3

–

) and nitrite (NO2

–

) 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 aerobic environments, nitrite 

concentrations tend to be low because nitrite is converted rapidly to nitrate, a more stable form 

(Guillette and Edwards 2005). Analyses of nitrate 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 nitrogen gas by denitrification is transported 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 either 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 and Morris 1986). Nitrate in body 

fluids can be converted to nitrite or it can accumulate because of hepatic detoxification of nitrite 

(Edwards et al. 2004). 

Several detrimental effects have been ascribed to nitrate. Concerns related to drinking water, 

including groundwater, have centered on nitrate interference with hemoglobin in the bloodstream 

resulting in a reduced ability to carry oxygen that leads to tissue hypoxia 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 synthesis, 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 levels 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, including crustaceans, insects, and mollusks. 
 214 

Table 3. Nitrate concentrations yielding 50% mortality in 96-hour tests (96-hour LC50) or other 

changes deemed relevant, e.g., altered growth, time to metamorphosis or reproductive 

characteristics (Camargo et al. 2005; Corrao et al. 2006; Mattson et al. 2007). 

Taxonomy Life stage 96-hour LC50

(mg L–1 NO3–N) 

Other threshold

(mg L–1 NO3–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 6.25 

 
 215 

Taxonomy Life stage 96-hour LC50

(mg L–1 NO3–N) 

Other threshold

(mg L–1 NO3–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 0.01–0.04 
 216 

Nitrate concentrations from 0.23 mg L–1 NO3–N to over 1,000 mg L–1 NO3–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 NO3–N for 

nymphs of mayfly species when they were subjected to a mixture of nitrate, nitrite and ammonia. 

A value of 2.2 mg L–1 was calculated as a safe concentration for long-term exposure of caddisfly 

larvae to nitrate alone. In general, early life history stages appear to be most sensitive, and 

freshwater invertebrates appear to be more sensitive than marine invertebrates. Thus, nitrate 

toxicity may have its most severe ecotoxicological effect on invertebrate growth and 

development rates, due to greater effects on larval phases. 

Toxicity values for the effects of nitrate on fishes range from 1.1 mg L–1 NO3–N for least 

observable and no observable effects on eggs to over 1,000 mg L–1 NO3–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 experiments or abnormalities correlated with nitrate 

concentrations in freshwater habitats. Impacts on eggs, embryos or fry 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, and longer-term exposures 

may produce increased tolerance (Knepp and Arkin 1973; Russo et al. 1981). 

Amphibians have been promoted as sentinel species or species that will provide an early warning 

of the ecological effects of toxicants. They suit this role because their skin increases uptake rates 

for contaminants as a byproduct of being adapted for respiration and hydration and because they 

have complex life histories spanning aquatic and terrestrial habitats. Negative effects of nitrate 

on amphibians include decreased activity, reduced feeding, disequilibrium, deformities, 

mortality, and altered timing of metamorphosis. Evidence points to nitrate acting as an endocrine 

disruptor. Overall, sensitivity to nitrate varies 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 effects in tadpoles. Factors such as water chemistry and the 

type of nitrogen compound introducing the nitrate 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 prepared 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 confound 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 populations is confounded by habitat 

destruction and the impacts of introduced species; however, nitrate toxicity should factor into 

amphibian conservation. 

Toxicity tests using birds and most mammals are lacking, and one study addressed nitrate effects 

on the reptile, Alligator mississippiensis. Preliminary findings indicate that nitrate concentrations 

between 0.01–0.04 mg L–1 NO3–N appear to affect alligators (Guillette and Edwards 2005). 

Plasma testosterone levels in both male and female juvenile alligators were negatively correlated 

with nitrate levels, and estradiol levels in juvenile females were positively correlated with nitrate 

levels. Toxicity to birds and mammals probably occurs through ingestion of contaminated water. 
 217 

Herds of cattle were unaffected by exposure to water contaminated with 0–44 mg L–1 NO3–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-BE_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 concentrations reported to cause effects on animals 

(Munch et al. 2006). However, nitrate concentrations exceed 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 uncertainty 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) toxicity, 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 concentrations is accumulating. Increases in 

concentrations of nitrate may initially lead to changes in traits of individual 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 reached 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 ecosystems would benefit from diagnostic studies of 

bottom-up and top-down interactions. These interactions, with their associated complexity and 

nonlinearity, ultimately determine the level of management needed to protect springs 

ecosystems. Furthermore, the complex and nonlinear nature of key interactions make it highly 

likely that management applied at real world, spatiotemporal scales will generate unforeseen and 

unforeseeable responses or emergent properties, which create a need to manage adaptively. 
 218 

RESEARCH TO IMPROVE MANAGEMENT OF NUTRIENTS IN SPRINGS 

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. The 

eutrophication progression scheme suggests that increased nutrient delivery to aquatic systems 

favors production of fast-growing algae that ultimately outcompete and displace 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 macrophytes 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 plants and algae do correlate with other physical 

and chemical parameters. However, we know little about their individual or combined effects or 

the form and strength of potential feedback mechanisms, including those linked to nutrient 

concentrations. Overall, numerous processes and their interactions probably 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 trophic webs, and 3) changes in habitats. Toxicity from elevated 

concentrations of ammonia and nitrate is not a major 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 webs and habitats, along with flow-on effects on 

faunal assemblages, are determined by complex interactions among bottom-up and top-down 

processes. At this time, we lack data to support rigorous evaluation of the role of nutrients in 

such changes. 

Overall, management of nutrients in spring ecosystems would benefit from compilation of 

inventories, collection of baseline data, and completion of diagnostic studies focused on bottomup and top-down interactions. The interactions, complexities and nonlinearities uncovered by 

these studies should influence the management of spring systems. Furthermore, the likelihood 

that unpredictable, emergent properties will arise when management is implemented at real 

world, spatiotemporal scales creates a need to manage adaptively. 

Designing, implementing, evaluating and applying research that improves management of 

Florida’s springs and spring-fed systems requires input from multiple stakeholders, including 

policy makers, managers, the public, and scientists 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 drivers and valued components of spring systems 

chosen with the best available information and, eventually, with guidance from the 

results of diagnostic studies 
 219 

− 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 temporal scales that support rigorous analyses 

of interactions (note that these inventories and baselines form the platform for 

adaptive management) 

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

• diagnostic studies that determine assimilation of nutrients by flora as key elements in 

nutrient budgets for spring systems, including contributions by epiphytes, macroalgae, 

vascular plants and microphytobenthos or microalgae found in sediments 

• diagnostic studies that determine how nutrients and other drivers affect overgrowth, 

shading and other relationships among periphyton, 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 include 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, including uptake and other physiological 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 relationships form the core of an adaptive 

management approach) 

− elucidate the direction, magnitude and form of limitations (e.g., nonlinear and 

nonreversible) to identify suitable targets and goals for management actions 

• diagnostic studies to identify and elucidate changes in composition and function of faunal 

assemblages related to changes in habitats in an effort to identify habitats 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 consumers in an effort to identify flora, 

grazers and trophic links to be protected or restored by management actions 

• diagnostic studies to identify and elucidate the toxic effects 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 
 220 

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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 hyacinth, 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 Invasive Plant Management spent approximately 

$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 unknown about 1) 

the history of these plants and their control in Florida; 2) the growth potential of the nonnative 

plants in springs as a function of elevated nitrate-nitrogen concentrations; 3) the social and 

ecological consequences of aquatic plant overgrowth; 4) the ecological risks associated with 

current aquatic plant control methods; and 5) the potential benefits of alternative aquatic plant 

management approaches in some springs systems. 

The Literature: What is Known 

• Major problems with nonnative plants in Florida began with the introduction of water 

hyacinth, a floating aquatic plant, into the St. Johns River in the late 19th 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 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 present in Florida before European colonization, 

or was introduced by early Spanish settlers. Chemical control programs have maintained 

water lettuce at low levels throughout Florida since the mid-1970s. 

• Hydrilla, a submersed aquatic plant, became established in several areas of Florida, 

including the Kings Bay/Crystal 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 ecosystems 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 overgrowth 
 232 

can have severe adverse effects on native plant communities, navigation, fisheries, and 

recreational desirability. 

• Water hyacinth and water lettuce emit allelopathic 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 nutrient uptake, shading, faunal habitat, direct 

filtration, and allelopathic mechanisms. 

• Ecosystem surveys indicate that water hyacinth, 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 significant 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 fluridone, a systemic herbicide 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 commonly used to control hydrilla in springs. 

• Established biological control organisms are known to adversely affect water hyacinth, 

water lettuce, and hydrilla in some springs systems, although biocontrol organisms 

generally do not maintain aquatic plant populations at low levels achieved by chemical 

control. A promising biological 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 other flowing waters have 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-nitrogen 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 with 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 pulses and ecological openings associated with 

the rapid loss of aquatic plant populations. 
 233 

• The effects of allelopathic compounds emitted 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. Variables such as dissolved nutrients, dissolved 

oxygen, biomass deposition, and floral and faunal community structure should be 

monitored before and after major aquatic plant control operations in Florida springs. 

• Water hyacinth and water lettuce are currently being managed for algal-suppression, 

nutrient recovery, and biomass utilization in a number of tropical countries, including 

places in which they are considered nonnative. While it is not known if such methods 

would be helpful in springs ecosystems, careful experimentation with floating plant 

treatment methods may be worthwhile in highly degraded springs 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 organisms 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 plants. Although there are more than a dozen nonnative aquatic 

plants established in Florida’s springs ecosystems, the vast majority of historic 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 fear that rising nitrate-nitrogen levels in 

Florida springs may further promote growth of these, and perhaps other, nonnative plants to the 

overall detriment of native plants and animals (e.g., Florida Springs Task Force 2000; Loper et 

al. 2005). 

 

Aquatic plant control programs that primarily target hydrilla, water hyacinth, and/or water lettuce 

are actively employed in several major springs systems, including Rainbow River, Silver River, 

Wekiva River, Weeki Wachee River, Wakulla River, and Kings Bay/Crystal River (Bureau of 

Invasive Plant Management 2007). Selective application of chemical herbicides is the primary 

operational method for aquatic plant control, although manual and mechanical harvest methods 

are also used in some springs systems. Approximately $173,000 was spent by the Florida 

Department of Environmental Protection (DEP) for aquatic plant control in springs ecosystems 

in 2005 – 2006. Over 80% of this amount was allocated for herbicidal suppression of hydrilla in 

Rainbow River, Weeki Wachee River, Wakulla River, and Silver River (Bureau of Invasive 

Plant Management 2007). 

This review chapter begins by generally describing the introduction, ecological 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-nitrogen 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 invasive plant proliferation on ecosystem 

structure and function in springs? 

3. What are the potential implications of aquatic plant control on ecosystem 

structure and function, particularly in terms of shifting springs toward more algaldominated ecosystem? 

4. For springs in which increased algal dominance is the primary conservation 

concern, are there opportunities for experimenting with alternative aquatic plant 

management strategies that attempt to utilize the known functions of extant 

invasive plants in support of long-term ecosystem recovery goals? 
 235 

Major Invasive Aquatic Plants: HIstory and Control in Florida 

Water Hyacinth 

Water hyacinth is a floating aquatic plant native to South America’s Amazon River drainage. 

Introduced into many countries throughout the subtropics and tropics in the late 19th and early 

20th centuries, water hyacinth is now commonly regarded as one of the world’s most problematic 

invasive plant species (Holm et al. 1977). A number of researchers have found that, under ideal 

conditions, water hyacinth is more 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 waters due to the synergistic effects of rapid deposition of 

organic matter from senescing leaves, restriction of phytoplankton production due to shading, 

and interference with atmospheric oxygen exchange (Penfound and Earle 1948; Joyce 1985). 

Development of anoxic conditions under water hyacinth 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 cover 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 specimen 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 multiplied to such an extent that extremely large 

floating mats of water hyacinth became established at many locations along the St. Johns River 

and its tributaries (see Figure 1). Boat navigation and logging operations 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 cattle forage (Buker 1982) and 

an organic mulch source for citrus trees planted in sandy soils (Maltby 1963), water hyacinth 

continued to spread rapidly into many of the state’s other drainage basins throughout the first 

half of the 20th 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 20th century. Reports from the Panhandle’s spring-fed St. Marks River suggest 

that water hyacinth was first introduced into that system as late as the 1970s (Bartodziej and 

Leslie 1998). 
 236 

The history and evolution of efforts 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 authorizing the United States Army Corps of Engineers to destroy 

water hyacinth in navigable waters, and the Florida 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 herbicide sprays were briefly used by the Corps of 

Engineers for hyacinth control around the turn of the century, but reports of severe cattle 

mortality associated with these sprays led both the Florida Legislature and U.S. Congress to 

prohibit use of these compounds in Florida by 1905 (Buckman and Company 1930). 

As a result of this prohibition, most hyacinth control 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 dislodge 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 blockages, they apparently were insufficient for 

abating the proliferation of water hyacinth. Thus, the spread and areal extent of water hyacinth 

coverage continued to increase in Florida throughout the first half of the 20th century, peaking at 

approximately 51,000 acres in the early 1960s (Schardt 1997; see Figure 3). 

Figure 1 – Water hyacinth overgrowth in St. Johns River. Illustration from March 19, 1898 

edition of Harper’s Weekly, as adapted by University of Florida Center for Aquatic and 

Invasive Plants. http://aquat1.ifas.ufl.edu/guide/stboatbg2.gif
 237 

By all accounts, the invention of the organic phenoxy herbicide 2,4-D in 1942 marked an 

important turning point in water hyacinth control. Unlike the inorganic herbicidal sprays used 

previously, 2,4-D effectively controlled 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 hormonal production. A number of 2,4-D-based spraying programs that 

targeted heavily concentrated populations of water 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 large-scale aquatic plant control 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 (Schardt 1997). Adopted into state law in 1974 as 

the official guiding principle for aquatic plant management, maintenance control is defined by 

the operational goal of maintaining water hyacinths (and other invasive nonnative aquatic plants) 

Figure 2 – Water hyacinth shredder in the Withlacoochee River, circa 1940. Photo by United 

States Army Corps of Engineers, as adapted by the University of Florida Center for Aquatic 

and Invasive Plants. http://plants.ifas.ufl.edu/guide/mechconold7js.jpg
 238 

at the “lowest feasible level” (Florida Statute 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 interfere with beneficial uses of water bodies can be maintained. 

Aside from the obvious objective of effectively suppressing nonnative plants, some research 

suggests that maintenance of plants at small populations 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 control 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 systemic herbicide) (Bureau 

of Invasive Plant Management 2007). Potential implications 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 operations on water hyacinth populations in 

Florida.Graph adapted from Florida Department of Environmental Protection, Bureau of 

Invasive Plant Management. http://www.dep.state.fl.us/lands/invaspec/images/Graph%20wh%20acres.jpg
 239 

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 lettuce 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 uncertainty 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 

18th century (Stuckey and Les 1984). Some researchers speculate that 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 that 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 near the Suwannee River’s Manatee Springs and 

several springs along the St. Johns River, indicating that these springs have had water lettuce 

populations for well over two centuries. Other historical accounts in springs include Webber’s 

(1897) sighting of large water lettuce populations in the Wekiva River during the late 19th

century, and Carr’s (1994) reports of water lettuce as a common component of springs 

ecosystems throughout north Florida during the 1940s. Field observations by Odum (1957) 

suggest that water lettuce growth in Silver Springs during the mid 20th 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.jpg
 240 

seasonal patterns and that large 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 large-scale displacement of water lettuce in Florida during the late 

19th and early 20th centuries (Schmitz et al. 1993; Carr 1994). However, large populations of 

water lettuce quickly emerged in Florida and other states after the commencement of water 

hyacinth control in the 1950s, likely due to water lettuce’s relative resistance to 2,4-D (Eggler 

1953; United States Army Corps of Engineers 1973). Researchers during the mid 1960s 

discovered that water lettuce was controlled effectively by diquat dibromide (Weldon and 

Blackburn 1967), and since that time 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 based solely on hand removal of plants has 

drastically reduced water lettuce populations 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 20th century, naturalized populations of 

hydrilla were identified in several Florida waterways by the early 1960s (Schmitz et al. 1993). 

Over subsequent decades the plant has rapidly spread into many aquatic systems throughout 

Florida and the southeast United States. Of the $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 was spent for hydrilla 

control in spring systems such as Wakulla River, Weeki Wachee River, Silver River, and 

Rainbow River in fiscal year 2005 – 2006. 

Several features make hydrilla an extremely effective 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 ability 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 interfere with flood control structures 

(Schmitz et al. 1997; Jones and Beardall 2005). In addition, hydrilla’s root system is 

characterized by many large tubers that often resprout after leafy growth is suppressed through 

herbicides or other disturbance (Schardt 1997). Some research suggests that hydrilla may also 

have lower light requirements for photosynthesis relative to native submersed plants, thereby 

allowing it to colonize deeper areas of water bodies that previously 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 phytoplankton production in underlying waters 

(Schmitz et al. 1993). Dense hydrilla also has been known to adversely affect sports fisheries by 

providing increased habitat cover that restricts prey availability for 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 benefit certain fish and wildlife populations. For example, it is 

commonly reported that moderate hydrilla coverage provides superior game fish habitat 

(Schmitz et al. 1993), which leads many fishermen to advocate for management strategies that 

maintain significant hydrilla coverage in popular fishing lakes (Jones and Beardall 2005). 

Hydrilla is also known to provide highly attractive habitat for manatees (Campbell and Irvine 

1977), certain macroinvertebrates (Schramm and Jirka 1989), and diverse assemblages of native 

water fowl (e.g., Johnson and Montalbano 1984; Esler 1990). Thus, one of the most complex 

challenges currently facing Florida’s aquatic plant managers is development of techniques that 

better maintain the positive habitat values of hydrilla in areas where it is permanently 

established, while still preventing severe overgrowth and spread of the plant into new areas 

(Jones and Beardall 2005). 

Modern hydrilla control methods in Florida rely heavily upon fluridone, a slow-acting systemic 

herbicide, and endothall, a fast-acting contact herbicide (Jones and Beardall 2005). Because 

effective treatment by fluridone requires long contact times that are prohibitively difficult and 

expensive to obtain in flowing systems, a dipotassium salt formulation of endothall (trade name 

Aquathol) is typically used for hydrilla control in springs (DEP 2005). However, serious 

concerns about the long-term sustainability of current chemical control methods are raised by 

recent findings of independently evolved fluridone-resistant hydrilla biotypes in Florida lakes 

frequently treated with fluridone (Michel et al. 2004). Although no endothall-resistant strains 

have been identified, many researchers argue that future development of resistance among 

hydrilla populations repeatedly treated with endothall, such as those in springs, is possible or 

perhaps even inevitable (Jones 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.jpg
 242 

Diquat and copper (including chelated copper and copper sulfate formulations) are the major 

herbicides aside from fluridone and endothall that are currently listed for hydrilla control 

(Langeland 1996; Bureau of Invasive Plant Management 2007). Copper and diquat were 

commonly used for hydrilla control 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 employed in Florida due to resistance concerns 

(Jones and Beardall 2005). However, increased use of copper for hydrilla control in some Florida 

springs is problematized by regulatory language that places restrictions on copper herbicide 

usage in areas frequented by manatees (State of Florida Division of Administrative Hearings 

1993). These restrictions on copper herbicides were adopted 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 Bay/Crystal River, in 

which copper was commonly used as an herbicide/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 NONNATIVE PLANTS IN SPRINGS 

The increased nitrate-nitrogen contamination observed in springs throughout Florida is of great 

concern to ecosystem managers and 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 the potential for nitrate-nitrogen to favor the 

proliferation of water hyacinth, water lettuce, hydrilla, and other invasive 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 treatment 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 conditions relevant to those in Florida springs. As discussed in more 

detail below, the available literature strongly suggests that nitrogen limitation of these 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 mitigation 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, potassium, and iron have also been found to limit 

water hyacinth growth (Gopal 1987). While nitrogen limitation to water hyacinths is frequently 

indicated when N/P ratios are less than 7 (Wilson et al. 2005), experimental trials indicate that, at 
 243 

very high nutrient levels, dissolved nitrogen typically will be depleted 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, calcium, 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 limited by these nutrients in most Florida spring ecosystems. While 

long-term enrichment trends observed in springs 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, report a water hyacinth biomass 

doubling time of 10 days at nutrient levels of 0.28 mg/l TN1

 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 the 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 nutrients were not a major limiting factor. 

Two recently developed water hyacinth models present nitrogen concentration values that could 

be expected to limit water hyacinth growth 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 nitrogen concentrations 

clearly fall within a range of direct 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 availability 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 for nutrient availability and replacement as 

affected by flow velocity, inferential caution is warranted due to important morphological, 

chemical, and climatic differences between the Tha-Chin River and typical Florida springs 

systems. 

Perhaps the most intriguing clue about potential nitrate-nitrogen thresholds for water hyacinth in 

Florida springs comes from Webber’s (1897) observation of water hyacinth plants showing a 

stunted growth form in the Silver River during the late 19th 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 scientific literature discussed above, it is 

reasonable to hypothesize that Webber’s (1897) observations may be indicative of nitratenitrogen serving as a physiological constraint to water hyacinth growth in the Silver River at the 

 1

 All nitrogen concentrations reported by Bartodziej and Leslie (1998) are in terms of TN. 

Relative contribution of nitrate-nitrogen to the TN measurement is not given. 
 244 

background water quality levels of the late 19th and early 20th century. Confirmation of this 

suggested relationship, however, would require detailed assays of water hyacinth growth 

response to variable nitrate-nitrogen concentrations in flowing mesocosms, and/or comparative 

field studies of water hyacinth growth characteristics 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 water hyacinth and water lettuce 

under experimental nutrient treatments (e.g., Aoi and Hayashi 1996) may make it reasonable to 

assume that the potential nutrient limitation thresholds 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 (Aoi and Hayashi 1996). Schmitz et al. (1993) hypothesize, while 

also noting the lack of specific studies testing this hypothesis, that increased levels of nitrogen 

may favor water lettuce to out-compete native 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 Florida springs. Productivity measurements given by 

Odum (1957) indicate that growth of water lettuce 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 19th century could 

suggest that background nutrient levels in that particular spring fed river were not a serious 

limiting factor for water lettuce growth. As suggested 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 enrichment 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 variety of evidence is suggestive of 

important relationships between hydrilla growth and increased nitrogen enrichment in Florida’s 

aquatic systems. For example, tank experiments clearly indicate that hydrilla’s growth and ability 

to out-compete native tape grass (Vallisneria americana) under high light conditions 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 significant relationship between dissolved total nitrogen and the 

extent of hydrilla occurrence in Florida lakes, although stronger correlative relationships were 

identified with other water quality parameters such as alkalinity, pH, and total dissolved 

phosphorus. 
 245 

It must be cautioned, however, that straightforward 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 dissolved nutrient availability for 

hydrilla and other submersed plants (Canfield and Hoyer 1988). Following this second point, 

Terrell and Canfield (1996) found that dissolved nitrogen availability for aquatic plants in Kings 

Bay/Crystal River, which had a mean TN concentration of approximately 0.26 mg/l at the time 

of study (neither nitrate-nitrogen contribution to TN nor sediment nitrogen were reported), 

greatly exceeded the biological demands necessary for maximum growth and accumulation of 

hydrilla. Based on this finding, Terrell and Canfield (1996) concluded that nutrient reduction 

likely was not an effective strategy for reducing growth of hydrilla in Kings 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 needed to determine if 

nitrate-nitrogen reduction is a feasible strategy for hydrilla suppression in the context of other 

Florida springs. 

NONNATIVE PLANTS AND ALGAL COMMUNITY DYNAMICS 

While overgrowth of nonnative invasive plants is a serious management issue, most conservation 

and management concern in Florida springs over recent 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 structural associations between native 

aquatic plants, epiphytic algae, and animal communities historically found in springs ecosystems 

(Odum 1957), it stands to reason that the obvious structural changes in plant community caused 

by invasive plant growth potentially could trigger structural changes in the algal and animal 

communities. 

This section utilizes scientific literature about major invasive plants to explore two mechanisms 

by which these plants might contribute to a successional trajectory in which undesirable algae 

become dominant in springs ecosystems. The first proposed mechanism, referred to as “boom 

bust,” focuses on the implications of the increased productivity, biomass accumulation, and 

vulnerability to catastrophic losses and rapid community reorganization that are associated with 

invasive plant communities. The second proposed mechanism focuses on the allelopathic 

properties of the invasive plants, with particular attention to the potential implications of 

algicidal allelopathy. 

Boom-Bust Hypothesis 

Perhaps the most basic functional trait shared by water hyacinth, water lettuce, hydrilla and other 

invasive aquatic plant species is higher productivity relative to competing native plants (Schardt 

1997). Thus, the introduction of a successful invasive 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 variety of stochastic natural events such as saltwater storm surges 
 246 

(Terrell and Canfield 1996), high water/floods (Bartodziej and Leslie 1998; Bureau of Invasive 

Plant Management 2007), and killing frosts (Joyce 1985; Schardt 1997). 

Figure 6 gives a simplified diagram of mechanisms by which a boom bust cycle associated with 

invasive plants potentially could lead to increased nuisance algae in springs systems. The 

diagram shows that one major ecosystem effect often associated with invasive plants is 

competitive displacement of native plant species (e.g., Schmitz et al. 1993; Bartodziej and Leslie 

1998), some of which may have previously suppressed 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 water 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 competitive displacement of native plants followed by 

a bust cycle characterized by destruction of dominant invasive plant populations may favor 

succession by opportunistic filamentous algae. Filamentous algae have been observed to fill in 

such ecological voids after aquatic 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 biomass 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 systems, have been documented in other ecosystem 

contexts (see Clugston 1963; Bicudo et al. 2007). 

A mechanistic feedback that may further favor proliferation of nuisance algae after a boom bust 

invasive plant cycle is implied by the work of Stevenson et al. (2007), who found that nutrient 

enrichment of sediments was significantly associated 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 plants, aquatic plant control activities were 

suggested as a possible source of enriched sediments documented 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 time 
 247 

relationships appears warranted in springs, particularly because invasive plant growth and 

control has been linked with nutrient 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 contributor 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 organic sediments in the 

Kissimmee River and Lake Okeechobee. In a highly 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-nitrogen, 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 enrichment in springs, it is important to note that a 

variety of factors such as flow velocity, external inputs of organic 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 exert significant influence over sediment nutrient 

levels. Detailed measurements of sediment nutrients, dissolved nutrients, and plant/algae 

community characteristics in discrete samples taken directly before, directly after, and in 

sequential intervals following aquatic plant control would provide invaluable information about 

the overall ecosystem effects of current management techniques on springs ecosystems. 

Allelopathy Hypothesis 

Algal community composition in springs and other streams are clearly shaped by ecological 

factors such as nutrient availability, flow velocity, light intensity, grazer abundance, and 

substrate quality (e.g., Stevenson et al. 2007). In addition to these 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 indicate that water hyacinth, water lettuce, and hydrilla all emit 

allelopathic compounds that restrict growth of algal and/or plant competitors. In the case of 

water hyacinth, Jin et al. (2003) conducted lab experiments indicating that various compounds 

extracted from water hyacinth roots had algicidal properties on Chlorella sp., Scendesmus 

obliquus, and undifferentiated phytoplankton that were comparable in activity to copper sulfate – 

a commonly used commercial algicide. Similarly, Aliotta et al. (1991) isolated several 

allelopathic chemicals from water lettuce that showed 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 hydrilla negatively affected the growth of vascular aquatic plant 
 248 

Ceratophyllm sp. through allelopathic mechanisms. However, Glomski et al. (2002) later argued 

that allelopathy is only a very minor factor in hydrilla’s overall competitive success against other 

aquatic plants. 

The potential implications of algicidal allelopathy in the floating plants are interesting to 

consider in the context of Florida springs systems. On the one hand, allelopathic emissions from 

floating plants may be a mechanism by which fringe mats of floating plants in spring runs serve 

as a constraint for algal growth and accumulation in habitat areas beyond those covered by the 

plants, including beds of native submersed plants located in the main stream flow. If this is the 

case, then the algicidal properties of floating plants may potentially serve as a buffer against 

algal overgrowth in nutrient-enriched 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 algicidal activity of floating plant allelopathic emissions on springsspecific algae species would require isolation of algicidal compounds 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 such compounds in the water of springs would be 

required to understand the extent to which allelopathic 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 the EPA for use in aquatic systems is the primary 

tool used by aquatic plant managers. The most obvious ecological concern associated with use of 

aquatic herbicides is non-target damage to native plants, algae, and animal communities, either 

through direct toxicology of herbicides or the ecological consequences of rapid plant senescence. 

Although aquatic plant managers and herbicide applicators take great precautions to avoid major 

non-target impacts, the complexities of ecosystems make some non-target damage unavoidable. 

This section reviews literature associated with two of the most plausible and commonly 

suggested mechanisms by which chemical control 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 direct toxicology and/or habitat destruction. 

Potential implications of these mechanisms for Florida springs, particularly in terms of 

succession towards filamentous algae/cyanobacteria communities, are also discussed. 

Herbicide Selection Hypothesis 

Development of herbicide resistance among algae has been repeatedly demonstrated at both the 

cellular level of individual species (i.e., a specific algal species evolves resistance to an 

herbicide) and the ecological level 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 major concern associated with 
 249 

repeated use of herbicides in aquatic systems is the potential selection for herbicide-resistant 

algal biotypes. In particular, recent 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, endothall, glyphosate, and 2,4-D. When relevant 

information is available, attention is particularly given to potential selection concerns associated 

with Lyngbya wollei. 

Copper 

Copper compounds are most often used in aquatic systems as algaecides, but are listed an 

approved method for control of hydrilla and other submersed plants species in Florida (Bureua of 

Invasive Plant Management 2007). Largely due to concerns about sediment contamination and 

bioaccumulation risks in manatees (O’Shea 1984; Facemire 1991; Leslie 1992), copper 

herbicides have not been widely used in Florida since the early 1990s. However, copper 

herbicides were used extensively for hydrilla control in the Kings Bay/Crystal River ecosystem 

during the 1970s and 1980s (Haller et al. 1983), and currently are being researched as a potential 

option for filamentous algae control 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 research, 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 increasingly copper resistant algal communities in a 

number of lakes and reservoirs throughout the northeastern 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 understanding of algal community dynamics in 

Florida spring systems, such as Kings Bay/Crystal River, where copper herbicides have been 

used historically. 

Diquat 

Scientific literature indicates that concerns associated 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 diquat do accumulate in aquatic 

sediments, most recent research suggests that diquat forms complex bonds with sediments that 

render it biologically unavailable (Emmett 2002). 

Bioassay research indicates that there is a wide range of tolerances among algal taxa to diquat 

exposure. Peterson et al. (1997) found that 50% inhibition of biomass production occurred 
 250 

among two green algae taxa (Scendesmus quadricauda and Selenastrum capricornutum) at 

concentrations of approximately 0.6 mg/l of diquat, while 50% growth inhibition for diatoms 

(Nitzschia sp. and Cyclotella meneghiana) and cyanobacteria (Oscillatoria sp. and 

Pseudoanabaena sp.; Microcystis sp.) occurred at concentrations 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 sensitivity 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% inhibition 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 observations 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 tolerances exhibited by algal-taxa is that repeat 

treatment with diquat may exert a chronic selective pressure that favors community-level 

dominance by diquat-resistant species and/or biotypes over time. Much like with the copper 

herbicides, diquat treatment history 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, which is generally referred to by the trade name 

Aquathol, is the primary herbicide used for control of hydrilla in flowing water systems. While 

Aquathol is not typically recommended for algae control purposes, an amine endothall salt (trade 

name Hydrothol) is registered for control of filamentous algae and submersed plants in Florida. 

However, Hydrothol is not commonly used in Florida’s public waters due to its relatively high 

toxicity to fish and other aquatic fauna. Field observations suggest 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 the 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 name “Rodeo”) is sometimes used for control of 

water hyacinths in Florida (Bureau of Invasive Plant Management 2007). While generally not 

used as an algaecide in aquatic environments, bioassays indicate that glyphosate does have 

algaecidal action. Wong (2000) reported significant growth inhibition of Scendedesmus 

quadricauda at 2 mg/l of glyphosate and complete inhibition at 20 mg/l. In very recent bioassay 

work with Microcystis auroginosa, Lopez-Rodas et al. (2007) reported considerable 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 concentrations of 120 mg/l. While the 

glyphosate concentrations of concern to Lopez-Rodas et al. (2007) are at least an order of 
 251 

magnitude over realistic field concentrations associated with aquatic plant control (Langeland 

2006), it is plausible that the lower end of concentrations showing growth inhibition in the Wong 

(2000) experiment may be reached during aquatic plant control operations. Bioassays testing 

glyphosate response curves of algal species in Florida springs would be needed to better 

determine algal-selection risks associated with this herbicide. 

2,4-D 

Algal toxicity bioassays with 2,4-D have found that 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 stimulatory effect is that 2,4-D’s 

mode of herbicidal action is mimicry of the plant hormone auxin, rather than direct targeting of 

photosynthesis mechanisms. Toxic effects on algae 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 plant 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, meaning that changes in algal community could be promoted by 

2,4-D through an enrichment mechanism, rather than a toxicity mechanism. Like with the other 

herbicides, specific bioassays testing growth response to 2,4-D among algal species found in 

Florida springs would be needed to better understand the community level effects of this 

herbicide. 

Attractor – Catastrophe Hypothesis 

A final mechanism by which invasive plants and their subsequent management may profoundly 

affect ecosystem structure in springs is through a sequential 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 become 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 “catastrophes” that affect nonnative plant 

communities. Although any habitat disturbance that destroys non-native plant communities is 

potentially a catastrophic sink for associated faunal communities, primary focus is given here to 

the non-target toxic effects and/or general habitat disturbance that may be associated with 

aquatic plant control activities. The major rationale for this focus 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 associated with water 

hyacinth in the St. Marks River clearly demonstrates the high habitat values that can be provided 

by this nonnative plant in Florida springs. Contrary to other ecosystem contexts in which large 

mats of water hyacinth have been clearly shown to depress dissolved oxygen, coverage by water 

hyacinth in the flowing waters of the St. Marks River did not show significant oxygen profile 

differences as compared to strap-leaf sag (Sagitraria kurziana) or other native plant 
 252 

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 community associated with water hyacinths was 

generally more diverse in terms of invertebrate taxa and included significantly 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 populations among water hyacinth in 

the St. Marks River indicated that water hyacinth 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 habitat for the common moorhen (Gallinula chloropus) (Bartodziej 

and Weymouth 1995). 

Figure 7 – Water hyacinth fringe in River Rise Spring #3, St. Marks River. Photo by Northwest 

Florida Water Management District (2006). 

http://www.nwfwmd.state.fl.us/rmd/springs/Wakulla_StMarks/photos/rise3_1.jpg
 253 

 

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 (treatment) 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 suggested by Carr (1994), it may be reasonable to assume that the similar 

structure of water lettuce and water hyacinth is likely to provide 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 large numbers of spring run crayfish and 

apple snails in water lettuce harvested from the Ichetucknee River in 2000 – 2001. Similarly, 

Corrao et al. (2006) report high concentrations of apple snails on hydrilla and apple snail eggs on 

water lettuce at Wacissa Springs. In other ecosystem contexts, hydrilla is known to support large 

macroinvertebrate populations (Schramm and Jirka 1989) and attract diverse assemblages of 

water fowl (Johnson and Montalbano 1984; Esler 1990). Dramatically increased manatee 

utilization of Kings Bay/Crystal River beginning in the 1960s (Kochman et al. 1985) coincides 
 254 

with the introduction and rapid spread of hydrilla, 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 habitat 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 Springs and herbicide suppression 

of plants containing snail eggs at Wacissa Springs may have had significant negative effects on 

apple snail populations, but noted that more detailed studies would be necessary to test these 

observations. Similarly, Evans (2007) observed that harvest of water lettuce from Ichetucknee 

Springs resulted in direct mortality of several faunal species, including 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 raised significant non-target concerns in Florida springs. Most 

notably, copper herbicide use in Kings Bay/Crystal River was discontinued in the late 1980s due 

to the finding of severe accumulation of copper 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 grazers and that sediment contamination can have long-term 

effects on the structure and function of benthic communities. Increased levels of Lyngbya wollei

often were observed in Kings Bay/Crystal River subsequent to hydrilla herbicide treatments (see 

Cowell and Botts 1994), which led Evans et al. (2007) to hypothesize that toxicological 

suppression of grazer communities associated with copper herbicide usage may have contributed 

to subsequent algal overgrowth. 

Risk assessments on four grazer species performed by Mastin and Rodgers (2000) indicate 50% 

lethal concentration dosages2

 (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 treatment 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), however, reported that poor control of hydrilla using standard 

dosage rates sometimes led aquatic plant managers to apply significant higher 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 decreases as a function of increased hardness due to 

carbonate-induced precipitation of copper compounds (Mastin and Rodgers 2000). Thus, it can 

be argued that non-target risks in Florida springs ecosystems from dissolved 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 dose (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 Florida springs ecosystems should be implemented 

cautiously and carefully monitored to minimize 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 assessment 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 studies, and suggested that a combination 

of direct toxicity and the destruction of aquatic plants through 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 hypothesize 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 little 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 glyphosate (Langeland 2006) at concentrations used for aquatic 

plant control. However, Wang et al. (1994) does document bioaccumulation of 2,4-D after water 

hyacinth treatment in some fish species, and a breakdown product of 2,4-D (2,4-

dicholorophenol) has been shown to have relatively high toxicity to earthworms (Roberts and 

Dorough 1984). Relyea (2005) has raised concerns about the effects of the Roundup formulation 

of glyphosate on amphibians, but Langeland (2006) notes that the toxicological effects reported 

by Relyea (2005) are associated with an adjuvant in Roundup that is not used in the aquaticapproved formulations of glyphosate (e.g., Rodeo). 

 

Aside from habitat loss and toxicology, the other major concern associated 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 approximately two weeks following the herbicide treatment, thereby 

resulting in “a detectable but temporary impairment of the aquatic community in Rock Springs 

Run” (Wetland Solutions, Inc. 2006). 

Some of the most detailed monitoring of aquatic plant control in a springs 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 concerns were raised after state 

park biologists observed a significant die-off of spring run crayfish following at least one 

subsequent Aquathol treatment (DEP 2006). In response to these concerns, DEP (2006) 

conducted a management experiment in which crayfish 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 were 
 256 

deployed nine days after the herbicide application and again retrieved four days later. Data 

sondes showed that dissolved oxygen dropped from as high as 4.3 mg/l to approximately 2.0 

mg/l after the herbicide treatment, and Aquathol exposure was measured 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 experiment. Although these results did not support the 

hypothesis that previously observed crayfish mortality was caused by aquatic plant control, the 

possibility of drastically reduced crayfish fitness as a result of prolonged 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 Wakulla River before (4/4/2002) and after 

(5/6/2002) Aquathol treatment of hydrilla (DEP 2002). 

RECOMMENDATIONS FOR FURTHER RESEARCH 

This chapter opened by asking four sets of questions related to nonnative plants in Florida 

springs. The first asked to what extent increased nitrate-nitrogen might be a factor in the spread 

of the invasive plants water hyacinth, water lettuce, and hydrilla, and at what level might 

invasive plant growth in springs be limited by nitrate-nitrogen. Available literature does not 

provide a clear answer for determining 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 flowing mesocosms and, in the case of hydrilla, with 

realistic sediment nutrient concentrations would be necessary to determine more accurate 

nutrient limitation parameters for nonnative plants in spring systems. 

The second question asked about the potential implications of invasive plant proliferation on 

ecosystem structure and function in springs. Available literature indicates that biomass 

accumulation, suppression of native plants, and algicidal emissions associated with nonnative 
 257 

plants may produce feedback mechanisms that together make springs systems more prone to a 

successional trajectory of overgrowth by nuisance algal 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 patterns in nonnative plant communities would be 

needed to better understand the relationships between nonnative plants and algal overgrowth. 

The third question asked about the potential implications of aquatic plant control on ecosystem 

structure and function. What emerged from a review of literature was a wide range of concerns 

about possible algal selection, habitat loss, faunal toxicology, and dissolved oxygen suppression 

that may be associated with aquatic plant control. With the notable exceptions of circa 1980s 

work conducted by several agencies in Kings Bay/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 concerns over crayfish mortality after Aquathol 

applications (DEP 2005), there has been very little monitoring 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, detailed 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 restoration of desired ecosystem functions in highly degraded springs and 

spring run ecosystems? Given the notorious history 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 context of springs ecosystems is undoubtedly controversial from the 

outset. However, discussions among ecological restoration practitioners indicate an increasing 

willingness to consider the proposition that, given the diverse realities of modern environmental 

change (e.g., climate change, nutrient enrichment, toxic contamination, and global species 

mixing), more adaptive approaches to nonnative species control should be considered 

(D’Antonio and Meyerson 2002; Ewel and Putz 2004; Gobster 2005). From an adaptive 

management perspective, the overriding questions might be: Are nonnative species management 

efforts directly leading to an even worse set of problems (e.g., succession 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 concerns facing many springs ecosystems is the 

proliferation of nuisance algae such as Lyngbya wollei, an adaptive management question 

implied from several lines of evidence presented above is the extent to which aquatic plant 

control may be a contributing factor in this proliferation. A complementary question is how 

alternative methods of aquatic plant 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 “control” 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 newly discovered nonnative plant populations from 

becoming irreversibly established are also clearly justified. Instead, the 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 the most sustainable longterm method for controlling water hyacinth, water 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 adjusted at a large scale in response to biocontrol 

introductions (Haller 1996). For springs in which current aquatic plant control methods may be a 

significant disturbance factor that selects for algal overgrowth, increased use of floating plant 

biocontrol organisms may be a promising alternative 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 significant control effects. 

Center et al. (2005) found that herbivory by water hyacinth weevils (Neochetina eichhorniae and 

N. bruchi) greatly reduced the competitiveness of water hyacinth in relation to water lettuce. 

While water hyacinth’s superior competitive abilities 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 weevil populations reduced the rate of water hyacinth expansion by 

approximately 10% over the course of a growing 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 studies or augmentation estimates for water hyacinth 

moth (Sameodes albiguttalis), another effective biological control 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 sites 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 ecosystem 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 competition, 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) detail 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 project to better understand the nitrogen reduction 

implications of allowing for greater colonization of water hyacinth and water lettuce in Lake 

George (SJRWMD 2006). Detailed ecosystem surveys by Bartodziej and Leslie (1998) in the St. 

Marks River clearly indicate the high habitat value of water hyacinth 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 numbers of resident and wintering manatees, it is 

reasonable to suspect that an increased availability of water hyacinth, which is known as a 

nutritious and sometimes preferred manatee forage source (Lomolino 1977), might reduce 

herbivory pressure that can adversely affect desirable native submersed plant populations 

(Hauxwell et al. 2004). Studies could include small scale pilot projects that study the ecosystem 

effects of variable water hyacinth and water lettuce coverage levels in springs, particularly in 

those now affected by severe algal overgrowth. Comparisons among springs ecosystems with no 

aquatic control of floating plants (e.g., St. Marks River) and widespread 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 associated with hydrilla in some springs 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 ability 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 associated with hydrilla in Florida springs 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 radically 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 discovered in Kings Bay/Crystal River in 1992, may 

effectively prevent hydrilla from topping out is potentially of great importance for other springs 

ecosystems affected by hydrilla. Similar to the field 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 experimental treatments containing biocontrol 

organisms Hydrillia pakistanae and Bagous hydrillae. Increased experimentation with release of 

such biocontrol organisms clearly should be a priority for Florida springs ecosystems affected by 

well-established hydrilla populations. 
 260 

Surprisingly, recent research in the Potomac River has identified significant positive correlations 

between hydrilla colonization 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 negative relationship to coverage by Lyngbya wollei, perhaps 

indicating that hydrilla – particularly if overgrowth is controlled 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 terms 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 consequences of different control strategies (e.g., 

biocontrol, herbicides, manual removal, native plant restoration, and various combinations of 

these) over time. 

CONCLUSION 

There is little question that nonnative plant invasions 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 enrichment of springs may have contributed to 

overgrowth by nonnative aquatic plants, current studies provide insufficient information for 

making a definitive determination of nitrate-nitrogen concentrations that would effectively limit 

water hyacinth, water lettuce, or hydrilla growth in springs. Scientific literature suggests that 

both nonnative plant overgrowth and aquatic plant control techniques to suppress such 

overgrowth have the potential to serve as severe disturbances that could promote succession to 

algal dominated states in springs ecosystems. Because it is extremely unlikely that nonnative 

plant species can be entirely eliminated from systems 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 springs 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 and successional impacts associated with current 

aquatic plant control methods; 2) increased research and experimentation with biocontrol 

organisms; and 3) careful experimentation with ecosystem recovery methods that utilize the 

nutrient uptake, algal suppression, and habitat values of floating plants. 
 261 

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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 Engineering Sciences, University of Florida 

Summary 

The relationship between nutrients and spring ecosystem structure and function primarily focuses 

on the state-wide increase in spring nitrate concentrations derived from anthropogenic sources 

and the concurrent observed visual decline of these ecosystems. However, the apparent 

correlation between increased nitrate loading and declining aesthetic appearance of spring 

ecosystems has only anecdotally provided evidence for a causative relationship. Organism-level 

studies, focused on single species of algae, vascular 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 population structure in springs. Firm 

conclusions linking rising spring nitrogen concentrations and changes to individual components 

of biotic communities remain illusive, in some measure due to the immense complexity of spring 

ecosystems and the large number of external forcing functions (energy and matter inputs) that 

shape their structure and functions. 

General systems theory predicts that any significant 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 the 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 responses 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 concentrations and declining ecosystem primary production and 

photosynthetic efficiency. While this possible cause 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 levels appear to have the opposite effect of 

reducing overall primary productivity. It can be theorized that spring ecosystem responses to 

nutrient increases integrate the observed complexity of their hundreds of important plants and 

animals. In their natural, moderate to low nutrient state, springs may be adapted for optimal 
 272 

efficiency of light utilization and maximum ecosystem metabolism (Maximum Power Theory). 

Rising nutrient concentrations may result in the competitive advantage of “weedy” algae and 

plant species that can capture light and spatial resources through higher rates of net productivity 

but lower gross productivity. Continuing elevated nutrient concentrations in concert with 

differential grazing pressure, reductions in flow, or loss of other top-down control mechanisms 

may be linked to observed plant community shifts from adapted submersed aquatic plants with 

sparse but highly productive periphyton communities to systems dominated by benthic and 

attached filamentous algae. Data are also presented 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 review that an ecosystem approach is essential for future research 

efforts to provide a greater understanding of the relative interactions between the myriad 

physical, chemical, and biological fluxes present in springs and their normal responses to rising 

nutrient levels. In summary these recommended applied research activities might include the 

following: 

• Detailed baseline investigations of the “normal” or “existing” ecosystem-level structure 

and response of a broad sample of first and second magnitude springs 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 forcing functions including discharge rates, 

groundwater chemistry, nutrient concentrations, and recreational intensities; 

• Multivariate statistical analysis of these baseline data to look for significant and 

repeatable patterns between independent and 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 nutrient conditions under realistic and relevant 

spring environmental conditions; 

• Whole-spring manipulation studies (if necessary) to test the effects of possible 

management techniques such as controlling levels of springshed nutrient loading, human 

recreational activities, alternative control methods for invasive species, and estimation of 

optimal consumer carrying capacities (e.g., manatee density). 

There exists a wide-spread recognition of the environmental and economic importance of 

artesian springs in Florida. A generous commitment of public dollars and a focused and logical 

research agenda will be critical to pulling these precious but threatened natural resources back 

from their current declining path. 
 273 

ECOSYSTEM STRUCTURE

Primary Producers 

The paradigm of eutrophication is that nutrient enrichment (nitrogen and or phosphorus 

primarily) will result in an increase in biomass in primary producers often with negative 

ecological and economic consequences (Cloern 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 estuarine systems, nitrogen is typically identified 

as the nutrient limiting primary production. The effects of nitrogen on marine and estuarine 

systems, the pathways for nitrogen transport between land and aquatic habitats, 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 commonly implicated as 

limiting for primary production and has been highly 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 nutrient 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 producers 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 the water column were significantly related to 

benthic algal biomass, while Francoeur’s (2001) meta-analysis of 237 nutrient enrichment 

studies in temperate streams found that 16% indicated a nitrogen (N) response, 18% indicated a 

phosphorus (P) response, 23% required 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 material, thus preventing or reducing the downstream 

export of inorganic nutrients. However, these lotic ecosystems can become saturated with 

respect to relatively conservative inorganic nutrients such as phosphorus and recalcitrant forms 

of organic nitrogen and in some instances may not be able to further assimilate inorganic forms 

through uptake, microbial utilization, and denitrification. It is unclear 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 limiting chemical constituents or forcing functions 

such as light availability. Once saturation occurs, nitrogen export to downstream ecosystems 

will increase proportionally to nitrogen loading (Bernot and Dodds 2005). Whether saturation of 

nitrogen or phosphorus has occurred in Florida spring ecosystems remains uncertain in most 

cases and is obviously a function of individual spring biogeochemistry. In at least one case 

within portions of the Wekiva River system, which is fed by Wekiwa and Rock Springs (WSI 

2007a) there is evidence that phosphorus saturation 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 increased nutrient loads and that if these thresholds are exceeded, 

harmful ecological consequences may occur. Our use of this concept has origin 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 examined in the context of ecosystem response to 

perturbation (Figure 1). An example of a subsidy-stress gradient is the productivity of hardwood 

forested swamps in response to flooding, in which forested swamp productivity was observed to 

increase under seasonal flooding conditions but to decline under permanent flooding (Conner 

and Day 1992). With regard to spring ecosystems, it is likely that a variety of physical, 

biological, and chemical inputs may operate individually or simultaneously within subsidy-stress 

gradients. For instance, stream velocity can replace nutrients and remove waste products at low 

velocities, but is capable of scouring at high velocities, and grazing 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 effects of herbivorous snails, carnivorous fish, and a 

trace metal (elemental cadmium) on primary productivity in stream mesocosms. 

It is hypothesized in this chapter that nutrients also function along a subsidy-stress gradient in 

aquatic ecosystems. At extremely low concentrations 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 likely to have unique performance curves for primary producers by 

functional group, i.e. algae vs. vascular plant, or even by species within these groups. 

Considering this, when monitoring ecosystems for impairment or the effects of nutrient 

gradients, whole ecosystem measures may produce diametrically opposite results than singlespecies studies and will be more informative for documenting overall ecosystem impairment than 

measures of individual functional groups. 

Algae 

Pseudo-phytoplankton – The development of plankton in lotic ecosystems is largely prevented 

due 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 either 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 periphytic primary producers, light, nutrients, 

temperature, and velocity. In flowing water systems, stream discharge or velocity will influence 

their rate of delivery downstream. Given that phytoplankton populations 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). However, should declines in spring discharge rates 

occur to the point that water residence 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 spring 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 input 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 which have examined pseudo-phytoplankton collected 

from spring ecosystems and their response to manipulations of ambient nutrient levels. Frazer et 

al. (2002) conducted nutrient addition bioassay of pseudo-phytoplankton collected from several 

locations of the Chassahowitzka, Crystal, Homosassa, and Weeki Wachee Rivers. These 

samples were allowed to grow under near-optimal light and temperature conditions under a 

variety of nutrient treatments. These treatments included a control (without any additional 

nutrient additions), nitrogen (N) addition (0.4 mg NO3-N/L), phosphorus (P) addition (0.04 mg 

PO4/L), silica (Si) addition (0.4 mg Si/L), N+P addition (0.4 mgNO3-N/L + 0.04 mgPO4/L), and 

N+P+Si (0.4 mg NO3-N/L + 0.04 mg PO4/L + 0.4 mg Si/L). For those pseudo-phytoplankton 

samples collected at the upper spring run, bioassays with additional nutrients did not produce an 

increase in growth compared to controls with ambient nutrient levels in 61% of the experiments, 

suggesting that the majority of locations were not nutrient limited. In the balance of 

experiments, 19% were phosphorus limited, 10% were nitrogen limited and 10% were co-limited 
 276 

by nitrogen and phosphorus. Given that all four of these spring ecosystems are supplied by 

waters with elevated nitrate concentrations relative 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 concentrations 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 NO3-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 nitrogen limitation in 75% of the experiments. The 

data from the Frazer et al. (2002) study suggest that when spring run nitrate concentrations 

exceed approximately 0.4 mg/L, nitrogen limitation is unlikely. 

A second springs’ study utilizing a pseudo-phytoplankton bioassay from material collected from 

Kings Bay, which contains multiple spring vents, was completed by Saindon (2005). This 

researcher compared pseudo-phytoplankton growth rates for experimentally 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. Saindon (2005) observed significant differences 

between sites in the relative growth rates of pseudo-phytoplankton, and that maximum biomass 

was influenced by both nutrient ratios 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) work was that maximum pseudo-phytoplankton 

biomass could be maintained at very low ambient 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, maximum pseudo-phytoplankton biomass appeared to 

increase with increasing nutrient availability. By way of comparison, 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, suggesting 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 reported 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 have (or had) a natural diversity of macroalgae 

(Whitford 1956). The mere presence of these species 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 macroalgae 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 positively correlated to nitrogen and phosphorus concentrations with a 
 277 

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 concentrations and indices of human activities 

within 1,000 m of the sampling site. Several possible explanations for the lack of better 

correlations between ambient nutrient concentrations and macroalgae abundance are that other 

factors limit and influence the standing crop. Principally, instantaneous field measures are 

incapable of incorporating prior environmental conditions which have caused the current state. 

Other important forcing functions affecting macroalgae were not measured as well, for instance 

in spring runs with strong flow and heavy recreational 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 Canfield 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 nitrate 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 matrix of nitrogen and phosphorus treatments 

maintained under relatively optimal growth conditions. From these experiments it appears 

nutrient concentrations affect the growth rate and ultimately the abundance of macroalgae as 

growth occurred in most nutrient treatments, even at very low nutrient concentrations, yet fastest 

growth rates occurred in high nutrient concentrations and were approximately two or more times 

greater than under low nutrient concentrations. Based on the bioassay experiments, it was 

preliminarily estimated that to initiate 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 are 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 category of primary producers which 

are commonly thought to respond to nutrient enrichment of spring water through an increase in 

their coverage, biomass, and/or occurrence of weedy 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 nutrient effects on periphyton abundance in a spring 
 278 

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, Florida. This study added nitrate, phosphate, and a 

combination of nitrate and phosphate to the upstream end of flow-through tubes containing glass 

microscope slides in the spring run. This experimental design allowed the amount of periphyton 

colonizing glass microscope slides to be compared between nutrient treatments and with a 

control treatment over a one month period. Increased epiphyte abundance (as measured by 

chlorophyll per unit area) was observed following the addition of phosphorus, as well as 

phosphorus in combination with nitrogen, when compared to the control (no nutrient addition 

treatment). The results suggested that phosphorus was the primary nutrient limiting periphyton 

abundance in the upper spring fed portion of this system, however nitrate additions supported 

additional periphyton growth over the control treatment and therefore nitrogen appeared to be 

secondarily limiting (Notestein et al. 2003). The Chassahowitzka spring system had elevated 

nitrate concentrations, approximately 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 absolutely limiting for it to have a biological effect 

in a spring ecosystem. 

In stream systems outside of Florida, there have been numerous publications on the relations 

between periphyton abundance and nutrients (e.g. Dodds et al. 1997, Francoeur 2001). 

Interestingly, there do not appear to be consistent trends or factors which control periphyton 

abundance in stream systems, which may be the case with Florida’s spring systems as well. As 

the following interstate examples show, increased nutrient concentrations can influence 

periphyton abundance, but the responses to nutrient 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 nitrogen, has been shown to increase periphyton 

abundance (Elwood et al. 1981, Peterson et al. 1993, Bothwell 1985). Other 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 increase periphyton abundance. Another factor that 

interacts with periphyton abundance 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 controls originating from the bottom up (i.e. light availability) or 

top down (i.e. grazers controls) and suggest that the factors controlling periphyton abundance in 

Florida’s spring systems needs further research, likely at the level of in situ complex mesocosms 

that can colonize with realistic populations and diversity of algae and invertebrates. 
 279 

Vascular Plants 

Studies of two Florida springs systems, the Chassahowitzka (Notestein 2001) and the 

Ichetucknee (Kurz et al. 2004), involved measuring the abundance of submersed vascular 

macrophytes and the concentration of water column nutrients. Chassahowitzka had ambient 

water column nitrate concentrations of approximately 0.45 mg/L and phosphate concentrations 

of 0.016 mg/L; while Ichetucknee had approximately 0.55 mg/L nitrate and 0.04 mg/L phosphate 

in the water column. These measurements are elevated compared to historic concentrations. 

Neither study showed a correlation between surface water nutrient concentrations and the growth 

rate and/or abundance of submersed aquatic macrophytes, which appears to be consistent with 

research from other non-spring fed stream systems. 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 between the field growth of P. pectinatus 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 which 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 ecosystems with current average 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 nutrient concentrations and submersed vascular 

macrophytes. 

Much concern over the apparent reduction from the historical abundance of submersed vascular 

plants in spring ecosystems exists, and is warranted due to the important habitat and the food 

web contributions these primary producers make (see Chapters 1 and 4). As noted in the section 

on periphyton above, the potential causes for declines in submersed vascular plant abundance 

could include resource competition from periphyton or filamentous and benthic algae, 

overgrazing by consumers, physical disturbance from recreation, and/or disease; 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 related 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 observed in 98% of Florida springs) can significantly 

reduce the growth of ammonia-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 the levels of organic anions within those species 

adapted to ammonia uptake by the roots (such as P. alpinus). This may explain the observed 
 280 

reduction in growth under high water column nitrate concentrations, although further 

experiments are needed to assess potential negative effects on other species. 

Primary Consumers 

Invertebrates 

Nitrate pollution in Florida springs has been implicated 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 concentration measurements 

in six Florida springs and conducted 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 nitrate concentrations. LC50 levels (the concentration at which 50% 

mortality occurred) could not be determined despite 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 EC50 (the effective concentration of a 

compound where 50% of its maximal effect is observed) for these apple 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 changes due to non-indigenous plant invasions, might explain the 

observed declines in apple snail abundance (Corrao et al. 2006). 

Studies on the effects of nitrate 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 the impact of nitrate 

must be assessed on a case by case basis. 

Vertebrates 

Studies examining the effects of nitrate on the reproduction and development of two vertebrate 

species found in Florida spring ecosystems, have been made for the Southern toad (Bufo 

terrestris) by Edwards et al. (2006) and the Eastern mosquitofish (Gambusia holbrooki) by 

Edwards and Guillette (2007). These researchers reported 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 concentration increased while in experiments utilizing 

spring water, tadpoles reared in high nitrate (30 mg/L) delayed metamorphosis, suggesting that 

water source was a factor besides nitrate concentration affecting Bufo development (Edwards et 

al. 2006). In their study of adult male mosquito fish collected from springs with varying nitrate 

concentrations, Edwards and Guillette (2007) observed that nitrate concentration (up to 5 mg/L) 

were significantly correlated with 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 ecosystems 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 data were compared to 

ambient nitrate nitrogen concentrations reported from fifteen of those springs. Although there 

were trends for lower fish and mussel species numbers 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-five species in the 1940s, 

Walsh and Williams (2003) collected twenty-nine species in 2002, and Munch et al. (2006) 

observed thirty-three species in 2004-05. Although the number of fish species doesn’t appear to 

have declined in correlation with increased nitrate, 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 indirectly affected by elevated nitrate concentrations 

and their apparent contribution to plant community 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 fish, 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 when designing experimental 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 hypothetical. If elevated nitrate concentrations 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 nitrate 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 which 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 professor, personal communication 2004) has 

suggested that if pregnant manatee 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 contamination of ground water and spring ecosystems in 

Florida. 
 282 

Humans 

A number of springs are utilized as a source for commercially bottled water (e.g. Gainer Springs 

in Bay County, Ginnie Springs in Alachua County, and Silver Springs in Marion County (Scott 

et al. 2002)). Given the widespread occurrence and generally increasing magnitude of nitrate 

contamination in Florida’s springs, there may be the potential of spring water 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 the average daily intake 

from all sources estimated at 75 milligrams. In healthy adults, about 5% is converted to nitrite 

by bacteria in saliva and additional bacterial conversion of nitrate 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 conversion of nitrate to nitrite. This is especially a 

concern for infants, whose gastrointestinal systems 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 nitrite in the human digestive system can in turn 

react with hemoglobin in the blood and creating methemoglobin, a form of the protein 

hemoglobin that is unable to transport oxygen. This results in a decreased capacity of the blood 

to transport oxygen causing a condition 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 adults are susceptible to long-term exposure to 

lower levels of nitrates and nitrites which can cause diuresis (an increase in the amount of urine, 

and starchy deposits and hemorrhaging of the spleen). Given the potential for human health 

impacts, the USEPA has designated a drinking water 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 greater 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, secondary, etc. heterotrophs and decomposers) 

which collectively comprise the living component of spring ecosystems (see Chapter 1). 

Because primary producers generate oxygen through photosynthesis and both producers and 

consumers use oxygen through respiration, 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 and 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, which 
 283 

shows average daily DO change at Silver Springs, 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 gross 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 1,200-m station due to the addition of the Back 

Channel Extension in the late 1970s. Estimated GPP rates corrected for the total aquatic area 

declined from about 15.7 to 15.6 g O2/m2

/d in the 1950s and the late 1970s, to about 11.2 g 

O2/m2

/d during 2004/2005 (from Munch et al. 2006). 

Gross and Net Productivity 

Odum (1957a) measured nitrate and gross primary productivity (GPP) in eleven Florida springs 

in 1955, which reveal a weak but negative relationship between nitrate concentration and GPP 

(R2

 = 0.14). The recent findings of lower GPP in Silver Springs at higher nitrate nitrogen 

concentrations are consistent with the patterns revealed by Odum’s earlier research (Table 1). A 

more recent study of community metabolism has been completed as part of the Wekiva River 

Pollutant Load Reduction Goal evaluation conducted 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 levels (WSI 2007a). The Wekiva River and Rock 

Springs Run were found to have higher nitrate nitrogen concentrations (averaging 0.69 and 0.84 
 284 

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 Springs were significantly 

higher than downstream concentrations (Wekiva – 1.1 vs. 0.36 mg/L; Rock Springs Run – 1.1 vs. 

0.55 mg/L, respectively). Gross primary productivity and PAR efficiency were found to be 

significantly lower in all of these spring-fed streams with higher nitrate nitrogen concentrations 

(Figure 3 and Table 2, WSI 2007a). This trend is consistent with the observations made at Silver 

Springs during the fifty-year retrospective study described by Munch et al. (2006). 

Table 1. A comparison of gross primary productivity (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 

(gO2/m2

/d) Date GPP (gO2/m2/d) Date GPP (gO2/m2/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 11.2 
 285 

y = -3.8294x + 5.7311

R2

 = 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 Primary

Productivity (g O2/m2/d)

WR-1

WR-2

RSR-1

RSR-2

ASC-1

JC-1

y = -3.4485x - 1.5075

R2

 = 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 Primary

Productivity (g O2/m2/d)

WR-1

WR-2

RSR-1

RSR-2

ASC-1

JC-1

y = -0.3993x + 7.2451

R2

 = 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

R2

 = 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

R2

 = 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 Creek, JC is Juniper Creek, RSR is Rock Springs 

Run (1 is upstream, 2 is downstream), WR is Wekiva 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 reference spring runs from, Alexander Springs and 

Juniper Springs (from WSI 2007a). 

Parameter Units Wekiwa Springs Rock Springs Reference Springs 

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 

O2/m2

/d 

2.12 5.39 3.76 1.41 1.00 1.20 5.71 5.67 5.69 

Community 

NPP 

g 

O2/m2

/d 

-3.50 -3.78 -3.64 -5.85 -8.59 -7.22 -0.39 -0.52 -0.46 

Community 

Respiration 

g 

O2/m2

/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 

O2/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 dissolved inorganic nitrogen concentrations less 

than 0.15 mg/L and soluble reactive phosphorus concentrations of 0.014 mg/L or less. These 

streams had GPP estimates ranging from less than 0.1 to 15 g O2/m2

/d, community respiration 

(CR) estimates ranging from about 2.0 to 11.1 g O2/m2/d, and photosynthetically active radiation 

(PAR) efficiencies ranging from 0.042 to 0.45 g O2/mol. Mullholland et al. (2001) reported that 

GPP and NPP were not correlated with either dissolved inorganic nitrogen or soluble reactive 

phosphorus concentrations but were significantly correlated with PAR, although CR was 

correlated with soluble reactive phosphorus concentrations. 

Other estimates of primary production from spring ecosystems include those made by Duarte and 

Canfield (1990a), who published plant community and productivity data for thirty-one Florida 

springs. Productivity data were not directly comparable to data collected by Odum (1957a, b), 

Knight (1980), WSI (2006), or Munch et al. (2006) as they were based on a short term, rapid 

assessment technique rather than on a full diurnal 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 productivity were correlated 

with degree of shading by shoreline vegetation. These results suggest that light may have been 

the primary limiting factor during their study. 

Respiration 

Changes in community respiration are likely to be influenced by nutrients if nutrients are 

influencing the abundance or productivity of primary producers. Changes in primary producers 

would in turn affect the consumer or decomposer community and hence the respiration of this 

part of the ecosystem. The practical differentiation of autotrophic versus heterotrophic 

respiration is not feasible in field studies. From Florida spring studies that have compared 

community respiration rates to water column nutrient 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 increase downstream in the Wekiva River and in 

Rock Springs Run as nitrate concentrations decreased. Community respiration also declined at 

Silver Springs at the same time nitrate nitrogen concentrations were increasing (Munch et al.

2006). 

Productivity/Respiration Ratio 

The productivity/respiration (P/R) ratio is commonly used to classify ecosystems as either 

autotrophic (P/R > 1) or heterotrophic (P/R < 1) depending on whether 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 respectively. 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 being 1.26 in June 2004 and a minimum monthly 

average of 0.73 in February 2004. In a comparison of spring run segments with nitrate 

concentrations ranging from natural ambient levels (< 0.05 mg/L) to elevated levels (> 1.0 

mg/L), P/R ratios were found to decline significantly in relation to increasing nitrate 

concentrations (Figure 3, WSI 2007a). This negative correlation suggests that the P/R ratio can 

be used to evaluate nutrient impairment of spring ecosystems. 

Photosynthetic Efficiency 

The efficiencies of photosynthetic primary producers have been measured most comprehensively 

at Silver Springs, where Odum (1957b) reported an average value of 1.09 g O2/mol of PAR 

(ranging from 0.79 to 1.53), while Knight (1980) reported an average value of 1.06 g O2/mol of 

PAR (ranging from 0.61 to 1.50), and Munch et al. (2006) reported an average value of 0.95 g 

O2/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 O2/mol, is equivalent to an estimated PAR efficiency of 7.6% 

using estimated conversion factors and PAR efficiency was observed to vary seasonally with the 
 288 

highest monthly average value of 10.5% (1.30 g O2/mol) measured in December 2004 and the 

lowest value of 5.5% (0.68 g O2/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 number 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 O2/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 efficiency correlated with increasing nitrate concentrations. 

Community Metabolism Conclusions 

As the previous examples illustrate, traditional measures of describing community structure 

which use biomass metrics allow few generalizations 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 the ecosystem components and may 

reduce variance caused by site-specific conditions. The observations summarized in this chapter 

of reduced primary productivity and ecological efficiency in correlation with increasing nutrients 

may have multiple explanations. 

First, Florida’s spring run ecosystems were likely adapted to naturally lower ambient nutrient 

conditions and higher nutrient levels could represent an imbalance 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 photosynthesis resulting from increased epiphytic algae 

growth and shading as a result of elevated water 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 photosynthetic rates at typical midday 

intensities As spring waters are generally clear and light transmission is excellent (Duarte and 

Canfield 1990b), epiphytic reductions in light availability to vascular plants may not reach the 

level of causing host plant death. In an experimental manipulation of epiphytes and light levels, 

Asaeda et al. (2004) noted that both epiphytic algae and low light reduce growth 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 algae-created shade occurred on the leaf surface. 

A third alternative explanation to reduced primary productivity in macrophytes in relation to 

elevated nitrate concentrations was provided by Boedeltje et al. (2005) who reported high water 

column nitrate nitrogen concentrations can 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 than ammonia uptake and assimilation (Runge 
 289 

1986, Marschner 1995). Boedeltje et al. (2005) hypothesized that nitrate-dominated nitrogen 

assimilation may lead to strong metabolic disturbances in species 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 are needed to assess potential 

negative effects on species found in the springs of Florida. 

Although the causative mechanisms which result in reduced primary productivity and ecological 

efficiency in correlation with increasing water column nutrient concentrations (especially nitrate 

and perhaps phosphorus) remain unknown; the averages, 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 other springs, spring runs, and 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 readily apparent that people prefer to swim in clear, clean water rather 

than turbid, colored, or phytoplankton-rich waters. It is likely that highly abundant macrophytes 

discourage water recreation as well. This presents a dilemma, in that, native benthic vascular 

plants represent a natural and ecologically important 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 visitors 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 perceived diminished appeal of a spring system, there 

would be associated declines in the economic impacts generated. Preservation of spring 

ecosystems should not be based solely on an economic basis, unless considerations for 

environmental services are factored in as well, but they provide support towards these systems 

protection. At this time insufficient quantitative public use data exist from most springs to 

develop correlations with increasing nutrient concentrations. 

Effects of Eutrophication on Human Recreation 

The type of recreation will obviously influence tolerance for the presence of SAV, with water 

dependent recreation perhaps having lower tolerance 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 exists, but Holland and Cichra (1994) reported that 

what respondents liked best about their trip to Rainbow River was: clear, clean water (35%), 

scenery (19%), quiet, relaxing (17%), and seeing wildlife and fish (11%) and that what 
 290 

respondents liked least about their trip: rainy weather (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 aquatic 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 recreational satisfaction. Similarly, increases in the abundance of 

vascular plants or macroalgae are likely to be viewed negatively by recreational users of spring 

ecosystems, given that people do not like to swim 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 perceived 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 these 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 recreation, and colonization by 

non-indigenous species. However, the widespread increase in nitrate 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 for observed changes in ecosystem metabolism. Strong (2004) 

examined temporal trends in water quality characteristics in 109 Florida springs. Parameters that 

were generally found to be increasing were specific conductance, alkalinity, hardness, and 

nitrate. Only pH concentrations were observed to be declining. There were no apparent trends 

in phosphorus concentrations. Mean nitrate concentrations in spring 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 declining water quality of springs are occurring. An 

illustration showing the range of nitrate plus nitrite concentrations (NOx-N mg/L, although nitrite 

is generally a minimal component of these aggregate values) for 130 Florida springs compiled 

from 2006 Florida Geological Survey data is shown in Figure 4. This figure presents the wide 

range of NOX-N concentrations that Florida springs exhibit and suggests that 83% of the 

surveyed springs have NOX-N concentrations greater than 0.1 mg/L, which is a commonly 

assumed background level for groundwater nitrate levels in Florida. 
 291 

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Spring ID

Average NO x-N (mg/L)

Stats NOx-N

Average 0.803

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 Mill 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 (Hillsborough) 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 NOX-N concentrations (mg/L) for 130 Florida springs (data 

from Florida Geological Survey, Scott et al. 2004). 

A similar illustration showing the range of total phosphorus concentrations (TP mg/L), for 130 

Florida springs compiled from 2006 Florida Geological Survey data is shown in Figure 5. 

Although total phosphorus (the aggregate of all forms of phosphorus) data are presented, in the 

case of groundwater, this total phosphorus is almost entirely composed of dissolved inorganic 

forms of phosphorus such as orthophosphate or soluble reactive phosphorus. This figure 

suggests that the total phosphorus concentrations of spring water are less variable than the wide 

range of NOx-N concentrations that Florida springs exhibit. Groundwater phosphorus 

concentrations are largely a function of the 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 separating mesotrophic and eutrophic lakes (Table 3, 

Forsberg and Ryding 1980). Although other indices of trophic status exist (e.g. the Florida 

Trophic State Index created by the Florida Department of Environmental Protection), their 

application to lotic systems has been problematic. This is principally due to the index’s inability 

to incorporate macrophytes and benthic algae that are not accounted for by standard methods 

which measure water column chlorophyll content in non-flowing systems. 
 292 

Because phosphorus levels have not been increasing in Florida’s springs, even though the 

groundwater is regionally rich in phosphorus, (Strong 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 NOx) and the resulting nitrogen to phosphorus ratios have 

become skewed to favor phosphorus limitation. This 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) relative to historic ambient conditions, 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 are largely beyond human control. Nitrate loading from 

anthropogenic activities can, however, be moderated and hence represents our best ability to 

effect a nutrient reduction strategy to benefit Florida spring ecosystems. 

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Spring ID

Average TP (mg/L)

Stats TP

Average 0.084

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 Ellaville Spring 92 Suwannee 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 Fanning 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 Mill Pond 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 total 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 system 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 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 

Springs ecosystem may be considerably less productive than it was fifty years ago (Munch et al.

2006). This appears to be a consequence of lowered gross primary productivity and efficiency of 

light utilization. This result appears to be counter-intuitive to the observation that nutrient levels 

(nitrate nitrogen) which support plant growth have increased in Silver Springs. 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 demands 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 maximize community metabolism, and the relatively 

recent increase of nutrients in the form of nitrate may have resulted in a stress rather than a 

subsidy (E.P. Odum et al. 1979). The subsidy-stress hypothesis may be the best model to 

describe the response of spring ecosystems to increasing nitrate concentrations; however, 

controlled field experiments need 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 retrospective study of Silver Springs by Munch et al.

(2006). The key findings from that 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% lower 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 similar to those observed for epiphytes and 
 294 

submersed aquatic plants. However, it is important to note that these estimates may not 

be an accurate indicator of primary productivity attributable to benthic mats, since mat 

biomass includes large proportions of bacteria, 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 records at Silver River. 

• 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 Knight’s 1978-79 study (Knight 1980). These declines in biomass were 

primarily due to large reductions in a few species (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 O2/m2

/d) in the 1950s and late 1970s to about 11.2 g 

O2/m2/d during the current study, a decline of about 27%. 

• Community respiration also declined from about 14.8 g O2/m2/d during the earlier studies 

to about 10.9 g O2/m2

/d during the Munch et al. (2006) study, a 26% reduction. 

• The resulting net community primary productivity declined from about 1.0 g O2/m2/d in 

the 1950s, to 0.80 g O2/m2/d in the late 1970s, to about 0.42 g O2/m2/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 O2/mol) of 

Photosynthetically Active Radiation (PAR) during Odum’s study to about 0.94 g O2/mol 

of PAR during the Munch et al. (2006) study, a decline of about 13%. 

• Average particulate export rates were found to be 72% lower during the Munch et al.

(2006) study compared to data published by Odum (1957). These particulates are a 

representation of the net system productivity, and their decline strengthens the evidence 

that overall productivity has declined in Silver Springs. 

The Munch et al. (2006) study observed that the percentage of developed lands within the 

springshed (5 mile radius) was positively correlated with actual nitrate concentrations and spring 

discharge was found to be inversely correlated with nitrate. The resulting correlation 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 concentration of Silver Springs water based on 

projected changes in land use and corresponding nitrogen loading 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 working in concert or at 

odds. Four possible hypotheses were offered 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 input of solar energy may have lowered GPP 

and PAR efficiency. 

• GPP may have declined as areas of optimal 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 velocities due to natural climatic conditions or 

consumptive groundwater uses may have reduced the previous subsidy needed for 

maximum plant/periphyton growth and 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 over the past 50 years. These factors include 

reductions in the volume of spring discharge, increases in riparian shading, reductions in the 

abundance of herbivorous fish, and increases in nitrate concentration and load. The authors of 
 296 

the Munch et al. (2006) study suggest that the corresponding increases in nitrate concentration 

over this time period cannot exclusively be responsible 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 remains warranted. Support for the concept that 

nitrate can act as both a subsidy at low concentration 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 measures of ecosystem productivity and nitrate 

concentrations. Based on these lines of available evidence, nitrate alone may not be the sole 

factor negatively impacting Florida spring ecosystems, but it is the single most important water 

quality parameter that has changed during the period of observed declines in spring ecosystem 

structure and function. 

RECOMMENDATIONS FOR FURTHER RESEARCH 

Despite a wide variety of springs’ research, there remains a significant knowledge gap between 

the real and perceived threats that nitrate pollution 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-term studies of nitrogen additions in 

lotic ecosystems and clear distinctions need to be made between ecosystem responses to shortterm or periodic increases in nitrogen loading and alterations in ecosystem functions due to 

chronic nitrogen loading.” In addition to the general research questions to be answered, 

Florida’s springs represent important cultural and economic resources which also provide realtime windows into the condition of Florida’s most essential natural resource – high-quality 

groundwater. 

With these caveats in mind, we suggest that nutrients are not the only stressors of concern in 

spring ecosystems. Recent work in the headwater areas of Wekiwa and Rock springs indicates 

that ecosystem metabolism can be negatively impacted 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 invasions from non-indigenous plant and animal 

species, aquatic weed management activity, erosion and deposition of sediments, and increasing 

stormwater inflows from surrounding urbanized areas. Future research to be conducted should 

attempt to effectively understand the separate and synergistic effects of these multiple forcing 

functions and the associated responses of our spring ecosystems. 

With the multiple stressors affecting spring ecosystems, any metric used to monitor their health 

should to the fullest extent possible 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 productivity, community respiration, productivity to 

respiration ratios, and photosynthetic efficiency. These metrics can be relatively easily estimated 

using upstream-downstream changes in dissolved oxygen concentrations and have the best 

ability to describe ecosystem level functions, and their value is enhanced when combined with 

measures of fauna and flora abundance and water chemistry, other typical 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 agencies, but improvements in the spatial and 

temporal sampling could be realized as well as improved assembly, analysis, and 

reporting of the resulting data. 

• Continue to expand and refine 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 nitrogen loading per LULC category, assessment 

of springshed land use changes on the nutrient 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 and as a result to limit or prevent springshed 

pollution would be enhanced. 

• Better utilize current technology of in situ data loggers in a representative set of major 

springs ecosystems so as to provide regular or continuous records of changing water 

quality conditions, including temperature, dissolved oxygen, pH, and specific 

conductance (and other parameters including nutrients as instruments and technologies 

become available). These measures, particularly if made in conjunction with levels of 

ambient solar radiation and spring discharge would allow the direct estimation of 

ecosystem metabolism and photosynthetic efficiency. Given the inter-specific range of 

physical and chemical conditions that Florida springs currently exhibit, simultaneous 

estimates of ecosystem metabolism would allow development of a response surface for 

comparison to changing environmental conditions (forcing functions). For example, 

multivariate analysis could be applied to determining the relative importance of a number 

of forcing functions on ecosystem metabolism, including identification of the 

concentration or intensity of each variable (e.g., nitrate) as it transitions from a subsidy to 

a stressor. 

• Quantify invertebrate, vertebrate, and floral diversity and abundance on a three-to-five 

year schedule at 1st and 2nd magnitude springs so that trends in these ecosystem 

components can be observed. This sampling should be conducted on springs that are also 

being monitored for ecosystem metabolism and should be scheduled on a monthly basis 

for at least one year during each three-to-five -year period so as to elucidate seasonal 

trends. 

• Conduct in situ mesocosm experiments which quantify the role that bottom-up processes 

(i.e. nutrient limitation) versus 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 responses by the primary producer community. 
 298 

• Better monitor and inventory human use activities 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 the springs’ resource. Human recreation levels 

or uses which are identified to be impairing spring health or degrading spring aesthetics 

should be modified through adaptive management strategies. The most successful 

example of this balance may be Ichetucknee Springs State Park which annually 

accommodates hundreds of thousands of in water recreationalists while maintaining 

sustainable populations of submersed macrophytes. 

Recommendations 

Based on the synthesis of information provided concerning the effects of nutrients on springs in 

Florida, the following recommendations are provided: 

• It is recommended that the Florida Department of Environmental Protection (DEP) 

Springs Initiative continue funding applied research under field conditions 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 importance of artesian springs in Florida. 

• It is recommended that counties with significant springs’ resources utilize existing 

governmental agencies to develop springs protection plans that address such components 

as land development regulations and land acquisition, along with other groundwater and 

spring protection measures in order to prevent further and, if possible, reverse the 

observed increases in nitrate-nitrogen concentrations. Regulatory programs and tax 

incentives should be utilized to reduce nitrogen loads to the artesian 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 support from the Florida Legislature develop 

applicable standards that define nutrient impairment in artesian springs. Springs that meet 

these impairment metrics should be moved as quickly as possible into the state’s Total 

Maximum Daily Load (TMDL) program. Nitrate concentrations significantly above 

historical spring levels could be used as a criterion for listing under the Impaired Waters 

Rule. For springs that are above recommended thresholds for nitrate nitrogen 

concentrations, these TMDLs could provide a basis for enhanced nitrogen-removal 

standards for municipal and on-site wastewater facilities as well as for stormwater 

management systems and the active and pro-active enforcement of nitrogen mass loading 

limits by existing and future point and non-point sources in the affected springsheds. 
 299 

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7 

Managing Nutrient Inputs to Florida 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 pollution of Florida springs is potentially as 

broad as the scope of human activities that contribute nutrients to the springshed. Regulations on 

the discharge of wastewater are key elements. The federal Clean Water Act and Safe Drinking 

Water Act provide a set 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 

the National Pollutant Discharge Elimination System (NPDES). Discharges to groundwater may 

be regulated under this program if there is a “significant nexus” to the quality of navigable 

surface waters. The Safe Drinking Water Act protects public water supplies by establishing 

minimum criteria for drinking water quality and requiring states to regulate the underground 

injection of pollutants. Groundwater quality standards 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 implements 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 direct jurisdiction of the federal program, such as 

nonpoint sources, agricultural discharges, and 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 treatment plants are generally only required to use 

secondary treatment, except in areas 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 also a general limitation on the degradation of any waters and a 

prohibition on the degradation of designated Outstanding Florida Waters (OFW). A narrative 

nutrient standard prohibits altering nutrient concentrations “so as to cause an imbalance in 

natural populations of aquatic flora or fauna.” Numeric standards can also be developed, most 

commonly as Total Maximum Daily Loads (TMDL). 
 306 

TMDLs can be adopted by either EPA or the Florida Department of Environmental Protection 

(DEP). Florida’s process begins with determining whether a particular waterbody is “impaired” 

based on “objective and credible data, studies and reports” demonstrating it does not meet water 

quality standards for a specific criterion, such as nitrogen or phosphorus. If there is sufficient 

data demonstrating 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 equitably 

allocated to the various sources and basins contributing pollutants. The DEP then assembles a 

group of stakeholders to develop a Basin Management Action Plan (BMAP) to develop 

strategies for implementing the adopted TMDL. NPDES permits may be revised as necessary to 

implement the BMAP. Other dischargers may be required to reduce their discharge “to the 

maximum extent practicable”. Nonpoint sources must demonstrate compliance by either 

implementing Best Management Practices (BMP) or conducting water quality monitoring. There 

are additional powerful incentives for implementing 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 DEP or a water management district. 

Stormwater is a significant source of pollutants 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 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 stormwater 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 pollution. 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 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 nutrients to springs. They can provide for improved stormwater and wastewater 

management. The Florida Legislature has specifically required local governments in the Wekiva 

Study Area to amend their comprehensive plans to enhance springs protection. It has also limited 

the ability of local governments to regulate agricultural land uses and the conversion of 
 307 

agricultural land to urban uses. Local governments are liable for compensating landowners who 

show an “inordinate burden”. A Consumer Fertilizer Task Force has recommended new 

restrictions on the ability of local governments to regulate the application of fertilizers. 

REGULATION OF DISCHARGES 

Clean Water Act 

The Federal Water Pollution Control Act Amendments of 19723

 created one of our nation’s most 

comprehensive environmental statutes. It was enacted in response to widespread 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 waters4” through a combination of research, planning, 

construction and regulation. The Act was amended in 1977 and renamed the Clean Water Act5

(CWA), but the fundamental structure and policies have remained intact for over 35 years. 

The CWA prohibits any discharge of pollutants that is not authorized by a permit or otherwise in 

compliance with the Act6. Discharges may be authorized through a permit issued by EPA or a 

state under a program called the National Pollutant 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. Technology-based effluent limitations require the 

use of specified levels of technology 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, however, 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 from 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 discrete conveyance” of pollutants9. Point sources might include 

wastewater treatment plants, industrial sources, concentrated animal feeding operations, or 

stormwater systems. Agricultural stormwater and irrigation return flows are expressly exempted 

from regulation, even though they might have the physical characteristics of a point source10. A 

spring itself might be considered a point source under this definition 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). 

10Id. 
 308 

Although the term “navigable waters” was defined in the Act as “waters of the United States11”, 

the extent to which discharges to groundwater are subject to the CWA is not clear12. Several 

judicial opinions held that Congress intended to exclude groundwater from the scope of 

regulation under this statute13. Other decisions have found jurisdiction over groundwater that is 

hydrologically connected to surface waters14. 

The Supreme Court’s recent decision in Rapanos v. United States15 has further confused the 

scope of CWA jurisdiction in a case involving the 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 jurisdiction required a “substantial nexus” to waters that are 

navigable16. In interpreting these opinions, the 11th Circuit Court of Appeals recently held there 

might be no CWA jurisdiction over a discharge of pipe processing wastewater to a perennial 

stream that connected to navigable waters17. 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 chemical, physical or biological integrity of the navigable waters. 

A similar test would have to be met for any discharge of nutrients to groundwater to be subject to 

CWA regulation. There would have to be a relatively direct hydrologic connection to navigable 

waters and there would have to be proof that the discharge significantly affected the quality of 

 

1133 U.S.C. §1362(7) 

12See generally, Thomas L. Casey, III, Reevaluating “Isolated Waters”: Is Hydrologically 

Connected Groundwater “Navigable Water” Under the Clean Water Act?, 54 Ala. L. 

Rev. 159 (2002). 

13Rice v. Harken Exploration Co., 250 F.3d 264 (5th 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). 

14U.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 Situation (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). 

15126 S.Ct. 2208 (2006). 

16The requisite nexus could be shown if the subject wetlands “either alone or in combination 

with similarly situated lands in the region, significantly affect the chemical, physical, and 

biological integrity of other covered waters more readily understood as ‘navigable’”. 126 

S Ct. 2248. Those effects cannot be “speculative or insubstantial” Id. 

17U.S. v. Robison, 505 F.3d 1208 (11th Cir 2007). 
 309 

those surface waters. The 9th Circuit Court of Appeals recently found such a significant nexus in 

a case involving the discharge of wastewater to a pond adjacent to a river18. Because the 

wastewater seeped into the river through a groundwater connection, affected the quality of the 

surface water and had a significant ecological connection through bird, mammal and fish 

populations, it was a regulated water19. 

State Discharge Permitting 

The DEP has assumed primary responsibility from EPA for implementation of the NPDES 

permitting program20. For discharges to groundwater that do not have a sufficiently close 

connection to a surface water, however, only state 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 Control Act prohibits the construction or operation 

without a permit of any stationary installation reasonably expected to be a source of water 

pollution21. An “installation” is any “structure, 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 impairment of . . .waters, or alteration of the 

chemical, physical, biological, or radiological integrity of . . .water in quantities or at levels 

which are or may be potentially harmful or injurious to human health or welfare, animal or plant 

life, or property or which unreasonably interfere with the enjoyment of life or property, including 

outdoor recreation unless authorized by applicable law.23” Finally, “waters” “include, but are not 

limited to, rivers, lakes, streams, springs, impoundments, wetlands, and all other waters . . 

.including underground waters24.” The distinction in federal law between “point” and 

“nonpoint” sources and between “navigable waters” and groundwater is thus not very significant 

in Florida25. The state, unlike EPA, has clear authority to regulate virtually any discharge to any 

water. 

Several sections of Florida statutes authorize the DEP to adopt rules and issue permits for 

 

18N.Cal. Riverwatch v. City of Healdsburg, 496 F.3d 993 (9th Cir. 2007). 

19496 F.3d 1000-1001. 

20Fla. Stat. §303.0885 (2007); Rule 62-620, FAC. DEP assumed NPDES authority for industrial 

and domestic wastewater discharges in 1995 and stormwater discharges in 2000. See 

generally, Terry Cole, National Pollutant Discharge Elimination System Program, Ch. 

12.2, Florida Environmental and Land Use Law (June 2007). 

21Fla. Stat. §403.087(1) (2007). 

22Id., §403.031(1). 

23Id., §403.031(7) 

24Id., §403.031(13) 

25If a discharge is regulated under Florida law, but not subject to the Clean Water Act, however, 

then certain provisions adopted to allow the state to qualify for 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 pollution26. If wastewater is discharged directly to groundwater via a well, then an 

Underground Injection Control (UIC) permit is required27. Generally, dischargers are required to 

provide “reasonable assurance” that water quality standards will not be violated and that 

appropriate levels of treatment are being implemented. 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 standards, depending on 

where the effluent is discharged. Discharges to groundwater can be made through a variety of 

systems, with different levels of public access28. 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 Maximum 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 regulatory uncertainty and delay should be reduced in favor of imposing the 

most protective requirements. For example, wastewater discharged to such estuaries as Sarasota 

Bay, Tampa Bay, Charlotte Harbor and their tributaries must meet advanced waste treatment 

standards29. Special requirements have recently been imposed on domestic wastewater 

discharges in the Wekiva Study Area that may serve as a model for other areas contributing 

recharge to springs30. These requirements were based on recommendations developed pursuant 

to the Wekiva Parkway and Protection Act31. The DEP first determined that 0.2 mg/l should be 

adopted as an initial target for nitrogen concentrations at the spring vents, based on unimpacted 

reference sites. The DEP also mapped the recharge areas according to their relative vulnerability. 

 

26Id., §§403.061(14), .087(2), (5), .088(2) (2007). The rules governing permitting are at Fla. 

Admin. Code r. 62-4 (2007). 

27Fla. Admin. Code r. 62-528 (2007). 

28Rule 62-610 provides different criteria depending 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. 

29Fla. 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). 

30Fla. 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 Legislature, Id. §369.316, and includes large 

parts of the springsheds of Rock Spring, Wekiwa Spring and other springs feeding the 

Wekiva River. For the basis of recommendations specific to wastewater discharge see 

Florida Department of Environmental Protection, A Strategy for Water Quality 

Protection: Wastewater Treatment in the Wekiva Study Area (December 2004). For 

background information on the various planning and coordination activities 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, with 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 are allowed in areas of less 

vulnerability, for systems that have lower rates of application and for smaller systems whose 

costs of additional treatment are much higher. Dischargers can qualify for relief from these 

requirements by making “an affirmative demonstration, based on relevant water quality data, 

physical circumstances, or other credible information, 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 wastewater34. Industrial 

dischargers must comply with technology-based effluent limitations developed for various 

industrial categories and adopted as rules, as well water quality-based effluent limitations35. New 

sources must meet new source performance standards (NSPS). Existing sources are required to 

implement the best available technology economically achievable (BAT). For conventional 

pollutants such as Biological Oxygen Demanding materials (BOD5) or Total Suspended Solids 

(TSS), a less stringent effluent limitation is applicable, termed best conventional pollutant 

control (BCT). 

Concentrated Animal Feeding Operations (CAFO) are a particular type of industrial discharger 

that generates substantial quantities of nutrients36. The national regulations adopted by EPA have 

been challenged by both industrial and environmental interests37. Final revisions have not yet 

been adopted38. Florida will be required to implement the final federal rules, but meanwhile 

such facilities are governed by rule 40 C.F.R. pt. 412. Facilities that are not subject to NPDES 

permitting are governed by rule 62-670, F.A.C.39 

 

32Fla. Admin. Code r. 62-600.550(4)(a)(2007). Technically these are termed “rapid-rate or 

restricted access slow-rate land application systems”. 

33Id., 62-600.550(7). This provision appears to allow the discharger to demonstrate that the 

discharge will achieve compliance with the target goal, not to challenge the goal itself. 

34Id., 62-660.200(2)(e). 

35Id., 62-660.400. DEP has adopted by reference the rules developed and adopted by EPA under 

the Clean Water Act. 

36 See generally, James H. Andreasen, Concentrated Animal Feeding Operations: A Program in 

Transition, Natural Resources and Environment 45 (Spring 2007). CAFOs produce 500 

million tons of manure annually. Id. at 45. 

37Waterkeeper Alliance v. U.S. Envt’l Protection Agency, 399 F.3d 486 (2d Cir. 2005). 

3871 Fed. Reg. 37744 (June 30, 2006). The existing rules are at 40 C.F.R. pt 412. 

39The 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: Significant Cases in the 

Works, Section Reporter (March 2005). Environmental and Land Use Law Section, The 

Florida Bar. Available at http://www.eluls.org, visited February 23, 2008. DEP is 
 312 

WATER QUALITY STANDARDS 

Florida has established classifications and standards for both ground and surface water40. One of 

 the primary regulatory criteria for any discharge is that it not cause or contribute to violations of 

water quality standards41. 

Groundwater Quality Standards 

The DEP generally regulates discharges of pollutants to groundwater and has established a 

scheme of groundwater classification and standards.42 The standards for groundwater are 

primarily designed to protect human health43. For example, all groundwater must be “free from” 

substances in concentrations which are carcinogenic, mutagenic, teratogenic, or toxic to human 

beings44.. The only protection for non-human organisms in the groundwater is for plants, animals 

and organisms native to the soil that are responsible for treating the discharge. A groundwater 

discharge is also prohibited from causing violations of the acute toxicity standard for surface 

water if it reaches the surface45, 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 reactions by recreational users could 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 contaminant levels (MCL) for drinking water46. The 

only nutrient listed in the MCL is nitrate/nitrite which is limited to 10 mg/l47. Groundwater 

quality standards thus protect spring discharges against 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 draft rules 62-620, 62-621.500(3) & (3)(a). 

40Florida, 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. 

41Fla. Admin. Code r.62-4.030, -4.070, -4.240(4) (2007). 

42See generally, Cynthia Christen, Groundwater Protection: An Overview, Ch 11.1, Florida 

Environmental and Land Use Law (Feb 2002). 

43Fla. Admin. code ch 62-520 (2007). 

44Id., 62–520.400(1)(b). 

45Rule 62–302.200(1) is the acute toxicity standard. 

46Rules 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 consumption through water supply systems. EPA 

adopts Maximum Contaminant Levels (MCL) as minimum criteria for drinking water. 

The SDWA also requires the states to regulate the discharge of pollutants to underground 

sources of water and MCLs have become commonly used to limit the contamination of 

groundwater and provide criteria for the remediation of contaminated sites. 

47Fla. 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 spring, however, surface water quality standards 

apply and a discharge of nutrients to groundwater cannot be permitted if it would cause or 

contribute to a violation of surface water quality standards48. Florida’s surface water quality 

standards include a policy statement on nutrients49. 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 standards50. 

Most spring discharges in Florida 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 use51. 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 interfere 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. The minimum criteria prohibit discharge components 

which produce nuisance or toxic conditions53. To the extent that nutrient discharges could have 

such effects, they would violate the minimum standards. In that case they would also be likely to 

 

48 Id., 62-4.070; 62-302.300(15). 

49Id., 62-302.300(13) 

50The moderating provisions include exemptions from water quality standards, Rule 62-4.243, 

mixing zones, rule 62-4.244 and site specific alternative criteria, rule 62-302.800. 

Dischargers who are violating water quality standards can also continue to operate under 

the terms of an administrative order, apply for a variance or seek reclassification of the 

affected waters. See generally, Jennifer L. Fitzwater, Relief from Florida Water Quality 

Criteria, 12.4, Florida Environmental and Land Use Law (August 2003). 

51Fla. Stat. §403.021 (2007); Fla. Admin. Code r. 62-302.300(11), (14) (2007). 

52Id., 62-302.300(14). A discharger can, however, show that a particular use is not attainable and 

overcome this prohibition. 

53Fla. 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 District, alleging that operation of the water management system in 

South Florida caused an imbalance of flora and fauna56. In 1992 the lawsuit was settled57, but it 

was not until 2004 that a numeric criterion for phosphorus was finally adopted58. 

Florida’s nondegradation policy and implementing rules also control nutrient discharges. In 

addition to protecting existing and designated uses, the policy prohibits any 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, includes “Man-induced nutrient enrichment 

(total nitrogen or total phosphorus). . .60” Factors for determining whether 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 balancing these factors. 

A related nondegradation standard protects waters that are designated as Outstanding Florida 

Waters (OFW) and thus entitled to the highest levels of protection61. 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 water quality cannot be lowered as a result of the 

proposed activity or discharge62. The term “existing ambient water quality” is defined as the 

water quality that could reasonably have been expected, based on the best scientific information 

 

54Id., 62-302.530(46). 

55Id., 62-302.530(47)(b). 

56United States v. S. Fla. Water Management Dist., Case No. 88-1886 (S.D. Fla.). 

57United States v. S. Fla. Water Management 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 comprehensive 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). 

59Fla. Admin. Code r. 62-302.300(17)(2007). 

60Id., 62-302.530 (47) (a) (Table). 

61Id., 62-302.700, 62-4.242(2). A slightly more stringent rule applies to Outstanding National 

Resource Waters, which are given additional protection 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). 

62Id.,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 qualifications 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 prior to a permit application63. 

Specific numeric nutrient standards can be established in several ways. The Legislature required 

the DEP to establish a numerical criterion for phosphorus in the Everglades64. 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 

development65 Most numeric criteria, however, are being 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 state66. 

TOTAL MAXIMUM DAILY LOADS 

The Clean Water Act required the states to assess their waters to identify those areas where 

compliance with technology-based effluent limitations is not sufficient to achieve water quality 

standards67. For those waters, the state is required to establish a total maximum daily load 

(TMDL) for relevant pollutants at a level sufficient 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 force compliance with the statutory mandate68. The result in 

Florida was a 1999 consent decree requiring EPA to establish TMDLs for almost 2000 

waterbodies by 201169. The Florida Legislature responded by enacting the Florida Watershed 

Protection Act70, which established a process for developing and implementing TMDLs by the 

state. Florida now has a very comprehensive process for assessing waterbodies, establishing 

TMDLs and developing Basin Management Action Plans (BMAP) to address water quality 

violations71. TMDLs and BMAPs are being developed to address nutrient pollution in many 

areas. Whenever the state fails to develop a TMDL that meets federal criteria or the requirements 

of the consent decree, then EPA must establish one72. The Florida Legislature has directed the 

DEP to take action to develop and implement TMDLs for several specific waterbodies, including 

 

63Id.,62-4.242(2)© 

64Fla. Stat. §§373.4592(2)(m), (4)(e)(2007). The statute 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. 

65Fla. Admin. Code r. 62-40.210 (20), .430 (2007). The Districts are authorized to develop 

strategies for meeting water quality standards in Surface Water Improvement and 

Management (SWIM) plans. Fla. Stat. §373.453(2)(e) (2007). 

66http://www.dep.state.fl.us/water/wqssp/nutrients/index.htm. 

6733 U.S.C. §1313(d) (2007). 

68OLIVER 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. 

69Florida Wildlife Federation v. Carol Browner, Case No. 4:98CV356-WS (N.D.Fla.). 

70Fla. Stat. § 403.067 (2007). 

71DEP TMDL documents are available at http://www.dep.state.fl.us/water/tmdl/index.htm. 

72EPA TMDL documents are accessible at http://www.epa.gov/OWOW/TMDL/. 
 316 

Lake Okeechobee and its tributaries73, the Caloosahatchee74, the St Lucie75 and the Wekiva76. 

Legislation has been proposed to require the development of TMDLs for additional springs77. 

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 begins by determining whether there 

is sufficient quality data to prove that a particular waterbody is “impaired”, as defined in the 

impaired waters rule78, to justify placement on the planning 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 criterion79. The 

rule establishes stringent quality standards for the data that can be used in this assessment. If 

numeric criteria have been established, then narrative or biological criteria cannot be used as the 

basis for showing impairment unless the DEP demonstrates the numeric criteria are not 

adequate80. 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 mean chlorophyll a values shall be the 

primary means for assessing whether a water should be assessed further for 

nutrient impairment. Other information indicating an imbalance in flora or fauna 

due to nutrient enrichment, including, but not limited to, algal blooms, excessive 

macrophyte growth, decrease in the distribution (either in density or areal 

coverage) of seagrasses or other submerged aquatic vegetation, changes in algal 

species richness, and excessive diel oxygen swings, shall also be considered82. 

Specific nutrient criteria are also specified for streams83, lakes84 and estuaries85. 

 

73Fla. Stat. §§ 373.4595(1)(e)-(h),(l), ©, (4) (2007). 

74Id., § 373.4595(5). 

75Id.,§ 373.4595(l), ©, (4). 

76Id., § 369.318(1), (8). 

77Proposed Florida Springs Protection Act, SB 2394 (02/28/2008). Earlier proposed legislation 

also relied on a modified TMDL process. 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 Clinic at the University of Florida Levin 

College of Law. See http://www.law.ufl.edu/conservation/springs_narrative.shtml 

(visited April 22, 2008). 

78Fla. Admin. Code r. 62-303 (2007). 

79Fla. Stat. § 403.067(3)(b) (2007). 

80Id., ©. 

81Id. 82 Fla. Admin. Code r. 62-303.350(1) (2007). 

83Id., 62-303.351. 

84Id., 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 nutrient enrichment, the DEP must have five years of data sufficient 

to develop a site-specific threshold for when an imbalance in flora or fauna occurs or to apply the 

generic thresholds for streams, lake and estuaries86. 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) causing the impairment.87” For waters impaired by nutrients, the DEP must 

identify whether it is limited by nitrogen, phosphorus or both88. 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 attainment of applicable water quality standards” 

in the future, with “reasonable progress” by the next cycle89. The verified list must be submitted 

to EPA for review and approval and is subject to challenge in state administrative courts as a 

final order90. 

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 standards91. In calculating 

the TMDL, the DEP must account for seasonal variations and “include a margin of safety that 

takes into account any lack of knowledge concerning the relationship between effluent 

limitations and water quality92.” The calculated load must then be allocated to the various 

sources contributing pollutants. The DEP may allocate loadings between point and nonpoint 

sources, among basins and among sources. The allocation must be “reasonable and equitable” 

and based on consideration of ten listed factors93. TMDLs are adopted as rules by the Secretary. 

Detailed allocations may be left to the Basin Management Action Plan. 

 85Id., 62-303.353. Open coastal waters are also included. 

86Id., 62-303.450. 

87Id., 62-303.710(1). 

88Id., 62-303.710(4). 

89Id., 62-303.600. An example of such a case is the Suwannee River Partnership. Ayşin 

Dedekorkut, Suwannee River Partnership: Representation Instead of Regulation, Ch 2 in 

JOHN T. SCHOLTZ AND BRUCE STIFTEL, EDS., ADAPTIVE GOVERNANCE 

AND WATER CONFLICT: NEW INSTITUTIONS FOR COLLABORATIVE 

PLANNING (2005). Resources for the Future, Washington, D.C. ISBN 1-933115-18-1. 

90Winston K. Borkowski, Total Maximum Daily Loads in Florida– The New Millennium, 12.5-5 

in Florida Environmental and Land Use Law (June 2007). 

91Fla. Stat. § 403.067(6)(a)2 (2007). 

92Id. 

93Id., §403.067(6)(b). 
 318 

Basin Management Action Plans (BMAP) 

Basin Management Action Plans (BMAP) are the foundation for implementation of an adopted 

TMDL. BMAPs are intended to integrate existing water management strategies to achieve water 

quality standards and equitably allocate pollutant reductions. They must include an 

implementation schedule, a basis for evaluating the plan’s effectiveness, feasible funding 

strategies and mechanisms to address future increases in pollutant loading. Phased 

implementation is allowed. Credits may be granted to sources that reduced pollutant loads prior 

to development of the BMAP. The plan must include provisions for monitoring water quality and 

revising the plan in response to the results94. 

 

The BMAP may include such nonregulatory measures as public works construction, land 

acquisition, education, waste minimization, pollution prevention, interagency agreements and 

further planning95. The BMAP must also affect permitting. NPDES permits may be reopened to 

incorporate conditions to implement the plan96. 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 pollutant 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 dischargers identified in the BMAP must demonstrate 

compliance by either implementing best management practices (BMP) or conducting water 

quality monitoring99. A further incentive for implementing BMPs is that the discharger 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 Assurance Trust Fund for contamination-related 

expenses101. 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 authority to adopt them for “agricultural pollutant sources”. The 

 

94Id., §403.067(7)(a)5. 

95Id., §403.067(7)(b)1.b-f. 

96Id., §403.067(7)(b)2.a. 

97Id, §403.067(7)2.b. 

98Id, §403.067(7)2.b.f. 

99Id., §403.067(7)(b)2.g. 

100Id., §403.067(7)(b)2.i. 

101Id., §403.067(7)(c)3. The shield also extends to research projects that are developing or 

demonstrating BMPs. Id. 

102Id., §403.067(7)©. The term “BMP” means what this section actually refers 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 verified at representative sites by the DEP103. An initial 

verification based on “best professional judgement” that BMPs are “reasonably expected to be 

effective” must be made by the DEP before BMPs are adopted as rules104. 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 period105. 

STORMWATER 

Stormwater conveys nutrients to both ground and surface waters, and one of the objectives of 

stormwater regulation has been to control those inputs. In Florida, stormwater may be regulated 

by any combination of the DEP, a water management 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 regulate the construction of 

new stormwater systems106. Florida’s water management districts also exercised authority over 

the construction and operation of surface water management systems under Part IV of the Water 

Resources Act of 1972. To streamline permitting and eliminate duplication, 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 district107. Older 

permitting criteria still apply, however, to many existing developments or ongoing vested 

development activities. A further complexity stems 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 effectiveness in controlling nutrients, the DEP has begun the 

process of developing a new statewide stormwater rule108. 

The Water Resource Implementation Rule sets forth the goals for stormwater permitting109. In 

addition to controlling pollution, the program is intended to maintain and restore salinity and 

flow regimes and groundwater recharge. It established the policy that if a stormwater system is 

designed to meet the criteria of the relevant rules, there is a rebuttable presumption that it will 

 

103Id., 403.067(7)(c)3 

104Id. 

105Id., 403.067(7)(c)4 

106Fla. Admin. Code r. 62-35 (2007). The Department of Environmental Regulation (DER), 

DEP’s predecessor, implemented the rule. 

107The 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 become a popular means of paying for stormwater 

system maintenance and improvements. 

108DEP is proposing to develop Rule 62-347. A draft is not yet available. See 

http://www.dep.state.fl.us/water/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). 

109Fla. 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 loading of pollutants. For stormwater discharges to Outstanding 

Florida Waters (OFW), the goal is a 95% reduction. Each of the water management districts has 

rules regulating the construction and operation of stormwater systems110. 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 recent review of the criteria and site conditions, 

however, determined that current permitted designs only remove 40-50% of the nutrients111. 

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 surface water quality. The retention of stormwater that recharges 

groundwater may contribute to achieving 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 special criteria for groundwater 

protection. In a Sensitive Karst Area designated for parts of Alachua and Marion Counties 

special criteria apply to the construction and maintenance of stormwater basins to minimize the 

potential for the development of sinkholes or solution pipes to convey stormwater directly to the 

underlying aquifer112. These criteria primarily help to maintain whatever treatment is provided 

by the system as designed and constructed. Deficiencies in treatment capability are somewhat 

addressed by requirements for additional soil in stormwater basins. 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 groundwater113. 

ERP permits apply to both construction and operation 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 inspection and maintenance are relatively weak. Most 

importantly, there are no means for requiring older systems to be upgraded to incorporate 

additional pollution control measures114. 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 (BMAP) of the TMDL program may provide a 

means for requiring increased pollutant reductions. 

 

110 St Johns River Water Management District, Id., 40C-42 (stormwater), 40C-41 (basin criteria), 

40C-44 (agricultural surface water); South Florida Water Management District, Basis of 

Review for Environmental Resource Permit Applications (July 22, 2007); Northwest 

Florida Water Management District, Id., 62-346; Southwest Florida Water Management 

District, Id., 40D-4, 40, 400 and Basis of Review; Suwanee River Water Management 

District, Id., 40B-4. 

111Harvey H. Harper and David M. Baker, Evaluation of Current Stormwater Design Criteria 

within the State of Florida (June 2007). Final Report to Florida Department of 

Environmental Protection, DEP Contract No. S0108. 

112Fla. Admin. Code r. 40C-41.063(7) (2007). 

113Id., 40C-41.063(3)(a). 

114The 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 that collect and convey stormwater to navigable waters clearly meet the 

federal definition of a “point source”. Nevertheless, because such systems often collect pollutants 

generated by many other sources and are controlled by government entities, Congress and EPA 

delayed requiring NPDES permits for many years. In 1990 EPA began to phase in a regulatory 

program115. Phase I required NPDES permits for larger municipal separate storm sewer systems 

(MS4), large construction sites and some industrial sites. In 1999, Phase II extended the program 

to smaller municipalities and construction sites116. Authority to implement the program has been 

delegated to the DEP, which has adopted NPDES stormwater rules117. These rules require 

monitoring, the development of a Stormwater Management Program (SWMP) and the reduction 

of pollutants discharged to the Maximum Extent Practicable118. Because MS4 stormwater 

permits cannot be issued for longer than 5 years119, 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 systems, a type of Onsite Sewage Treatment and Disposal System 

(OSTDS), are used to treat the wastewater of approximately 1/3 of Florida’s population and are a 

major source of artificial 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 water121.” Traditional septic 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 water supply, the minimum lot size 

is one half acre. Septic systems must be sized according to the projected load and meet standards 

for the quality of the soils, depth to the water table, and setbacks from wells and surface water 

bodies. Generally, there must be 42 inches of suitable soils below the bottom of the drainfield122

and it must be at least two feet above the wet season water table123. The drainfield must be set 

back at least 75 feet from a permanent surface water and 15 feet from a swale or retention area. 

These standards provide for treatment of wastewater 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 subdivisions124. Variances may also be granted to relieve hardships 

where there is no reasonable alternative for waste disposal and the onsite system will not 

 

11555 FR 47990 (Nov. 16, 1990). 

11664 FR 68722 (Dec 8, 1999). 

117Fla. Admin. Code r.62-624 (2007) (MS4). 

118Id., 62-624.500. 

119Id., 62-624.300(5). 

120Florida Department of Health, Nitrogen Impact of Onsite Sewage Treatment and Disposal 

Systems in the Wekiva Area 1-20, 3 (June 30, 2007). 

121Fla. Stat. §381.0065(1) (2007). Fla. Admin. Code r. 64E-6 provides standards for OSTDS. 

122Id., 64E-6.006(1). 

123Id., 64E-6.006(2) 

124Fla. Stat. §381.0065(4)(g)(2007). 
 322 

adversely affect health or “significantly degrade the groundwater or surface waters125.” 

Alternatives to traditional systems that are designed by engineers to meet specified performance 

standards must be approved126. 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 Department127. 

Whenever a central sewer system becomes available, the owner of an OSTDS can be required to 

connect to it128. 

More restrictive criteria for OSTDS have been adopted by statute for the floodways of the 

Suwannee and Aucilla Rivers129 and the Florida Keys130. Recommendations for improved 

regulation of OSTDS have been made for systems installed in the Wekiva Study Area131. 

Local governments can adopt more stringent regulations on construction, performance, setbacks 

or maintenance of OSTDS through local comprehensive plans and land development regulations. 

For example, in the Green Swamp Area of Critical State Concern, two counties have adopted 

ordinances requiring periodic maintenance. Wakulla 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 affect land use, but do not provide the kind of 

comprehensive planning and control of land use that is required to manage 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 Act132

(hereinafter “Growth Management Act”) requires every local government in the state to adopt, 

periodically revise and implement a local comprehensive plan. The local comprehensive plan 

must be “consistent” with the goals, objectives and policies of the State Comprehensive Plan133

and the strategic regional policy plan of the relevant Regional Planning Council134. The Florida 

 

125Id., §381.0065(4)(h). 

126Id., §381.0065(4)(j). 

127Id., §381.0065(4)(j)4. 

128Id., §381.00655. 

129Id., §381.0065(4)(t)(2007). The regulatory floodway is defined to include the 10 year 

floodplain. 

130Fla. Admin. Code r. 64E-6.017-.0182 (2007). 

131Florida 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). 

133Id., ch 187. 

134Consistency of the local comprehensive plan with the state comprehensive and the strategic 
 323 

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 and implement the goals, objectives and policies of 

the adopted comprehensive plan135. It also requires that land development be consistent with the 

adopted plan136. 

A local government’s comprehensive plan must incorporate 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, Drainage, 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 lands137. The Sanitary Sewer, Solid Waste, Drainage, 

Potable Water, and Natural Groundwater Recharge Element requires identification of natural 

drainage features/groundwater recharge areas, assessment of current land use regulations related 

to these issues, and objectives and policies for implementation of land use regulation to protect 

drainage and recharge functions138. The Conservation Element must identify natural resources, 

including groundwater, and incorporate objectives and policies to conserve such resources139. 

Finally, the Intergovernmental Coordination Element requires analysis of current 

intergovernmental coordination, specific objectives for future coordination, and policies 

addressing each objective140. This element could serve a crucial role in encouraging the 

intergovernmental coordination necessary for effective protection of springsheds that lie in more 

than one local government. Coordination also must exist to adequately account for springs 

protection during water supply planning and under water management district regulatory 

programs. Sources of water supply adequate to provide for planned growth must be identified. 

Furthermore, when the Department of Community 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 their comprehensive plans every seven years141. 

This evaluation includes consideration of water management district regional water supply plans 

 

regional policy plan exists if the local plan is compatible with and furthers 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 policies of the plan. Id., § 163.3177(10)(b). 

Consistency with regional water supply plans is also effectively required. Id., §3177(6). 

135Id.,§163.3194, .3202. 

136Id.,§163.3194(1). 

137Id., §163.3177(6)(a) (2006). Fla. Admin. Code r. 9J-5.006 (2007). 

138 Fla. Stat. § 163.3177(6)© (2007. Fla. Admin. Code r. 9J-5.011(2007). 

139Fla. Stat. § 163.3177(6)(d) (2007). Fla. Admin. Code r. 9J-5.013 (2007). 

140Fla. Stat. §163.3177(6)(h) (2007. Fla. Admin. Code r. 9J-5.015 (2007). 

141Fla. 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 for input to help protect springs. Only the DCA, 

however, has the authority to determine that an amendment is not in compliance with the Act143. 

The DCA and the DEP have published a handbook for local governments to use in developing 

policies for springs protection144. The DCA is also developing a model land development code 

for springs protection145. 

The state has assumed a stronger role in supervising local land use decisions in certain areas of 

the state. The Governor and Cabinet, sitting as the Administration Commission, is authorized to 

designate “areas of critical state concern” in order to promote protection of and reverse 

deterioration of water resources in those areas146. The statute authorizing areas of critical state 

concern specifically refers to environmental resources and aquifer recharge areas147. Local 

comprehensive plans and land development regulations must be consistent with Principles for 

Guiding Development adopted by the Administration Commission148. The protection of water 

quality has clearly been an important goal for designating ACSC such as the Florida Keys, Big 

Cypress, and Green Swamp149. While designation as an area of critical state concern 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 and Protection Act, the Legislature required each 

local government within the Wekiva Study Area to amend appropriate elements of its 

comprehensive plan to implement a master stormwater 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 areas, karst features, and sensitive natural habitats 

including Longleaf Pine, Sand Hill, Sand Pine, and Xeric Oak Scrub150.” Legislation proposed 

to protect Silver and Rainbow Springs would require local governments to “adopt local land 

 

142Id., §163.3191(1)© (2007). 

143Id., §163.3184 (2007). DCA’s determination is, however, subject to challenge before an 

administrative law judge and an ultimate determination, based on that record, by the 

Administration Commission. 

144Fla. 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/publications/index.cfm (visited 

Jan 8, 2008). 

145http://www.dca.state.fl.us/fdcp/DCP/springs/index.cfm (visited Jan. 8, 2008). 

146Fla. Stat. §380.05 (2007).. 

147Id., §380.05(2)(a) (2007). 

148Id., § 380.05(1)(b)2 

149http://www.dca.state.fl.us/fdcp/dcp/acsc/index.cfm (visited Feb 13, 2007). 

150Fla. Stat. §369.321(3) (2007). 
 325 

development regulations and other local requirements to minimize harmful impacts to the 

springs151.” 

State statutes can also limit the authority of local governments to plan and regulate for springs 

protection. For example, agricultural activities do not fall within the scope of “development” for 

the purposes of comprehensive plans and thus are not regulated by comprehensive plans152. 

Local governments may not enforce new regulations for agricultural activities already regulated 

by or subject to best management practices promulgated by DEP, the Department of Agriculture 

and Consumer Services (DACS), a water management district, or a federal entity153. Thus, while 

agriculture is a significant contributor of nutrients to springs and springsheds in many areas of 

the state, counties and municipalities have limited authority to improve their regulation of 

agricultural practices. 

Local governments have also been prohibited from restricting the density of residential and 

commercial development in so-called “agricultural enclaves154” and cannot impose an 

“inordinate burden” on any landowner without payment of compensation155. 

One area of current controversy concerns the authority of local governments to regulate the use 

of fertilizers within their jurisdictions. Many local governments have adopted or proposed 

regulations to limit nonagricultural fertilizer use156. Apparent concern for regulatory overreaching and confusion led the 2007 Legislature to create a Consumer Fertilizer Task Force157. 

Among the charges to the was the following: 

Recommending methods to ensure local ordinances 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 standards158. 

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 

 

151Proposed legislation 10-16-07, adding section 369.406(1), F.S. 

152Fla. Stat. § 380.04 (2007), referred to by Id., §163.3164(6). 

153Id., §§ 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). 

155Id., §§70.001-.80. 

156See the undated compilation of “local fertilizer use ordinances” at 

http://consensus.fsu.edu/Fertilizer-Task-Force/legislation.html (visited April 22, 2008). 

157S. 10, Ch. 2007-232, Laws of Florida, codified at Fla. Stat. §576.092 (2007). 

158Id., §576.092(2)(a)4. 

159Florida Department of Agriculture and Consumer Services, Florida Consumer Fertilizer Task 

Force, Final Report to the 2008 Florida Legislature, 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 environment160. The 2008 Legislature is considering bills to 

implement the Task Force recommendations161. 

 

Available at http://consensus.fsu.edu/Fertilizer-Task-Force/index.html (visited April 22, 

2008). 

160 Id. 

161S2352 and H1267, 2008 Regular Session, Florida Legislature. 
 327 

8 

Synthesis of Nutrients and Springs: Knowns, Unknowns, Research Priorities 

and Management Implications 

Mark T. Brown 

Center for Environmental Policy and 

Department of Environmental Engineering Sciences 

University of Florida 

Summary 

This summary of the state of the knowledge related to nutrient effects on spring ecosystems and 

biota draws from the proceeding seven chapters of this report. It represents a synthesis of the 

most important findings; greater details can be found within each of the chapters. This synthesis 

chapter is organized into three parts: 

1) a summary of the “knowns” regarding nutrient effects on springs and spring biota, 

2) the unknowns and suggestions for future research, and 

3) suggestions for a research and adaptive management program for spring ecosystems in 

Florida. 

Florida springs are undergoing change. It is clear that nutrient loads within springsheds, 

particularly nitrate-N, from anthropogenic sources are increasing. These loads have apparently 

increased the quantities of nitrogen in shallow and deep ground waters. Nitrogen, but not 

phosphorus, concentrations in the discharges of many springs have increased in the past several 

decades. There are numerous other biotic and abiotic factors that have also changed in some, but 

not all springs, such as increased recreational use, decreased discharges, decreased O2

concentrations in discharges, increased aquatic weed control, invasive species, and increased 

salinities. By themselves or in combination, these factors may explain 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 nitrate-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 regions 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 prior to delivery to the 

soil and ultimately to the aquifer. 

Few studies have been able to provide a threshold value for nitrate-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 the 
 328 

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 management and research program that faces these 

uncertainties directly and uses them to learn and change is proposed. This proposal combines 

management interventions with targeted monitoring 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 Springshed Action Programs (ASAPs) as a way of developing a 

statewide response to the declines in ecosystem health observed of many springs. The proposed 

ASAPs will develop an environment within which meaningful research, education, social 

learning, and management can occur for the betterment 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 Environmental Protection (DEP) for a summary and 

interpretation of the available information on the effects of nutrients on spring organisms and 

systems. DEP’s mandate was to prepare a review 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 could 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 findings 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 (UF) 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 and 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 Fauna – 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 Ecosystems – Wetland Solutions Incorporated 

7. Managing Nutrient Inputs to Florida Springs: 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, summarizes 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 concentrations are increasing in the water discharging 

from springs (Chapter 2). Indeed, increasing nitrate concentrations in freshwater is not just a 

Florida or karst problem, but rather a global problem. Human populations have altered the global 

nitrogen cycle and other biogeochemical cycles through land use changes, fertilizer use, fossil 

fuel combustion and other pathways. With population and land use change comes nutrient 
 330 

enrichment. Florida’s karst region has experienced unprecedented population growth and 

changes in land use over the past several decades, with a consequent transfer of nutrients to the 

relatively unprotected groundwater. 

Among the nutrients of concern, nitrogen, particularly nitrate, appears 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 conservative solute, with denitrification apparently being 

negligible (Chapter 2). Third, although Florida’s geology is naturally rich in phosphorus, there 

does not appear to be a trend of increasing phosphorus 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 transformation of N appear to be 

the aspect of springs that has received the most study. Some general principles and caveats have 

emerged from this work (Chapter 2): 

• Emphasis on nitrate-N loading has been prioritized because of strong evidence for 

anthropogenic enrichment on groundwater N in the karst environment of North 

Florida while contemporaneous evidence for 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 increases 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 rapidly 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 spring flow derives may be extremely 

large, and hence nominal residence times can be long and extremely variable 

between springs. 

• Some aging measurements of water emerging from springs suggest, on average, it 

has spent between 10 and 30 years in the subsurface, 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 been 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, attenuation withing springsheds appears to be significant 
 331 

(~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. Atmospheric 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 emerging from springs appears to provide 

evidence of mixed sources (i.e., mineral fertilizers, organic wastes), but use of 

isotopes for management purposes requires flowpath tracking and consideration 

of transformations along the flow path. 

Biogeochemical transformations of nutrients in spring systems (Chapter 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 prominent in biofilms, algal mats, and the hyporheic 

sediments resulting from a variety of processes including: organic matter decomposition, 

nitrogen fixation, nitrification /denitrification, iron reduction, 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 Florida 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 nutrient storage, making reversal of current 

conditions where algae dominate spring primary producers problematic. 

In summary, control of nitrate-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 regions 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 prior to delivery to the soil and ultimately to the aquifer. Given 

that springsheds are heterogeneous surfaces of topography, soils, and vegetation, with subsurface 

confining layers that are not ubiquitous, detailed land use planning should figure prominently in 

controlling loads. Best Management Practices (BMPs) in confined regions are of value 

(although, there is some disagreement on the effectiveness 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 nitrate enrichment. 

Effects of Nutrients on Spring Organisms and Systems: What we know

There is no question that Florida’s spring ecosystems are undergoing change. An apparent 

increase in the dominance of filamentous algae in many springs, losses of desirable submerged 

aquatic vegetation (SAV, e.g., native vascular species such as Vallisneria americana or 
 332 

Sagittaria kurziana), increases in invasive vegetation forms, and anecdotal evidence of the 

disappearance of grazers and higher order animals all point to significant structural and 

functional changes in spring ecosystems. 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 the 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 relationship between these phenomena (Chapters 4 and 6). Further, the 

belief that increased algal biomass has led to the displacement of native macrophytes appears to 

be based largely on extrapolation from laboratory studies or qualitative observations (Chapter 4). 

Overall, this review of the literature suggests, that there is little direct 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 parallel a eutrophication progression scheme proposed for shallow and 

relatively open coastal systems (see Duarte 1995). This scheme suggests that increased delivery 

of dissolved nutrients to aquatic systems will favor production of fast-growing algae that extract 

their nutrients primarily from the water column. These algae may ultimately out-compete and 

displace slower growing vascular plants that rely primarily on nutrients in sediment and require 

more light. In the scheme, algal growth can create feedback loops that exacerbate the shift to 

algal dominance and decouple the system from nutrient concentrations in the water column. For 

example, increased organic matter or eutrophication can lead to decreased oxygen levels that can 

stress vascular plants and displace or kill grazers 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 productivity than water-column concentrations, 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 effects such as loss of higher trophic level organisms have had a 

cascading effect on vegetative community structure. 

Effects on Flora and Fauna 

The general pattern that emerges from 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 available results indicate that, if 

limitation exists, both nitrogen and 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 empirical investigations that suggest the potential for a negative 

effect of epiphytes or macroalgae on vascular plants in Florida’s spring systems, although 

none clearly demonstrated it or clarified causal mechanisms. 
 333 

• Algal mat assemblages may be limited by the availability of dissolved inorganic carbon, 

in Florida’s springs. 

• Algal mat thickness in Florida’s springs was weakly and positively correlated with 

nitrogen and phosphorus concentrations, but cover and biomass were not significantly 

correlated with nutrient concentrations. 

• Empirical data suggest that there is no statistically significant link between surface water 

nutrients and production of macrophytes at this time; however, light availability, 

especially as mediated by riparian shading, has been implicated as a limiting factor for 

SAV growing in Florida’s spring-fed streams. 

• The literature suggests that nitrate 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 possibility 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 limiting 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 coastal, spring-fed systems, perhaps masking the 

effects of nutrient enrichment. 

• There is some evidence from a survey of springs in Florida that a negative relationship 

exists between dissolved oxygen concentrations and benthic algal coverage. 

Nutrients can affect fauna in springs through direct 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 habitats. 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 concentrations reported to cause effects on 

animals (Munch et al. 2006). However, nitrate concentrations exceed 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. 

• Invertebrate assemblages vary among different vegetated and unvegetated habitats, which 

suggests that changes in the types and amounts 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 importance of factors beyond nutrients. 

• Fish assemblages and standing stocks vary among different habitats, which suggests that 

changes in habitats may lead to changes in fish assemblages. 

• Evidence of changes in fish assemblages based on repeated surveys in springs that have 

experienced increases in nitrate concentrations (e.g., Silver and Wekiva) is equivocal, 

with significant potential for confounding of drivers. 

• Fish biomass and phosphorus concentrations were related in a nonlinear manner, with 

increases in fish biomass being correlated with 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 arise through interrelated 

and interacting mechanisms. The outcomes bear directly 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 Florida 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 growth 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 downstream export rather than nutrient limitation. 

Overall the interrelationships between invasive plants and spring ecosystems can be summarized 

as follows: 

• Available scientific studies have not definitively 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 function of increased sediment N (albeit 

from tank experiments rather than in situ experiments). 

• Field surveys in Florida lakes indicate a marginally significant relationship between 

dissolved total N and the extent of H. verticillata occurrence. 

• In one long-term study of ecological communities associated with E. crassipes in the St. 

Marks River, faunal habitat values 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 depress DO, coverage by E. crassipes in the flowing waters of the St. 

Marks River did not result in significant oxygen profile differences as compared to 

Sagittaria kurziana (strap-leaf sag) or other native plant communities. 

• The literature suggests that aquatic plant control techniques (e.g. herbicide applications or 

mechanical harvesters) that are used to suppress 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 described as driven by “bottom-up” or 

“top-down” control processes (or combinations of both). Bottom-up control describes processes 

that originate at the bottom of the trophic ladder and are driven generally by exogenous forcing 

functions such as, light, flow regime, temperature, or nutrient availability that act to control 

growth and productivity of primary producers. Top-down control, on the other hand, originates 
 335 

at the higher end of the trophic ladder and generally takes the form of endogenous effects like 

changes in grazing pressure, seed dispersal, and other controlling feedbacks from within. 

However, the exogenous verses endogenous effects should not be taken to mean that top-down 

only implies factors originating within the ecosystem. In fact exogenous drivers like the removal 

of top carnivores or the loss of grazers can ultimately result in top-down cascades of impacts. 

Detailed study of two spring ecosystems, Silver Springs and the Rock Creek, Wekiva Spring 

complex, suggests that there is evidence for ecosystem scale effects of nitrate-N enrichment. 

Using a subsidy stress theory to explain the decline in gross primary production (GPP), it is 

suggested that increases in nitrate-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 today. This declining ecosystem productivity 

documented at Silver Springs was highly correlated with increasing nitrate nitrogen 

concentrations during the 50-year period of available data. Declining spring flows, increased 

shading by riparian trees, and altered fish populations were also observed to be correlated with 

declining ecosystem production at Silver Springs and could offer alternate or cumulative 

explanations of the observed ecosystem changes. The second study conducted in the Wekiva 

River and Rock Springs Run also found an inverse correlation between nutrient (total nitrogen 

and total phosphorus) concentrations and ecosystem metabolism. These spring run ecosystems 

also had other significant environmental stresses caused by humans, including decreases in 

discharge, intensive exotic plant management efforts, and disturbance due to recreational 

activities. Studies of whole 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 ecosystems exists primarily as a collection of 

regulations for the discharge of wastewaters through the National Pollutant Discharge 

Elimination System (NPDES). Additionally, certain sources of nutrient pollution that do not fall 

within the direct jurisdiction of the federal 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 aquatic flora or fauna.” 

Numeric standards can also be developed, as Total Maximum Daily Loads (TMDL), however, 

development of TMDLs for a particular waterbody requires that it be “impaired”. Local 

comprehensive plans and land development regulations can be developed to limit the intensity of 

land development to minimize potential nutrients contributions to springs. They can also provide 

for improved stormwater and wastewater management. 

Overall, the current framework for managing spring 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 program initiatives of a variety of agencies and 

programs including: Florida Department of Environmental Protection, Florida Department of 

Health, DEPs Bureau of Invasive Plant Management, Florida Water Management Districts, 
 336 

Florida Fish and Wildlife Conservation Commission, and local citizens representing important 

stakeholders. 

PART 2: SPRINGS RESEARCH 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 requires 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 everything, 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 significant change. They are 

under assault from a variety of stressors including reduced flows, increased nutrients (principally 

nitrate-N), human disturbance, changes in specific conductance, 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) making generalizations using key principles and 

across all springs to be dangerous at best. 

Our review of the available 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. Yet manage we must in the face of 

significant uncertainty. It is with this uncertainty in mind that we make the following 

recommendations regarding research needs. There are many areas of research 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 concerted belief that the research 

approach should be a systematic one. An approach that incorporates many of the individual 

suggestions into a holistic comparative, cross scale, 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 improve 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 strategies. It is a critical factor necessary in the 

development of TMDLs for spring sheds. As part of a systematic research program, 

continued development of spatially explicit models of nutrient and water budgets is 
 337 

strongly encouraged to better understand the reasons for 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 individual springs. The link 

between discharge and age is an important research need because of its implications for 

regulatory minimum flows and levels and for better understanding 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 the primary objective to determine the degree 

of sorption saturation along principal groundwater flow-paths. 

• Indications are that there may be significant spatial and temporal 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). Research is needed to determine if declines 

in DO or variability in DO levels are related to the increase in the dominance of 

filamentous algae in many springs, either through negative impacts on grazers, or impacts 

on SAV respiratory functions. 

Within-system nutrient dynamics are not well understood and there is a general lack of data on 

characterization of spring systems for biogeochemical purposes. We propose the following 

studies that will elucidate biogeochemical pathways and processes, enhancing spring 

management in the long run 

• Our understanding of nutrient dynamics in Florida spring systems resembles a black box 

approach, where changes in water column 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 distribution play a key role in structuring the biogeochemistry of 

the spring system; therefore, better understanding is needed 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 relative bioavailability of various nutrient forms 

stored in sediments and water column is important to document as are the fluxes between 

sediment, vegetation, and water column. 

• Research is needed to understand the fate of nitrogen within the spring system. This 

would include understanding of biotic uptake processes (algal/macrophyte uptake), 

importance of different N sources (NH4

+

, NO3

-

, DON), and the occurrence, rates and 

importance of biological N2-fixation. Whole spring run N budgets 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 fauna always will be uncertain for two primary 

reasons: 1) we cannot afford the amount of science that would come close to eliminating 

uncertainty and 2) unforeseen and unforeseeable responses will arise as emergent properties of 

these complex systems. In support of sustainable use and management of springs systems, we 

recommend a targeted research program comprising inventories, baselines, 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 “answerable” questions, e.g., questions with welldefined scopes and appropriate levels of precision, and detailed research projects designed to 

provide useful answers. Issues that arose during this review serve as starting points for such 

discussions. Key unknowns and examples of research directions are: 

• A lack of understanding regarding the roles of various types of flora in nutrient budgets 

hampers management decisions regarding assimilation 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, macroalgae, epiphytes, 

microphytobenthos or microalgae found in sediments, and other types of flora. 

• A lack of understanding regarding the links among nutrients and other drivers of primary 

productivity hampers efforts to predict impairment, such as overgrowth or shading of 

vascular plants by algae, set criteria to reverse impairment, or make decisions about 

restoration; and it can be addressed by diagnostic studies that 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 regarding 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 diagnostic 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 flora, grazers, and trophic links to 

be protected or restored by management actions. 

• A lack of understanding regarding the population and assemblage 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 elucidate the toxic effects 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. 

Invasive Plants and Springs: What we don’t know

It is surprising that more research has not been 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 strong incentive for a more holistic approach to 

understanding their short and long term effects. The study of spring ecosystems offer ample 
 339 

opportunities to investigate the role of invasive plants and the possible detrimental effects of 

their management in the changes that are occurring within spring ecosystems. We offer the 

following research topics related to invasive plants and their management in spring systems. 

• Diagnostic studies (nutrient and organic matter 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 characterize the effects on sediment nutrient 

accumulation and ecosystem succession.

• Since the available literature suggests that even historically characteristic nutrient 

concentrations at Florida springs could support invasive plant 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 and studies and observations that indicate 

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 

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 selection, habitat loss, 

faunal toxicology, and dissolved oxygen suppression that may be associated with aquatic 

plant control, a significant diagnostic program of study is warranted. This should include 

detailed in situ studies of the above under differing aquatic plant control regimes 

(including both mechanical and chemical treatment). The funds for such studies should 

be included in aquatic plant control budgets as a matter of course. 

Ecosystem Scale Effects: What we don’t know

Spring ecosystems are exceedingly complex, but potentially 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 nitrate concentrations) and a 

response (such as algal or plant growth and abundance). We know that springs, like other 

ecosystems, have feedback and control pathways that make predicting cause and effect 

relationships somewhat problematic. Yet that is exactly what management interventions require. 

Given X intervention, we expect Y result. Unfortunately the result is often Z; something we did 

not intend at all. 

Ecological monitoring of the multi-level effects of stressors on ecosystems is important for 

understanding the full implications of management interventions. It is not sufficient to monitor 

only the target of a particular intervention, because frequently 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 essential if management is to learn 

from past and refine future interventions. With 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 metabolism, trophic structure, 
 340 

and key forcing functions. Data need to be collected at a spatial and temporal resolution 

to facilitate separation of transient behaviors from long-term trends. 

• Elucidating changes in primary production, ecological efficiency, species composition, 

and succession, that result from changes in forcing functions such as nutrient 

concentrations, dissolved oxygen, salinity, temperature, and shading are necessary if we 

are to understand ecosystem management interventions. This can be accomplished 

through in-situ whole system studies in a variety of springs. 

• Elucidating effects of changes in herbivory, which could lead to primary producer 

effects that appear similar in symptom to nutrient 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, discharge, carbon dioxide 

concentration, nutrient concentrations, oxygen concentration, specific conductance or salt 

contents, and grazing pressure) and the responses of individual species and trophic levels. 

Any changes in individual organisms could then be related to changes in ecosystem 

structure and functioning and coupled with whole ecosystem studies (above) might help 

to elucidate feedbacks and heuristic potentials. 

• Indicators and assessment tools for spring ecosystems are essential. Understanding 

change requires indices that are sensitive 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 “we scientists” were pontificating on the cause 

while the patient was sick and dying on the operating table...a particularly 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 potential diseases, and relatively low-risk of clinical trial error as 

a result of the ecological changes now being experienced 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 springs be selected that represent a variety of 

stressors acting separately and in combination as a means of testing various hypotheses regarding 

the impacts of stressors on spring ecosystems and biota. Tests might involve, for instance, 

choosing springs with low versus high DO, presence 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. 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 states and hysteresis effects in which, even after 

long periods of time, the state of the spring system may be determined by its history. In other 

words removing a stressor may not be sufficient to “restore” a spring ecosystem to a former state 
 341 

as there may be considerable lag in spring ecosystems 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 elucidate spatial variability and temporal dynamics of stressors and 

impacts. Most of the current spring monitoring data has a resolution that makes discriminating 

spatial variability impossible and lacks the temporal sensitivity to account for potential pulse 

events that may have significant impact (low DO/nitrogen/high salinity events, for example). 

Scanning though the list of research suggestions, it is apparent that answers to the many 

uncertainties surrounding Florida Spring management will not come quickly. A sustained 

research program will necessarily require a sustained source of funding. Here we suggest a 

simple source of funding that is tied directly to the resource and that is independent of water 

management district funding or state tax dollars....a source of funding that with appropriate 

social marketing can be thought of as a positive 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 contribution” 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 single-serve bottles per person 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 toward springs research, the sale of bottled water in Florida 

would generate $19.4 million/yr (this does not include tourism sales of bottled water which could 

easily push the total available funding to nearly $21 million per year). If marketed correctly this 

could result in a sustainable source 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 the 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 tax, rather we believe that if presented correctly to the bottling industry 

they would be extremely willing to add a sentence on their bottling labels indicating that with the 

sale of each bottle of water they are contributing 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 are important, unique resources and as with most 

natural resources in Florida, they are under considerable pressure from changes in environmental 

drivers, direct human use, and indirect alteration of the chemical and physical environment in 

which they exist. Equally apparent is the fact that many Florida springs are exhibiting change in 

biological components. While exact causes are uncertain, in the previous sections, we document 

numerous stressors that might be grouped into at least four general classes: physical disruption, 

changes in quantity of discharges, changes in chemical composition of discharges, and changes 

in populations of higher trophic organisms which have top-down cascade effects on lower 
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trophic levels. It is quite possible but highly unlikely that the observed changes in biological 

indicators in some springs may be caused by a single stressor. More probable is the possibility 

that springs exhibiting change are 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 this time that our ability to 

understand causality between stressors and indicators in spring ecosystems is sufficient to allow 

such a fine grain approach to management. Thus, 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 (Grumbine, 1994). The focus on integration of these 

ecological objectives with the simultaneous attainment of social and economic objectives is what 

distinguished ecosystem management from earlier natural resource management approaches. 

Recognizing the often pathological insistence of managers to stick 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 only a whole systems approach to management but an adaptive one 

as well (Holling, 1978; Walters, 1986). 

The Springs Initiative: Protecting Florida’s Springs 

In 2001, the Department of Environmental Protection (DEP) made springs protection a priority 

through the establishment of the Florida Springs Initiative (FSI). Since then, DEP committed 

several million dollars to fund scientific research, water quality and biological monitoring, 

education and outreach, landowner assistance projects, and springs restoration (FSI, 2007). In the 

six years since its establishment, the FSI has accomplished much in the four areas of its 

responsibility: 

• Research and monitoring; 

• Landowner assistance in the form of planning, management, and restoration; 

• Restoration/protection activities and educational outreach 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 called 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 Alliance (RA, 2008) and others 

(Holling, 1978, 1986; Gunderson et al. 1995; Habron, 2003: Folke et al. 2005). A casual 

understanding of Adaptive Management leads to the belief that monitoring management 

activities and occasionally changing them constitutes 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 criticism of the current approach of the FSI, only 
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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 system, comprised of a multitude 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 Approach to Springs Protection 

An adaptive approach to natural resource management involves exploring alternative ways to 

meet management objectives, predicting the outcomes of alternatives based on the current state 

of knowledge, implementing one or more of these alternatives, monitoring to learn about the 

impacts of management actions, and then using 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 about the systems but more importantly, they are 

learning whether the intervention is likely to result 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 information and adjust management if the experiment suggests that 

adjustment is necessary. A main organizational principle of AM is that it requires partnerships of 

policymakers, managers, scientists, and other stakeholders who learn together how to create and 

maintain sustainable resource systems. In other words, 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 complex 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. The complexity of springs issues do not lend 

themselves to traditional single focus resource management that uses informed- trial-and-error of 

applying management interventions, watching the impacts, seeing what works, and then 

adjusting future interventions according to observations. Nor does their resolution lie in either 

private ownership or top-down government control. Instead, the complexity of issues and causes 

must be addressed in an informed process of Adaptive 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. The 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 framework 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, 
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knowledge, science, and power (Stringer et al. 2006). This requires that information flows be 

horizontal, among groups, as well as vertical, among institutional levels.

Managing in the Face of Uncertainty 

The National Research Council (2004) 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 understood. The careful monitoring of 

outcomes both advances scientific understanding and helps adjust policies or operations as part 

of the iterative learning process. AM is not an end in itself or a trial and error process, but rather 

a means to more effective decisions and enhanced benefits through learning while doing. 

Springs are complex, coupled natural and social systems, sometimes called social ecological 

systems. Their complexity leads to uncertainty about the exact structure of the system, the 

relationships between components, and causes and 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 interventions based on the hypotheses and designed 

as scientific experiments. If posed correctly 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 springs ecosystems, the causes of which are somewhat uncertain, we 

must learn as we manage. The opportunity to impose an adaptive management framework on the 

process so that research generates new knowledge and information that flows directly from 

management decisions is irresistible and too 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 and preserving the management system. There is a fundamental 

difference in approach between management that is adaptive and Adaptive Management; the 

former stresses preservation of the status quo, avoidance of risk, and concern with the costs of 

management. The latter accepts change and the uncertainty 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 the creation of up to six Adaptive 

Springshed Action Programs (ASAPs). Called Action Programs instead of management 

programs, we wish to stress that these programs are more than management. They are new 

management, research, and education structures that incorporate institutional verticality (i.e., 

state, regional, and local institutions) along with a horizontal group structure of socio-economic 

stakeholders, managers, and scientists organized as a working group to adaptively manage the 
 345 

spring and its springshed. They are akin to BMAPs 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 changing, a belief that 

through appropriate management it is possible to affect positive change, a willingness to 

implement interventions as scientifically valid experiments, the ability to monitor, collect and 

analyze data, and finally the readiness to change interventions should the data suggest it is 

appropriate to do so. 

The springsheds should be chosen to represent a cross-section of geographical, ecological, 

physiological, sociological, and political settings. As a first cut, we suggest the following eleven 

spring systems, from which 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 currently unaffected by any nutrient increases or 

flow reductions

• Wakulla Springs – an important north Florida spring 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 related to invasive species 

• Homosassa Springs Complex – a system with significant baseline 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 Florida 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 and 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 agricultural landscape 

with the highest nitrate N concentrations in the state 

• Wacissa River Springs – springs in a relatively undeveloped rural landscape with 

minimal nutrient impacts 

By choosing several spring systems simultaneously, we suggest that communication between 

spring ASAPs is just as important to management, learning, and science as within ASAP 

communication. Frameworks for communication between 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 that build a collaborative space for discussions 

between the policy, science, user, and manager groups. These discussions should focus, using a 
 346 

“systems approach”, on elucidating leading problems and issues within each of the spring 

systems, developing hypotheses concerning the causes of problems and effects of interventions, 

and develop scenarios about the future. These 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 process. A deeper understanding of the resource 

issues, management options, and future scenarios are likely outcomes. 

The current bag of spring management tools (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 (whether impacts are from nutrients, 

reduce discharges or human disturbance) will require negotiation. Who pays and how much are 

part of reducing impacts. The proposed ASAPs represent a democratic, responsive bottom-up 

management structure that can garner support and allocate responsibility without heavy-handed 

top-down regulation.

It should be noted that the ASAP approach suggested here is in addition to the needed research 

proposed in the previous section of this chapter. 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 small scale to better understand which interventions 

are most likely to yield desired ecological effects. These are not mutually incompatible, and are 

likely wholly complementary. The research proposed in the previous section is designed to 

provide baseline information about spring ecosystems and responses 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 concerning 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 use the existing spring working 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 statement 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 working 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 facilitating bottom-up discussion about how to 

process and use this information moving forward. All too often, working group 

meetings are characterized by a series of presentations that are not followed by 

serious discussion, synthesis, and argument among stakeholders about the 

information presented, much less the articulation 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 evaluating 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 that critical dialogue, 

collaborative development of research 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 apparent declines in spring ecosystem condition. 

The social framework of the ASAP team should 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 management strategy to restore SAV vegetation in a spring 

run. The appropriate agencies as members of the ASAP team would develop the necessary policy 

and implementation strategy. 

In summary, the key difference between the ASAPs and the current springs working groups is 

related to membership, organization, and empowerment. 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 testable 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 initiative could coordinate the 

ASAPs. We propose the DEP utilize a neutral facilitator whose primary job would be the 

development of consensus between the tripartite members of each ASAP. 

Adaptive Spring and Springshed Research 

Research within the AM context is designed for 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 “pure science” as it must respond to often-conflicting 

group demands of managers, politicians, citizens, developers, environmentalists, and users. It 

must be able to operate under conditions of complexity, uncertainty and risk, and generally be 

applied over large spatial and temporal scales. In addition AM research must cross many 

ecological and organizational boundaries 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 available information with inductive inference to the particular case in 

point. Managers want to move forward with scientific certainty that their intervention will 

achieve the desired result. Science is left wanting; wishing for the time to do the research, 

synthesize the results, and answer the questions. 

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-economic 

constraints. Overall, our understanding of Florida spring systems would benefit from cross 

system comparison. It is important that the science/research component foster the development 

of broad integrative studies that bring policy and management into focus and generate the kinds 
 348 

of knowledge and understanding that will inform and develop better management as well as 

social learning rather than just the accumulation of more information. In the absence of the 

ASAP system and the integrative program of study it would foster, the individual research 

questions given in the preceding section of this chapter remain as important research priorities. 

Most are significant questions about the controlling 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. 
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