United States Office of Water EPA-822-B-01-003
Environmental Protection Agency 4304 October 2001
oEPA Nutrient Criteria
Technical Guidance Manual
Estuarine and Coastal
Marine Waters
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U.S. Environmental Protection Agency
Nutrient Criteria
Technical Guidance Manual
Estuarine and Coastal Marine Waters
October 2001
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Disclaimer
This manual provides technical guidance to States, Indian Tribes, and other authorized jurisdictions to
establish water quality criteria and standards under the Clean Water Act (CWA), to protect aquatic life
from acute and chronic effects of nutrient overenrichment. Under the CWA, States and Indian Tribes are
to establish water quality criteria to protect designated uses. State and Indian Tribal decisionmakers
retain the discretion to adopt approaches on a case-by-case basis that differ from this guidance when
appropriate and scientifically defensible. Although this manual constitutes EPA's scientific
recommendations regarding ambient concentrations of nutrients that protect resource quality and aquatic
life, it does not substitute for the CWA or EPA's regulations; nor is it a regulation itself. Thus, it cannot
impose legally binding requirements on EPA, States, Indian Tribes, or the regulated community, and
might not apply to a particular situation or circumstance. EPA may change this guidance in the future.
Cover Photograph: Somewhere on the Chesapeake Bay. Supplied by David Flemer as a duplicate copy
from the Chesapeake Biological Laboratory Photo Archives, University of Maryland; date unknown but
earlier than 1972.
Nutrient Criteria—Estuarine and Coastal Waters
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CONTENTS
Contributors xi
Acknowledgments xiii
Foreword xv
Executive Summary xvii
Chapter 1. Introduction and Objectives 1-1
1.1 Backround 1-1
1.2 Definition of Estuaries and Coastal Systems 1-2
1.3 Nature of the Nutrient Overenrichment Problem in Estuarine and Coastal Marine Waters 1-5
Scope and Magnitude of the Problem 1-5
1.4 The Nutrient Criteria Development Process 1-8
Preliminary Steps 1-8
Strategy for Reducing Human-Based Eutrophication 1-10
Nutrient Criteria Development Process 1-11
Chapter 2. Scientific Basis for Estuarine and Coastal Waters Quantitative Nutrient Criteria .. 2-1
2.1 Introduction 2-1
Purpose and Overview 2-1
Some Important Nutrient-Related Scientific Issues 2-2
River-to-Ocean Continuum: Watershed/Nearshore Coastal Management Framework 2-5
2.2 Controlling the Right Nutrients 2-10
Overview 2-10
Some Empirical Evidence for N Limitation of Net Primary Production 2-11
Some Threshold Responses to Nitrogen Overenrichment 2-13
Effects of Physical Forcing on Net Primary Production 2-13
Other Physical Factors 2-23
2.3 Nutrient Loads and Concentrations: Interpretation of Effects 2-21
Conceptual Framework 2-24
Examples 2-24
2.4 Physical-Chemical Processes and Dissolved Oxygen Deficiency 2-27
2.5 Nutrient Overenrichment Effects and Important Biological Resources 2-28
Benthic Vascular Plant Responses to Nutrients 2-28
Other Examples of Important Biotic Effects of Nutrient Overenrichment 2-29
2.6 Concluding Statement on Nitrogen and Phosphorus Controls 2-32
Chapter 3. Classification of Estuarine and Coastal Waters 3-1
3.1 Introduction 3-1
Purpose and Background 3-1
Defining the Resource of Concern 3-2
3.2 Major Factors Influencing Estuarine Susceptibility to Nutrient Overenrichment 3-2
Dilution 3-3
Water Residence Time 3-3
Stratification 3-3
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CONTENTS (continued)
3.3 Examples of Coastal Classification 3-4
Geomorphic Classification 3-4
Man-Made Estuaries 3-6
Physical/Hydrodynamic Factor-Based Classifications 3-6
Other Considerations 3-9
Summary 3-11
3.4 Coastal Waters Seaward of Estuaries 3-11
Geomorphic Classification 3-12
Nongeomorphic Classification 3-13
Chapter 4. Variables and Measurement Methods To Assess and Monitor
Estuarine/Marine Eutrophic Conditions 4-1
4.1 Introduction 4-1
4.2 Causal and Response Indicator Variables 4-2
Nutrients as Causal Variables 4-2
Response Variables 4-6
Measures of Water Clarity 4-7
Dissolved Oxygen 4-8
Benthic Macroinfauna 4-9
4.3 Field Sampling and Laboratory Analytical Methods 4-9
Field Sampling Methods 4-9
Laboratory Analytical Methods 4-13
Water Column Nutrients 4-13
Sediment Analyses 4-15
Determination of Primary Productivity 4-16
Phytoplankton Species Composition 4-16
Macrobenthos, Macroalgae, and Seagrasses and SAV 4-17
Chapter 5. Databases, Sampling Design, and Data Analysis 5-1
5.1 Introduction 5-1
5.2 Developing Regional and National Databases for Estuaries and Coastal Waters 5-1
Data Sources 5-3
EPA Water Quality Data 5-3
National Oceanographic and Atmospheric Administration (NOAA) 5-5
Rivers and Streams Water Quality Data 5-6
USGS San Francisco Bay Program 5-6
State/Tribal Monitoring Programs 5-7
Sanitation Districts 5-7
Academic and Literature Sources 5-8
Volunteer Monitoring Programs 5-8
Quality of Historical Data 5-9
Location Data 5-9
Variables and Analytical Methods 5-9
Laboratory Quality Control 5-9
Data Collecting Agencies 5-10
Time Period 5-10
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CONTENTS (continued)
Index Period 5-10
Representativeness 5-10
Gathering New Data 5-10
5.3 Sampling Design 5-10
Sampling Protocol 5-11
Sampling Technique 5-12
Initial Considerations 5-12
Specifying the Population and Sample Unit 5-13
Specifying the Reporting Unit 5-14
Sources of Variability 5-14
Alternative Sampling Designs 5-16
Monitoring Programs 5-18
Citizen Monitoring Programs 5-20
5.4 Quality Assurance/Quality Control 5-21
Representativeness 5-21
Completeness 5-21
Comparability 5-21
Accuracy 5-22
Variability 5-22
5.5 Statistical Analyses 5-22
Data Reduction 5-22
Frequency Distributions 5-23
Correlation and Regression Analyses 5-23
Tests of Significance 5-24
Chapter 6. Determining the Reference Condition 6-1
6.1 Introduction and Definition 6-1
6.2 Significance of Reference Conditions 6-1
6.3 Paucity of Similar Estuarine and Coastal Marine Ecosystems 6-4
6.4 Approaches for Establishing Reference Conditions 6-4
In Situ Observations as the Basis for Estuarine Reference Condition 6-5
Areal Load Approach to Identification of Reference Condition 6-11
Chapter 7. Nutrient and Algal Criteria Development 7-1
7.1 Introduction 7-1
7.2 Role of Regional Technical Assistance Groups 7-2
7.3 Classification 7-3
7.4 Descriptive Background Information 7-3
Estuarine Watershed Characterization 7-3
Within Estuarine System Characterization 7-4
7.5 Elements of Nutrient Criteria 7-5
Reference Condition 7-5
Historical Information 7-5
Models 7-5
Antidegradation Policy and Attention to Downstream Effects 7-6
The RTAG 7-7
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CONTENTS (continued)
7.6 Hypothetical Examples of Nutrient Criteria Development Deliberations 7-7
Scenario 7-7
7.7 Evaluation of Proposed Criteria 7-8
Guidance for Interpreting and Applying Criteria 7-9
Do the Criteria Protect Designated Uses? 7-9
Restoration Goals 7-11
Sampling for Comparison to Criteria 7-11
7.8 Nutrient Criteria Interpretation Procedures 7-12
Decisionmaking Protocol 7-12
Multivariable Enrichment Index 7-12
Frequency and Duration 7-13
7.9 Criteria Modifications 7-14
7.10 EPA, State, or Tribe Responsibility under the Clean Water Act 7-14
7.11 Implementation of Nutrient Criteria into Water Quality Standards 7-14
Chapter 8. Using Nutrient Criteria to Protect Water Quality 8-1
8.1 Managing Point Source Pollution 8-1
The Clean Water Act and Water Quality Standards 8-1
Protecting Designated Uses 8-2
Maintaining Existing Water Quality 8-3
General Policies 8-4
Providing Flexibility in Implementation 8-6
NPDES Permits 8-7
Look to the Future ... Pollutant Trading 8-10
8.2 Managing Nonpoint Source Pollution 8-11
Nonpoint Sources of Nutrients 8-12
Efforts to Control Nonpoint Source Pollution 8-13
National Estuary Program 8-14
Atmospheric Deposition 8-15
Coastal Nonpoint Pollution Control Programs 8-16
Farm Bill Conservation Provisions 8-17
8.3 Comprehensive Procedure for Nutrient Management 8-19
Step 1: Status Identification 8-19
Step 2: Background Investigation 8-20
Step 3: Data Gathering and Diagnostic Monitoring 8-21
Step 4: Source Identification 8-23
Step 5: Management Practices for Nutrient Control 8-24
Step 6: Detailed Management Plan Development 8-26
Step 7: Implementation and Communication 8-26
Step 8: Evaluation Monitoring and Periodic Review 8-26
Step 9: Completion and Evaluation 8-27
Step 10: Continued Monitoring of the System 8-27
8.4 Resources 8-28
Chapter 9. Use of Models in Nutrient Criteria Development 9-1
9.1 Introduction 9-1
Use of Empirical Models in Nutrient Criteria Development 9-2
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CONTENTS (continued)
Use of Mathematical Models in Nutrient Criteria Development 9-3
9.2 Model Identification and Selection 9-3
Model Identification 9-4
Model Selection 9-8
9.3 Model Classification 9-9
Level I Models 9-10
Level II Models 9-13
Level III Models 9-14
Level IV Models 9-15
Summary of Model Capabilities 9-17
9.4 Use of Models for Nutrient Investigation 9-17
Model Calibration and Validation 9-17
9.5 Management Applications 9-22
Load-Response Analysis 9-22
Acceptable Nutrient Loads 9-23
Case Study Example 9-24
References R-l
Appendixes
Appendix A: Conditions for Bloom Development: Interplay among Biogeochemical, Biological,
and Physical Processes A-l
Appendix B: Additional Information on the Role of Temperature and Light on Estuarine
and Coastal Marine Phytoplankton B-l
Appendix C: Additional Information on Flushing in Estuaries C-l
Appendix D: NOAA Scheme for Determining Estuarine Susceptibility D-l
Appendix E: Comparative Systems Empirical Modeling Approach: the Empirical Regression
Method to Determine Nutrient Load-Ecological Response Relationships
for Estuarine and Coastal Waters E-l
Appendix F: Selected Theoretical Approaches to Classification of Estuaries and Coastal Waters . . F-l
Appendix G: Examples of Nutrient Concentration Ranges and Related Hydrographic Data
for Selected Estuaries and Coastal Waters in the Contiguous States
of the United States G-l
Appendix H: Preliminary Statement of Proposed Near Coastal Marine Nutrient Sampling and
Reference Condition Development Procedure H-l
Case Studies
San Francisco Bay Program: Managing Coastal Resources of the U.S CS-1
Long Island Sound - Hypoxia CS-4
NP Budget for Narragansett Bay CS-15
Tampa Bay Case Study CS-19
Restoring Chesapeake Bay Water Quality CS-27
A Perspective from Washington State CS-42
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CONTENTS (continued)
Figures
Figure 1-la. Draft aggregation of Level III ecoregions for the National Nutrient Strategy illustrating
those areas most related to coastal and estuarine nutrient criteria development 1-3
Figure 1-lb. Coastal provinces 1-4
Figure 1-2. The eutrophication process 1-6
Figure 1-3. Expanded nutrient enrichment model Source: Brickeretal. 1999 1-7
Figure 1-4. Elements of nutrient criteria development and their relationships in the process 1-9
Figure 1-5. Derivation of the reference condition and the National Nutrient Criteria Program
using TP, TN, and Chlorophyll a as example variables 1-9
Figure 1-6. Flowchart of the nutrient criteria development process 1-12
Figure 2-1. Idealized scheme defining the coastal ocean and the coastal zone 2-2
Figure 2-2. Schematic representation of contemporary (Phase II) conceptual model of coastal
eutrophication 2-5
Figure 2-3. Salinity zones 2-7
Figure 2-4. Schematic illustrating the central role of phytoplankton as agents of biogeochemical
change in shallow coastal ecosystems 2-7
Figure 2-5. Transport of nutrients to Laholm Bay, Sweden 2-12
Figure 2-6. Summary of nitrogen:phosphorus ratios in 28 sample estuarine ecosystems 2-14
Figure 2-7. Factors that determine whether nitrogen or phosphorus is more limiting in
aquatic ecosystems 2-15
Figure 2-8. Cartoon diagrams of three physical forcings that operate at the interface between SCEs
and the coastal ocean (tides), watershed (river inflow), and atmosphere (wind) 2-19
Figure 2-9. Simple schematic diagram showing the influences of river flow on ecosystem
stocks and processes examined in this study 2-20
Figure 2-10. Scatter diagram showing the relationship between the rate of decline in
dissolved-oxygen concentrations in deep water and average deposition rates
of total chlorophyll a during the spring-bloom period 2-20
Figure 2-11 Schematic diagram of coastal plain estuary types, indicating direction and
a-d. degree of mixing 2-22
Figure 2-12. Net transports in estuaries resulting from estuarine flows and mixing 2-23
Figure 2-13. Net movement of a particle in each layer of a two-layered flow system 2-23
Figure 2-14. The fraction of landside nitrogen input exported from 11 North American and
European estuaries versus freshwater residence time (linear time scale) 2-25
Figure 2-15. Scatter plots of water column averaged chlorophyll a at a mesohaline station versus
several different functions of total nitrogen loading rate measured at the fall line of
the Potomac River estuary 2-26
Figure 2-16a. Primary production by phytoplankton as a function of the estimated annual input of
dissolved inorganic nitrogen per unit volume of a wide range of marine ecosystems. . 2-30
Figure 2-16b. Primary production by phytoplankton as a function of the annual input of dissolved
inorganic nitrogen per unit area of a wide range of marine ecosystems 2-30
Figure 2-16c. Fisheries yield per unit area as a function of primary production in a wide range of
estuarine and marine systems 2-31
Figure 2-17. Comparative evaluation of fishery response to nutrients 2-32
Figure 3-1. Idealized micronutrient-salinity relations showing concentration and mixing of
nutrient-rich river water with nutrient-poor seawater 3-10
Figure 3-2a. Relationship between the mean annual loadings of dissolved inorganic nitrogen
and the mean annual concentration of chlorophyll a in microtidal and macrotidal
estuaries 3-11
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CONTENTS (continued)
Figure 3-2b. Relationship between the mean annual concentrations of dissolved inorganic
nitrogen and chlorophyll a in microtidal and macrotidal estuaries 3-12
Figure 6-1. Environmental quality scale representing reference conditions and potential nutrient
criteria relative to designated uses 6-2
Figure 6-2. Hypothetical frequency distribution of nutrient-related variables showing quantities for
reference or high-quality data and mixed data 6-7
Figure 6-3. Hypothetical example of load/concentration response of estuarine biota to increased
enrichment 6-9
Figure 6-4. An illustration of the comparison of past and present nutrient data to establish
a reference condition for intensively degraded estuaries 6-10
Figure 6-5. Areal load estimate approach to nutrient reference condition determination 6-14
Figure 7-1. Generalized progression and relationship of the elements of a nutrient criterion 7-2
Figure 7-2. Hypothetical illustration of developing a TN criterion in an estuary 7-8
Figure 8-1. Components of water quality standards 8-2
Figure 8-2. "Threefold framework" of evaluation 8-25
Figure 9-1. Eutrophication model framework 9-6
Figure 9-2. Use of models in load-response analysis 9-23
Figure 9-3. Use of models in determining allowable loads 9-24
Figure 9-4. Shipps Creek site map and salinity monitoring locations 9-25
Figure 9-5. Model results for existing conditions 9-27
Figure 9-6. Model results for 50% reduction in WWTP load 9-28
Tables
Table 2-1. Categorization of the world's continental shelves based on location, major river,
and primary productivity 2-8
Table 2-2. Estuaries exhibiting seasonal shifts in nutrient limitation with spring P limitation
and summer N limitation 2-12
Table 2-3. DO, nutrient loading, and other characteristics for selected coastal areas and a MERL
mesocosm enrichment experiment 2-16
Table 3-1. General drowned river valley estuarine characteristics 3-5
Table 3-2. Classification of coastal systems based on relative importance of river flow, tides,
and waves to mixing 3-9
Table 4-1. Suggested methods for analyses and monitoring of eutrophic conditions of
coastal and marine environments 4-10
Table 6-1. Summary of estuarine and coastal nutrient reference condition determinations 6-5
Table 6-2. Requisite assumptions for establishing watershed-based reference conditions 6-12
Table 7-1. Example of an enrichment index using the middle portion of a hypothetical estuary . . 7-13
Table 8-1. States for which the nonpoint source agency is not the water quality agency 8-14
Table 8-2. States and territories with coastal nonpoint pollution control programs 8-17
Table 9-1. Basic model features 9-18
Table 9-2. Key features of selected models 9-19
Table 9-3. Calculation spreadsheet for Shipps Creek estuary 9-26
Table 9-4. Chesapeake watershed nitrogen deposition under varying management schemes for
emissions of nitrogen atmospheric depositions precursors 9-29
Table 9-5. Water quality state variables used in CBEMP 9-31
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x Nutrient Criteria—Estuarine and Coastal Waters
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CONTRIBUTORS
Jessica Barrera (Hispanic Association of Colleges and Universities/University of Miami)
Robert Cantilli (U.S. Environmental Protection Agency)
Ifeyinwa Davis (U.S. Environmental Protection Agency)*
Edward Dettmann (U.S. Environmental Protection Agency)
Jen Fisher (University of Georgia)
David Flemer (U.S. Environmental Protection Agency)*
Thomas Gardner (U.S. Environmental Protection Agency)*
George Gibson (U.S. Environmental Protection Agency)*
Debbi Hart (U.S. Environmental Protection Agency)
James Latimer (U.S. Environmental Protection Agency)
Scott Libby (Battelle)*
Greg Smith (GLEC, Inc.)
CarolAnn Siciliano (U.S. Environmental Protection Agency)*
Jack Word (MEC Analytical Systems)*
* Denotes primary authors
WORK GROUP MEMBERS
Ifeyinwa Davis (U.S. Environmental Protection Agency)
David Flemer (U.S. Environmental Protection Agency)
John Fox (U.S. Environmental Protection Agency)
George Gibson (U.S. Environmental Protection Agency)
Debbi Hart (U.S. Environmental Protection Agency)
Suzanne Bricker (National Oceanic and Atmospheric Administration)
Dorothy Leonard (National Oceanic and Atmospheric Administration)
Scott Libby (Battelle)
Greg Smith (GLEC, Inc.)
Jack Word (MEC Analytical Systems)
Nutrient Criteria—Estuarine and Coastal Waters xi
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xii Nutrient Criteria—Estuarine and Coastal Waters
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ACKNOWLEDGMENTS
The authors wish to thank the following peer reviewers for their assistance in the preparation of this
manual: Don Boesch (University of Maryland) and Hans Paerl (Institute of Marine Sciences, University
of North Carolina), Matthew Liebman (U.S. EPA Region I), Michael Bira (U.S. EPA Region VI), and
Kenneth league (U.S. EPA Region VI). Walter Nelson and Peter Eldridge (U.S. EPA Office of
Research and Development) provided comments on an early draft of the manuscript, and Thomas
Brosnan (National Oceanic and Atmospheric Administration, Damage Assessment Center), Michael
Kemp (Horn Point Laboratory, CES, University of Maryland), Jonathan Pennock (Dauphin Island
Laboratory, University of South Alabama), Hassan Mirsajadi (Delaware Department of Natural
Resources and Environmental Control), and David Tomasko (State of Florida Southwest Florida Water
Management District) provided formal peer review comments on a final working draft. Additional
comments on the working draft were provided by Suesan Saucerman (U.S. EPA Region IX); John (Jack)
Kelly, James Latimer, and Edward Dettmann (U.S. EPA ORD); Lewis Linker (U.S. EPA Chesapeake
Bay Program); Laura Gabanski (U.S. EPA Office of Wetlands, Oceans, and Watersheds); Joel Salter
(Office of Wastewater Management); Mimi Dannel and Marjorie Wellman (U.S. EPA Office Science
and Technology); CarolAnn Siciliano (U.S. EPA Office of General Council); and Cynthia Moncreiff
(University of Southern Mississippi). Treda Smith (U.S. EPA Office of Water) assisted in compiling
references. Edits and suggestions made by the peer review panel were incorporated into the final version
of the manual. State agencies and private interest groups also offered comments and they were addressed
where possible in this manual. We appreciate the work of Joanna Taylor, The COM Group, Inc., who
patiently and graciously made repeated format changes to this manuscript as it evolved over the several
months. The authors of the case studies are thanked for their contributions.
Estuaries and coastal waters are a diverse suite of ecosystems, and differences in methods and
approaches are to be expected. This and subsequent manuals are not intended to be singular, one-time
publications. As experience accumulates, future editions will be prepared to reflect new understanding.
Nutrient Criteria—Estuarine and Coastal Waters xiii
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xiv Nutrient Criteria—Estuarine and Coastal Waters
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FOREWORD
This manual is intended for State/Tribal and Federal agency personnel actively engaged in water resource
management data collection, assessment, planning, and project implementation. Consequently, it
incorporates both a scientific rationale and enough of the "nuts and bolts" of nutrient criteria
development and management to help both initiates and those experienced in water resource
management.
These nutrient criteria development and management efforts are directed at anthropogenic sources.
Inherent "natural" background levels are not and should not be subject to management. Our
responsibility is to abate human-caused eutrophication in estuaries and coastal or "near coastal" (out to
about 20 nautical miles) marine waters.
To distinguish between natural background enrichment and human impacts, it is necessary to identify
localities that experience minimal human influence. Ambient nutrient measurements at these sites may
then be compared to similar sites that do experience human influences. The difference in nutrient
measurements is the difference between a reference site and a test site. A reference condition is a
collection of measurements from several reference sites that incorporates a central tendency statistic.
Because of differences in geologic parent material, climate, and geography, reference conditions are
different from one region to another. Similarly, waterbodies, especially estuaries, often respond
differently to nutrient inputs. Lakes and reservoirs are different from streams and rivers, and estuaries
and coastal marine waters have characteristics different from both. Criteria have to be designed for
particular waterbody types and the regions in which they lie.
The primary variables of concern in criteria development are two causal enrichment variables: total
phosphorus (TP) and total nitrogen (TN). These nutrients are essential to algal and plant production and
are the base of the food chain that supports all other life in the system. Also, two initial response
variables usually are the first indicators of biological growth reaction to enrichment. One is chlorophyll
a, which indicates photosynthesis and biomass production; the other is Secchi depth, a measure of water
clarity or a measure of turbidity, reflecting planktonic growth in the absence of inorganic suspended
material. In many marine and estuarine instances dissolved oxygen concentration (DO) and macrophyte
growth and density are also important measures and, where indicated, may be included as initial response
variables. Other measures can also be used, but these have been selected by EPA as of primary concern.
Nutrient criteria consist of judicious incorporation of present reference condition information about the
primary variables, together with a knowledge of historical conditions and trends in the nutrient quality
of the resource. These two factors, possibly augmented by data extrapolations or models, are analyzed
objectively by a panel of regional specialists well versed in the biology, physics, and chemistry of the
systems of concern. The criteria are also evaluated with respect to the possible consequences of their
implementation on downstream receiving waters. All of these elements are required for the
development of a nutrient criterion.
Nutrient Criteria—Estuarine and Coastal Waters xv
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With this information, the status of a given water resource can be determined, management plans can be
made, and management efforts can be evaluated.
The best possible understanding of the physical, chemical, and biological interrelationships in the
environment is important in nutrient criteria development and the subsequent management response.
However, effective nutrient criteria can and should be developed even in the absence of an in-depth
scientific investigation of the ecological processing of nutrients in the estuarine and marine environment.
An adequate number of proximal reference sites and current knowledge of the system are sufficient to
initiate criteria development and proposed management responses. A conservative, environmentally
responsible start can be made toward alleviating nutrient pollution, subject to adjustment as more
scientific knowledge is obtained and verified.
The reference condition approach to criteria development was peer reviewed by the USEPA Science
Advisory Board in 1990 and 1994 and judged to be scientifically defensible. It is also likely to be the
most cost-effective approach.
xvi Nutrient Criteria—Estuarine and Coastal Waters
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EXECUTIVE SUMMARY
This manual is designed for use by State, Tribal, and Federal water resource managers as they address the
cultural enrichment of their waters in conjunction with the EPA National Nutrient Criteria Program. It is
intended to provide a stepwise sequence of actions leading to the development of nutrient criteria for
estuarine and near-coastal marine waters to be used in correcting this overenrichment problem.
The premise of the National Nutrient Criteria Program is that many, if not most, of our nation's estuarine
and coastal waters are moderately to severely polluted by excessive nutrients (Bricker et al.1999),
especially nitrogen and phosphorus. This nutrient pollution affects not only the biotic integrity of the
waters and the decline of valuable fish and shellfish, it has the potential to cause harm to the public
health through hazardous algal blooms and the propagation of waterborne diseases. To address this
problem, EPA uses a regionalized, waterbody type specific approach to the development of nutrient
criteria or benchmarks for management decisionmaking. These criteria are based on the measurement of
the most natural (or least impacted by human development) waters of a given type in a given area
reflecting the condition to be expected in that region if human impacts are not a factor or are at least
minimized. The variables of specific concern are total phosphorus and total nitrogen as causal variables,
algal biomass (e.g., chlorophyll a for phytoplankton and ash-free dry weight for macroalgae), and water
clarity (e.g., Secchi depth) as early response variables. In waters that already experience hypoxia,
dissolved oxygen should be added as a response variable. EPA encourages States and Tribes to consider
additional response indicators such as seagrasses and algal species composition.
This natural ambient background or "reference condition" is an important element of the nutrient criteria
to be developed. The other elements are: an understanding of the historical status and trend of the water
resource to help put the reference condition in perspective; models of the nutrient data to help better
understand historical and present information and to project future consequences; concern and attention
to the effects of any criteria development on downstream receiving waters; and the objective compilation
and assessment of all of this information by a skilled body of regional experts...the "Regional Technical
Assistance Group" or RTAG. The regional criteria so developed are guidelines the States and Tribes of
the continental United States can use as they prepare their own criteria and standards for the
improvement and protection of the nation's coastal waters.
The first of the actions needed to reach this criteria objective is the organization and utilization in each
EPA Region of an RTAG consisting of specialists from State and Federal natural resource management
agencies versed in the management and scientific principles most appropriate to that region and those
waters. These are water resource managers, oceanographers, chemists, land use specialists, biologists,
estuarine ecologists, statisticians, and similar local civil service experts employed by the State or Federal
government. Academicians, special interest groups, and environmental group representatives are also
important participants in the criteria development process and may assist the RTAG in its efforts.
The first requirement of the RTAG is the review and refinement of ecoregional determinations as
most appropriate to the area. These are the geographic boundaries surrounding the similar estuarine and
Nutrient Criteria—Estuarine and Coastal Waters xvii
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coastal marine waters for which the criteria will apply. They are based on the EPA Ecoregion concept
and incorporate attendant coastal Provinces, both of which are based on geographic and geologic
similarities of landforms and parent material. The importance of this regionalization is the effort to deal
with waters all having a similar inherent background nutrient loading and response characteristic. Once
the regional boundaries and perhaps subregional divisions are completed, the RTAG investigates the
physical classification of the waters into similar estuaries or coastal reaches or embayments for criteria
development. In many instances the estuaries may be unique and require specific criteria.
Within the classification scheme developed, reference sites are identified as those areas suffering the
least cultural development or impact, and the compilation of similar reference sites becomes a reference
condition. The manual describes the scientific rationale for the variables selected, the dynamics of the
receiving waters, and potentially confounding physical and chemical interrelationships influencing
criteria development. It also describes sampling and analytical techniques for data gathering and
processing to develop the reference conditions as well as several options for the compiling of this
information. These include: (1) recognition and measurement of an excellent water body of ideal
nutrient water quality with the aim of preserving this state; (2) in situ reference site determinations for
moderately degraded waters; (3) hind casting for historical information from past higher nutrient quality
conditions to determine the reference condition when no reference sites remain; (4) use of loading
estimations from reference quality subestuarine tributary systems and projection to the estuary; and (5)
options for establishing coastal nutrient reference conditions including a Nutrient Criteria Program pilot
demonstration project.
Once the reference condition(s) has been determined, the RTAG then addresses the historical
perspective; considers the need for models to project future consequences; considers the potential effect
on receiving waters; and employs its own good judgment in collectively determining the appropriate
criteria values for each of the variables to protect the waters of concern and their designated uses. A
procedure is also suggested to equate the multiple criteria variables in a comprehensive dimensionless
index score. The manual concludes with a chapter on model development and applications to the criteria
program, and a chapter describing the application and implementation of nutrient criteria with emphasis
on EPA Standards and Monitoring Divisions and a description of a comprehensive ten step sequential
technique for water resource management.
This comprehensive progression from data collection to reference condition determination to criteria
development and management responses, is intended to help users achieve the restoration and protection
of the nutrient water quality of the nation's estuarine and near-coastal marine water resources.
xviii Nutrient Criteria—Estuarine and Coastal Waters
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CHAPTER 1
Introduction and Objectives
Background
Definition of Estuaries and Coastal Systems
Nature of the Nutrient Overenrichment Problem in
Estuarine and Coastal Marine Waters
Man has had a long and intimate association with the sea. It has borne his commerce and brought food
to his nets; its tides and storms have shaped the coast where his great cities have grown; the broad
estuaries have provided safe harbors for his ships; and the rhythm of its tides has taught him the
mathematics and science with which he now reaches for the stars (U.S. Department of the Interior 1969).
1.1 BACKGROUND
Nutrient overenrichment is a major cause of water pollution in the United States. The link between
eutrophication—the overenrichment of surface waters with plant nutrients—and public health risks has
long been presumed. However, human health concerns such as (1) Escherichia coli and the spread of
disease in sewage-enriched waters; (2) trihalomethanes in chlorine-treated eutrophic reservoirs; (3) the
incidence of nutrient-stimulated hazardous algal blooms in eutrophic estuarine surface waters with
suspected attendant human illnesses, including recent Pfiesteria investigations; and (4) the relationship of
phytoplankton blooms in nutrient-enriched coastal waters of Bangladesh to cholera outbreaks (Scientific
American, December 1998) all suggest that overenrichment pollution is not only an aesthetic, aquatic
community problem, but also a public health problem.
The purpose of this document is to provide scientifically defensible technical guidance to assist States,
authorized Tribes, and other governmental entities in developing numeric nutrient criteria for estuaries
and coastal waters under the authority of the Clean Water Act (CWA), Section 304a. The objective is to
reduce the anthropogenic component of nutrient overenrichment to levels that restore beneficial uses
(i.e., described as designated uses by the CWA), or to prevent nutrient pollution in the first place. The
primary users of this manual are State/Tribal and Federal agency water quality management specialists
and related interest groups. The manual is intended to facilitate an understanding of cause-and-effect
relationships in these complex systems and serve as a guide for nutrient criteria development, a resource
of technical information, a summary of the scientific literature, and a brief technical account of the
ecological structure and function of estuaries and coastal waters to facilitate an understanding of these
complex systems.
To combat the nutrient enrichment problem and other water quality problems, EPA published the Clean
Water Action Plan, a presidential initiative, in February 1998. Building on this initiative, EPA developed
a report entitled National Strategy for the Development of Regional Nutrient Criteria (U.S. EPA 1998a).
Criteria form the scientific basis, or yardstick, for ensuring that a desired result will occur because of a
particular form of environmental stress, in this case nutrient overenrichment. The strategic report
outlines a framework for development of waterbody type-specific technical guidance with emphasis on
the reference condition approach that can be used to assess nutrient status and develop region-specific
Nutrient Criteria—Estuarine and Coastal Waters
1-1
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numeric nutrient criteria. This technical guidance builds on that strategy and provides guidance for
nutrient criteria development for estuaries and coastal waters. Because estuaries and coastal waters lie at
the interface of the land and include various ecoregions and their rivers, this manual departs somewhat
from the freshwater manuals (e.g., Lakes and Reservoirs, EPA-822-BOO-001, and Rivers and Streams,
EPA-822-B-00-002; also available on the EPA web site: www.epa.gov/ost/standards/nutrient.html in
PDF format) and considers both land-based ecoregions and coastal ocean provinces as the geographic
framework. The freshwater nutrient guidance manuals used the ecoregion and subecoregion as the
predominant geographic operational units.
Because of differing geographic and climatic conditions among the East, Gulf, and West Coasts, uniform
national criteria for estuarine and coastal waters are not appropriate; they should be developed at the
State, regional, or individual waterbody levels. Figures l-la,b illustrate the pertinent ecoregions
(including geologic province) of the continental United States associated with coastal and estuarine
waters. In some cases, multiple criteria may be required for large systems with extended physical
gradients. This manual therefore does not provide guidance on how to set nationwide criteria, but
provides State water resource quality managers with guidance on how to set nutrient criteria themselves
relative to EPA regional criteria. This approach is in contrast to toxic chemical criteria, which tend
toward single national numbers with appropriate modifiers (e.g., water hardness for metals). It explores
some approaches to classification of estuaries and coastal shelf systems. The ability to develop useful
classification schemes is still in a highly developmental stage and needs considerable improvement. The
manual describes a minimum set of variables that are recommended for criteria development and
describes methods for developing appropriate values for these criteria. It also provides information on
sampling, monitoring, data processing, modeling, and approaches to implementation and management
responses.
1.2 DEFINITION OF ESTUARIES AND COASTAL SYSTEMS
It is important to have a clear view of the ecosystems that are the focus of this manual. The term
"estuary" has been defined in several ways. For example, a classical definition of estuaries focuses on
selected physical features—e.g., "semi-enclosed coastal waterbodies which have a free connection to the
open sea and within which sea water is measurably diluted with freshwater derived from the land"
(Pritchard 1967) (see Kjerfve 1989 for expanded definition). This definition is limited because it does
not capture the diversity of shallow coastal ecosystems today often lumped under the rubric of estuary.
For example, one might include tidal rivers, embayments, lagoons, coastal river plumes, and river-
dominated coastal indentations that many consider the archetype of estuary. To accommodate the full
range of diversity, the classical definition should be expanded to include the role of tides in mixing,
sporadic freshwater input (e.g., Laguna Madre, TX), coastal mixing near large rivers (e.g., Mississippi
and Columbia Rivers), and tropical and semitropical estuaries where evaporation may influence
circulation. Also, reef-building organisms (e.g., oysters and coral reefs) and wetlands (e.g., coastal
marshes) influence ecological structure and function in important ways, so that biology has a role in the
definition.
1-2 Nutrient Criteria—Estuarine and Coastal Waters
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Draft Aggregations of Level III Ecoregions
for the National Nutrient Strategy
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estuarine criteria development.
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Figure 1-lb. Coastal provinces.
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As will be shown, water depth plays a role in the relative importance of sediment-water column fluxes of
materials, including nutrients. These features paint a picture of high ecosystem diversity, where
prediction of susceptibility to nutrient overenrichment is still a scientific challenge and often requires a
great deal of site-specific information. It is because of this diverse response that reference conditions are
a part of nutrient criteria development.
Coastal waters are defined in this manual as those marine systems that lie between the mean highwater
mark of the coastal baseline and the shelf break, or approximately 20 nautical miles offshore when the
continental shelf is extensive. This area will hereafter be referred to as coastal or near-coastal waters.
Most States have legal jurisdiction out to the 3-nautical-mile limit. However, coastal oceanic processes
beyond this limit may influence nutrient loading and system susceptibility within the 3-mile zone.
1.3 NATURE OF THE NUTRIENT OVERENRICHMENT PROBLEM IN ESTUARINE
AND COASTAL MARINE WATERS
Scope and Magnitude of the Problem
Nutrient overenrichment problems are perhaps the oldest water quality problems created by humankind
(Vollenweider 1992) and have antecedents that extend into biblical history. The basic cause of nutrient
problems in estuaries and nearshore coastal waters is the enrichment of freshwater with nitrogen (N) and
phosphorus (P) on its way to the sea and by direct inputs within tidal systems. Eutrophication, an aspect
of nutrient overenrichment, is portrayed in Figure 1-2. In recent decades, atmospheric deposition of N
has been an important contributing factor in some coastal ecosystems (Vitousek et al. 1997, Paerl and
Whitall 1999).
In U.S. coastal waters, nutrient overenrichment is a common thread that ties together a diverse suite of
coastal problems such as red tides, fish kills, some marine mammal deaths, outbreaks of shellfish
poisonings, loss of seagrass and bottom shellfish habitats, coral reef destruction, and hypoxia and anoxia
now experienced as the Gulf of Mexico's "dead zone" (NRC 2000, Rabalais et al. 1991). Additionally,
recent evidence suggests that nutrient enrichment can exacerbate human health effects (Colwell 1996).
These symptoms of nutrient overenrichment often are preceded by primary symptoms (e.g., an increase in
the rate of organic matter supply, changes in algal dominance, and loss of water clarity) followed by one
or more secondary symptoms listed above (Figure 1-3). Nixon (1995) defined eutrophication as an
increase in the rate of supply of organic matter to a waterbody. In this manual, nutrient overenrichment
is defined as the anthropogenic addition of nutrients, in addition to any natural processes, causing
adverse effects or impairments to beneficial uses of a waterbody. The scientific literature still uses
overenrichment and eutrophication as synonyms. The terms have different meanings, however, because
eutrophication is a natural process in freshwater lakes and presumably in coastal marine waters. An
argument can be made that nutrient stress on coral reefs can cause a loss of symbiotic algae (i.e.,
dinoflagellates), resulting in loss of organic matter and death of the coral colony, a condition not
consistent with eutrophication in the strict sense.
Nutrient Criteria—Estuarine and Coastal Waters 1-5
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The Eulrophication Process
River
Freshwate
Nutrients (t
Sediments,
organic car
Healthy bcnthic community
(worms, snails, bivalves,
crustaceans)
& oxygen consul
Figure 1-2. The eutrophication process. Eutrophication occurs when organic matter increases in an ecosystem.
Eutrophication can lead to hypoxia when decaying organic matter on the seafloor depletes oxygen, and the
replenishment of the oxygen is blocked by stratification. The flux of organic matter to the bottom is fueled by
nutrients carried by riverflow or, possibly, from upwelling that stimulates growth of phytoplankton algae. This flux
consists of dead algal cells together with fecal pellets from grazing zooplankton. Sediment coupled nitrification-
denitrification is shown as well as NO3 transport into sediments when it can be identified. Source: modified from
CENR2000.
Despite several decades of progress in reducing nutrient pollution from waste treatment facilities,
nutrient runoff from farms and metropolitan areas, often far inland, has gone unabated or actually
increased (The Pew Oceans Commission: www.pewoceans.org; Marine Pollution in the United States:
Significant Accomplishments, Future Challenges, 2001; Mitsch et al. 2001). Interestingly, early marine
scientists considered nutrients as a resource, not a problem (Brandt 1901), and reflected on ways to
fertilize coastal seas to increase biological production. In fact, in the 1890s Brandt concluded that N was
the primary limiting nutrient in marine waters and that nitrification and denitrification were important
processes in the N cycle.
Nutrient overenrichment of estuaries and nearshore coastal waters from human-based causes is now
recognized as a national problem on the basis of CWA 305b reports from coastal States that list waters
whose use or uses are impaired; these figures vary from 25% to 50% of the waters surveyed. The
National Oceanic and Atmospheric Administration's (NOAA) National Estuarine Eutrophication
Assessment (Bricker et al. 1999) indicated that about 60% of the estuaries out of 138 surveyed exhibited
moderate to serious overenrichment conditions. Nutrient overenrichment of coastal seas now has
international implications (NRC 2000) and is especially well documented for coastal systems of Europe
1-6
Nutrient Criteria—Estuarine and Coastal Waters
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External Nutrient Inputs Primary
and Susceptibility Symptoms
Influence of
Physical and
Nitrogen and Biological
Phosphorus -*• Processes
(i.e. freshwater
inflow, flushing,
wetlands uptake,
filter feeders)
Decreased
-*- Light Availability
Secondary Potential Effects and
Symptoms Use Impairments
. Loss of Habitat
^ Submeraed ^ Commercial Fishing
Aquatic Recreational Fishing
Vegetation Tourism
Algal 1 NuisancefToxic 1
7*- D°m'nance p* Algal Blooms |—
Changes 1 1
Increased 1
~~*" Organic Matterl
Production 1
Low
^ Dissolved
Oxygen
Increase of Algal Toxins
Commercial Fishing
i— »• Recreational Fishing
Human Health Problems
Swimming
Tourism
Fish Kills
». Commercial Fishing
Recreational Fishing
^ Aesthetic Values
Tourism
Loss of Habitat
^ Commercial Fishing
Recreational Fishing
Tourism
Offensive Odors
~*~ Aesthetic Values
Tourism
D^.COffiRGTOlJpDELlNOAACA*DS EUTROPHICATION SURVEY DATA CORE GROUP
Figure 1-3. Expanded nutrient enrichment model. Source: Bricker et al. 1999.
(Justic 1987, Jansson and Dahlberg 1999, Gerlach 1990, cited in Patsch and Radach 1997, Radach 1992),
Australia (McComb and Humphries 1992), and Japan (Okaichi 1997). The problem is likely
underreported for developing nations. Currently, the European Union has initiated an effort to develop
nutrient criteria for surrounding fresh and marine waters (personal communication, U. Claussen, German
Environmental Protection Agency).
In summary, these examples demonstrate that both N and P may limit phytoplankton biomass production
depending on season, location along the salinity gradient, and other factors. Nutrient overenrichment
problems have been present from early history, especially in estuaries downstream of cities, and the
nutrient criteria development approach that follows is a new element in EPA's effort to address these
longstanding problems.
Nutrient Criteria—Estuarine and Coastal Waters
1-7
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1.4 THE NUTRIENT CRITERIA DEVELOPMENT PROCESS
Preliminary Steps
It is impossible to recommend a single national criterion applicable to all estuaries. Natural enrichment
varies throughout the geographic and geological regions of the country, and these subdivisions must be
considered in the development of appropriate nutrient criteria. For example, "drowned river estuaries"
may exhibit a range of inherent or ambient natural enrichment conditions from less than 1.3 (iM TP in the
thin soils of the Northeast to 2.6 (iM TP in the delta regions of the South and Gulf of Mexico.
Although lakes and reservoirs and streams and rivers may be subdivided by classes, allowing reference
conditions for each class and facilitating cost-effective criteria development for nutrient management,
except for barrier island estuaries and mangrove bays in a given area this is not feasible for estuaries. A
major distinction between this manual and the one prepared for lakes and reservoirs is that estuarine and
coastal marine waters tend to be far more unique, and development of individual waterbody criteria
rather than for classes of waterbodies (such as glacial temperate lakes) is a greater likelihood. Also,
estuaries will likely require classification by residence time or subdivision by salinity or density
gradients.
Consequently, it will be necessary in many cases to determine the natural ambient background nutrient
condition for each estuary or coastal area so that the eutrophication caused by human development and
abuse can be addressed. Human-caused eutrophication is the focus of this manual, but the development
of nutrient criteria, frequently on a waterbody-specific basis, will require another major distinction for
coastal marine criteria development. In the absence of comparable reference waterbodies, the historical
record of inherent and cultural enrichment may be particularly significant to developing reference
conditions of a particular estuary or coastal reach. The historical perspective is always important to
criteria development, but in this instance it may also be essential to reference condition determination.
An outline of the recommended process for coastal and estuarine criteria development is as follows: (1)
Investigation of historical information to reveal the nutrient quality in the past and to deduce the ambient,
natural nutrient levels associated with a period of lesser cultural eutrophication, (2) determination of
present-day or historical reference conditions for the waterbody segment based on the least affected sites
remaining, such as areas of minimally developed shoreline, of least intrusive use, fed by those tributaries
of least developed watersheds, (3) use of loading and hydrologic models to best understand the density
and flow gradients, including tides, affecting the nutrient concentrations, (4) the best interpretation of
this information by the regional specialists and Regional Technical Assistance Group (RTAG)
responsible for developing the criteria, and (5) consideration of the consequences of any proposed
criteria on the coastal marine waters that ultimately receive these nutrients to ensure that the developed
criteria provide for the attainment and maintenance of these coastal uses. This concept, as illustrated in
Figure 1-4, is the basis for the National Nutrient Criteria Program and is explained throughout this text.
In deriving the reference condition (Figure 1-5), the extreme values of hypereutrophy on one hand and
pristine or presettlement conditions on the other can be estimated from monitoring, historical records,
1-8 Nutrient Criteria — Estuarine and Coastal Waters
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Historical
Information
Describing
Trends
Reference
Condition
• TP
.TN
• Algal Biomass
• Water Clarity
• Other Variables
Models
Describing
System
Dynamics
Assessment by
Regional Technical
Assistance Group
• Evaluation for Downsteam Effects
NUTRIENT CRITERIA
Figure 1-4. Elements of nutrient criteria development and their relationships in
the process.
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Area of Increasing Cultural Eutrophication
Regional Reference Condition
Range of
Potential
Criterion
Value
Pristine, Pre-Settlement Conditions
Figure 1-5. Derivation of the reference condition and the National Nutrient Criteria Program using TP,
TN, and chlorophyll a as example variables. Clarity or Secchi depth would be on a reversed scale.
Protectivity nutrient criteria should be between pristine conditions and present reference conditions, i.e., the
most "natural" attainable.
Nutrient Criteria—Estuarine and Coastal Waters
1-9
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and paleoecological determinations. The reference condition and the derived criteria are scientifically
based estimates expected to be a present-day approximation of the natural state of the waters approaching
but not likely duplicating pristine conditions. They include a conscious decision to use areas of least
human impact as indicators of low cultural eutrophication. A measure of practical judgment is also
necessary where scientific methods and data are not adequate.
The use of minimally impacted reference sites has been adapted from biological criteria development and
is endorsed by EPA's Science Advisory Board (U.S. EPA 1992). Minimal impacts provide a baseline
that should protect beneficial uses of the Nation's waters. The term "minimally impacted" implies a high
percentage of conditions in reference locations and a low percentage of conditions in all locations (i.e.,
some enrichment is allowed, but not enough to cause adverse local effects or adverse coastal receiving
water effects). The upper end of the data distribution range from reference sites represents the threshold
of a reference condition, whereas lower percentiles represent high-quality conditions that may not or
cannot be achieved. The upper 25th percentile represents an appropriate margin of safety to add to the
minimum threshold, excludes the effect of spurious outliers, and serves as a sufficiently protective value.
Where sufficient data are available, comparison and statistical analysis of causal and response variables
can help determine effect thresholds and further refine reference conditions (see Figure 6-2).
Establishing the reference condition is but one element of the criteria development process. Reference
condition values are appropriately modified on the basis of examination of the historical record (most
important), modeling, expert judgment, and consideration of downstream effects.
Strategy for Reducing Human-Based Eutrophication
Six key elements are associated with the strategy for reducing human-based eutrophication (U.S. EPA
1998):
EPA believes that nutrient criteria need to be established on an individual estuarine or coastal water
system basis and must be appropriate to each waterbody type. They should not consist of a single
set of national numbers or values because there is simply too much natural variation from one part
of the country to another. Similarly, the expression of nutrient enrichment and its measurement
vary from one waterbody type to another. For example, streams do not respond to phosphorus and
nitrogen in the same way that lakes, estuaries or coastal waters.
• Consequently, EPA has prepared guidance for these criteria on a waterbody-type and
region-specific basis. With detailed manuals available for data gathering, criteria development, and
management response, the goal is for States and Tribes to develop criteria to help them deal with
nutrient overenrichment of their waters and protect designated uses.
• To help achieve this goal, the Agency has initiated a system of EPA regional technical and financial
support operations, each led by a Regional Nutrient Coordinator—a specialist responsible for
providing the help and guidance necessary for States or Tribes in his or her region to develop and
adopt criteria. These coordinators are guided and assisted in their duties by a team of inter-Agency
1-10 Nutrient Criteria—Estuarine and Coastal Waters
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and intra-Agency specialists from EPA headquarters. This team provides both technical and
financial support to the RTAGs created by these coordinators so the job can be completed and
communication maintained between the policymaking in headquarters and the actual environmental
management in the regions.
EPA will develop basic ecoregional coastal ocean province nutrient criteria for waterbody types.
The Regional Teams and States/Tribes can use these values to develop criteria protective of
designated uses; the Agency also may use these values if it elects to promulgate criteria for a State
or Tribe. These criteria, once adopted by States and authorized Tribes into water quality standards,
will have value in two contexts: (1) as decisionmaking benchmarks for management planning and
assessment and (2) as the basis of National Pollution Discharge Elimination System (NPDES)
permit limits and Total Maximum Daily Load (TMDL) target values. The Standards and Health
Protection Division of the EPA Office of Water will be developing implementation guidance for
these latter applications.
EPA plans to provide sufficient information for States and Tribes to begin adopting nutrient
standards by 2003.
States/Tribes are expected to monitor and evaluate the effectiveness of nutrient management
programs implemented on the basis of the nutrient criteria. EPA intends the criteria guidance to
reflect the "natural," minimally impaired condition of a given estuary or coastal water or the class
of these systems, respectively. Once water quality standards are established for nutrients on the
basis of these criteria, the relative success or failure of any management effort, either protection or
remediation, can be evaluated.
Thus, the six elements of the National Nutrient Criteria Program describe a process that encompasses
taking measurements of the collective water resources of an area, establishing nutrient criteria for
evaluating the discrete waters within that region, assessing individual waterbodies against these criteria
and associated standards, designing and implementing the appropriate management, and, finally,
evaluating its relative success.
Nutrient Criteria Development Process
The activities that compose the nutrient criteria development process are listed below in the order
generally followed, and the subsequent chapters of this document follow this sequence. Figure 1-6
presents a schematic illustration of the process with parallel, corresponding chapter headings.
Nutrient Criteria—Estuarine and Coastal Waters 1-11
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Chapters 1-3
Chapter 4
Chapter 5
Chapter 6
Chapter 7
Chapter 8
>
t
RTAG
Ecoregion / Coastal
Province Evaluation
|
I
Physic
Variables & Methods
Selection
Existing Databases
Sampling Design &
New Data Collection
Data Analysis
J
Develop Reference
Conditions
H
Consider Additional
Elements
Criteria
Development
Using Nutrient Criteria to
Protect Water Quality
Figure 1-6. Flowchart of the nutrient criteria development process.
1-12
Nutrient Criteria—Estuarine and Coastal Waters
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• Preliminary Steps for Criteria Development (Chapter 1)
Establishment of Regional Technical Assistance Groups
The Regional Nutrient Coordinator in each EPA multistate region should obtain the involvement of key
specialists (e.g., estuarine and marine ecologists, water resource managers, oceanographers, stream and
wetland ecologists, water chemists, and agricultural and land-use specialists) with respect to the
waterbodies of concern. These experts should be recruited from other Federal and State agencies.
Experts from academia and industry may serve as technical advisors on an as need basis but not official
voting members of the RTAG.
Particular Federal agencies of interest are the U.S. Geological Survey (USGS); Natural Resources
Conservation Service (NRCS); National Oceanic and Atmospheric Administration (NOAA); National
Marine Fisheries Service (NMFS) and National Ocean Survey (NOS); U.S. Department of the Interior;
National Park Service (NPS); National Seashores; the U.S. Fish and Wildlife Service (USFWS); U.S.
Army Corps of Engineers (USAGE); and, in certain areas of the country the Bureau of Land Management
(BLM) or special government agencies such as river basin commissions and inter-State commissions.
Similarly, for information and education activities, the National Sea Grant Program and for agriculture,
the USDA Cooperative Extension Service are valuable resources.
State agencies with responsibilities relevant to this effort are variously named, but are commonly referred
to as Department of Natural Resources, Department of Water Resources, Department of the
Environment, Department of Environmental Management, Fisheries and Wildlife Management, State
Department of Agriculture, State Department of Forestry, or other land-use management agencies. Most
state land-grant universities have faculty talent important to natural resource and nutrient management,
and almost all colleges and universities have applied science faculty with research interests and talents
appropriate to this initiative.
In selecting participants for the group, diverse expertise is an obvious prerequisite, but willingness to
cooperate in the group effort, integrity, and a lack of a strong alternative interest are also important
factors to consider in selecting these essential people who must make collective and sometimes difficult
determinations.
The experts chosen will constitute the RTAG, which will be responsible for developing more refined
nutrient criteria guidance for their respective estuaries and coastal waters. The RTAG should be large
enough to have the necessary breadth of experience, but small enough to effectively debate and resolve
serious scientific and management issues. A membership of about 30 approaches an unwieldy size,
although that number may initially be necessary to maintain an effective working group of half that size.
EPA expects that States and authorized Tribes will use the information developed by the RTAGs when
adopting nutrient criteria into their water quality standards. The RTAG is intended to be composed of
scientists and resource managers from Federal agencies and their State counterparts. The RTAG should
not delegate its responsibility with the private sector. The perspectives of private citizens, academicians,
and special interest groups are important, and these and other members of the public may attend RTAG
Nutrient Criteria—Estuarine and Coastal Waters 1-13
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meetings and offer opinions when invited, but the final deliberations and decisions are the responsibility
of the Federal and State members of the RTAG—the States when adopting nutrient criteria into their
water quality standards, and the EPA when determining whether to approve or disapprove such criteria.
They must also be able to meet and debate the issues without undue outside influence.
As a matter of policy, however, EPA encourages the RTAGs to regularly provide access and reports to
the public. The meetings should generally be open to the public and the schedule of those meetings
published in the local newspapers. At a minimum, RTAGs are encouraged to hold regular "stakeholders"
meetings so that environmental, industrial, and other interests may participate via a separate public forum
associated with responding to the group's efforts. It is important that citizens and public groups be
involved, and any significant determinations of the RTAG should include a public session at which a
current account of activities and determinations is presented and comments acknowledged and
considered. In addition, where specific land uses or practices are addressed, those property owners,
farmers, fishermen, or other involved parties should be consulted in the deliberation and decisionmaking
process.
It is reasonable to expect the RTAG to meet monthly, or at least quarterly, with working assignments and
assessments conducted between these meetings. To coordinate activities among the 10 RTAGS, and with
the National Nutrients Team, regular conference calls are recommended. At these sessions, new
developments in the Program, technical innovations and experiences, budgets, and policy evolutions will
be conveyed and discussed. In the same context, an annual meeting of all Regional Nutrient
Coordinators, State representatives, and involved Federal agencies should be held each spring in or near
Washington, DC. At this meeting, major technical reports are presented by specialists and issues
significant to the Program are discussed.
The composition and coordination discussed above are intended to establish the shortest possible line of
communication between the State, region, and national Program staff members to promote a rapid but
reasoned response to changing issues and techniques affecting nutrient management of our waters. This
format is also designed to be responsive to the water resource user community without becoming a part
of user conflicts.
Delineation of Nutrient Ecoregions/Coastal Province Appropriate to the Development of Criteria
The initial step in this process has been taken through the creation of a national nutrient ecoregion map
consisting of 14 North American subdivisions of the coterminous United States (Figure 1-1). These are
aggregations of Level III ecoregions revised by Omernik (2000). Alaska, Hawaii, and the U.S.
Territories will be subdivided into nutrient ecoregions later, with the advice and assistance of those
States and their governments.
The initial responsibility of each RTAG will be to evaluate the present ecoregional map with respect to
variability on the basis of detailed observations and data available from the States and Tribes in that EPA
region. This preliminary assessment will further depend on the additional nutrient water quality data
1-14 Nutrient Criteria—Estuarine and Coastal Waters
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obtained by those States. The databases, especially with respect to selected reference sites, may be used
to refine the initial boundaries of the map in each EPA region.
EPA recognizes that the coastal margins of these ecoregions will be of the greatest concern to the States
developing estuarine and coastal marine criteria, but in some instances watersheds will extend a
considerable distance inland. In any case, the consistent application of the ecoregion concept facilitates
both upstream and inland coordination by the RTAGs and States and integrates the coastal efforts with
rivers, lakes, and streams.
• Scientific Basis (Chapter 2)
Chapter 2 emphasizes the role of physical processes interacting with biological processes in modulating
the expression of nutrient enrichment effects and the potential of inaccurately assessing cause and effects
in developing management plans.
• Physical Classification (Chapter 3)
The next step in evaluating the data is to devise a classification scheme for rationally subdividing the
population of estuarine and coastal marine waters in the State or Tribal territory. Because identification
of overenrichment is the objective of nutrient criteria development, trophic classification per se should be
avoided, as should any classification based on levels of human development. Physical characteristics
independent of most human-caused enrichment sources are far more appropriate.
However, as stated above, many estuarine and some coastal marine areas will probably require individual
attention and development of reference conditions that are site-specific or at least specific to waterbody
segments. Within these contiguous segments, the reference stations should have similar residence time,
salinity, general water chemistry characteristics, depth, and grain size or bottom type.
Once the waters have been subdivided and classified, it is important to select the key indicator variables
of concern and determine how much information is available on the enrichment status of these stations.
• Selection of Indicator Variables (Chapter 4)
Chapters 4 through 7 describe the variables for which EPA anticipates developing 304 (a) criteria for
nutrients in estuaries and coastal waters and how they should be sampled, preserved, and analyzed.
Although a wide variety of indicator variables may be possible, this technical manual describes
development of numerical criteria for total phosphorus (TP) and total nitrogen (TN) as primary nutrient
causal variables of eutrophication, and measures of algal biomass (e.g., chlorophyll a for phytoplankton
and ash-free dry weight for macroalgae) and a measure of water clarity (e.g., Secchi depth or electronic
photometers) as primary variables of eutrophic response. In those systems that have hypoxia or anoxia
problems, dissolved oxygen also should be added as a primary response variable. States or Tribes may
elect to include other indicators as well, but the four primary variables and dissolved oxygen as indicated
are recommended as the essential indicators. Other variables are loss of seagrass/submerged aquatic
vegetation (SAV), benthic macroinfauna, iron, and silica as well as other indicators of primary and
secondary productivity.
Nutrient Criteria—Estuarine and Coastal Waters 1-15
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State and Federal agency records are the basis for an initial data search. In many States, water quality
information resides in more than one agency. For example, Maryland has a Department of Natural
Resources and a Department of the Environment, both of which retain water quality records. To
compound the data search problem further, States may also have pertinent data sets in their Department
of Fisheries and Department of Public Health. It is wise to initiate the search for information with calls
and questionnaires to colleagues in the State or Tribal agencies likely to be involved so an appropriate
list of contacts and data sets can be compiled. In doing so, regional Federal agencies should not be
overlooked either. These include the agencies described above in the selection of RTAG members.
• Nutrient Data Collection and Assessment (Chapter 5)
EPA has initiated the data collection and assessment process by screening the existing STORET and
ODES databases for information on lakes, reservoirs, streams, estuaries and coastal waters with respect
to the four initial parameters, and dissolved oxygen where appropriate (see reference to Chapter 4 above).
These primary variables were originally selected for robustness and conservativeness of estimation;
however, the preliminary screening of the STORET data revealed that these measurements are also
relatively abundant in the database.
Although this is an entirely appropriate starting point for nutrient criteria development, States and Tribes
are not required to confine their investigations and data selection to only these variables. States and
Tribes are encouraged to select additional measures that contribute to the best assessment of the
enrichment of their regional waters and protect designated uses. In particular, it is advisable to use both
causal indicators and response indicators as mentioned above.
Combining nutrient and biological system response information will yield the most definitive and
comprehensive criteria. To use only causal or only response variables in the criteria puts the State or
Tribe in jeopardy of not protecting the designated uses. For example, a highly enriched estuarine system
with a rapid flushing rate may appear to be in attainment when only the biota and dissolved oxygen are
measured, but the load of nutrients being delivered downstream in its coastal discharge plume is
degrading the receiving waters. Using a balanced combination of both causal and response variables in
the criteria, together with careful attention to tidal and seasonal variability, should mitigate against false-
positive or false-negative results.
Chapters 4 and 5 both discuss proper sampling, preservation, and analysis of samples. Seasonality,
spatial distribution of sample sites, composite versus discrete sampling, and fixed station versus stratified
random sampling are also explored.
Establishing an Appropriate Database
Review of Historical Information. Historical information, including sediment core analysis, is important
to establish a perspective on the condition of a given waterbody. Has its condition changed radically in
recent years? Is the system stable over time? What is the variability? Has there been a trend up or down
in trophic condition? Only an assessment of the historical record can provide these answers. Without
this information, the manager risks setting reference conditions and subsequent criteria on the basis of
1-16 Nutrient Criteria—Estuarine and Coastal Waters
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present condition alone, which may in fact be a degraded state. Valid historical information places the
current information in its proper perspective and is particularly important to coastal and estuarine
nutrient criteria development because of the difficulty in establishing classes and the scarcity of reference
waterbodies.
Data Screening. The first step in assessing historical or current data is to review the material to
determine its suitability to support nutrient criteria development. Anecdotal information and
observations are valuable, but the sources must be carefully considered. Fishermen's accounts, local
sport-fishing news stories, and observational logs of scientific field crews are all legitimate sources of
information, but they are subject to different levels of scrutiny before a trend is determined. The same
applies to databases. Nutrient information gathered for identifying failing wastewater treatment plants
cannot be assessed in the same light as similar data collected to determine overall water quality or trophic
state. The analytical procedures used, type of sampling design and equipment, and sample preservation
are other variables that must also be considered in any data review and compilation. Once this screening
is done, the compiled data may be sorted according to station location, physical characteristics, relative
depth, time, and date, and then analyzed for the establishment of reference conditions.
• Establishing Reference Conditions (Chapter 6)
Candidate reference locations can be determined from compiled data with the help of regional experts
familiar with the waters of the area. Classification will be an important first step and should be based on
physical characteristics of the waterbodies, including morphology, geological origin, and hydrologic
factors such as residence time, flow characteristics, tidal processes, and freshwater-saltwater
interchanges. An estuary may then be subclassified into lower, medium, and upper salinity regimes.
Specialists can also help to select the least culturally impacted sites or stations within each area.
Three candidate approaches are recommended for development of tidal estuarine reference conditions.
Two more approaches use loading information within the fluvial watershed. A sixth approach is
described for coastal waters. Where several replicate systems occur, each classified as near-pristine
based on recent data (e.g., past 10 years), then one can apply a frequency distribution approach, and this
manual recommends that the upper 75th percentile be used as a starting point. If some minor nutrient
enrichment is present, then all the data would be considered and, in this case, the lower 25th percentile is
suggested. In the case of significant nutrient-based environmental degradation, where reference sites
cannot be identified from current monitoring data, then hind-casting with ambient data is recommended.
There are three approaches: (1) empirical in situ data analysis, (2) sediment core or paleoecological
analysis, and (3) model hind-casting. Interpretation of this approach is potentially sensitive to
confounding by physical factors (e.g., freshwater inflows). The watershed approach is load-based. Here,
one attempts to locate a relatively nutrient-unenriched tributary, or stream segment, that is approximately
representative of the watershed, and extrapolate the nutrient load for the entire watershed. This can be
done empirically or, preferably, with models. The coastal approach focuses on changes in the nutrient
regime of estuarine plumes and waters some distance from such plumes. An index approach is described
that accounts for variability and facilitates identification of natural enrichment (e.g., upwelling). Long-
term monitoring is required to distinguish anthropogenic effects from natural variability.
Nutrient Criteria—Estuarine and Coastal Waters 1-17
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• Criteria Development (Chapter 7)
Nutrient Criteria Components
The move from data review and data gathering to criteria development involves a sequence of five
interrelated elements:
• Examination of the historical record or paleoecological evidence for evidence of a trend.
Determination of a reference condition using one of several alternative approaches. Remember that
the reference condition, however derived, is only part of the criteria development process.
• Use of empirical modeling or surrogate data sets in some instances where insufficient information
exists. This may be the case especially in estuaries with insufficient hydrological data, or
significantly developed or modified watersheds.
Objective and comprehensive interpretation of all of this information by a panel of specialists
selected for this purpose (i.e., the RTAG). These experts should have established regional
reputations and expertise in a variety of complementary fields such as oceanography, estuarine
ecology, nutrient chemistry, and water resource and fisheries management.
• Finally, the criterion developed for each variable should reflect the optimal nutrient condition for
the waterbody in the absence of cultural impacts and protect the designated use of that waterbody.
Second, it must be reviewed to ensure that the proposed level does not entail adverse nutrient
loadings to downstream waterbodies. In designating uses for a waterbody and developing criteria
to protect those uses, the State or Tribe must consider the water quality standards of downstream
waters (40 CFR 131.10 (b)). This concern extends all the way to coastal waters, but in practice the
immediate downstream receiving waters are the area of greatest attention for the resource manager.
The criteria must provide for the attainment and maintenance of standards in downstream waters.
A criterion for that estuary or subclass of estuary will not protect downstream water quality
standards, it should be revised accordingly.
Once the initial criteria (either Regional or State/Tribal) have been selected, they can be verified and
calibrated by testing the sampling and analytical methods and criteria values against waterbodies of
known conditions. This ensures that the system operates as expected. This calibration can be
accomplished either by field trials or by use of an existing database of assured quality. This process may
lead to refinements of either the techniques or the criteria.
Criteria are developed for more than one parameter. For example, all reference sites of a given class may
be determined to manifest characteristics of a particular level for TP concentration, TN concentration,
algal biomass,and water clarity. These four measures, and dissolved oxygen as appropriate, become the
basis for criteria appropriate to optimal nutrient quality and the protection of designated uses. The policy
for criteria attainment will be developed by the State or Tribe in consultation with EPA.
1-18 Nutrient Criteria—Estuarine and Coastal Waters
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When the estuarine or coastal marine segment in question reveals high TN and TP concentrations, but not
the expected high algal biomass and low water clarity, further investigation is indicated before deciding
whether criteria have been met. Flushing rates, inorganic turbidity, water color, or toxins may be
additional factors influencing the condition of the estuary.
Assessing Attainment With Criteria
An action level then is established for the nutrient criteria that have been selected for each indicator
variable. The list includes two causal variables (TN and TP) and three primary response variables (e.g.,
when dissolved oxygen problems occur this will add an additional variable to the response variables.
Failure to meet either of the causal criteria should be sufficient to prompt action. However, if the causal
criteria are met, but some combination of response criteria are not met, there should be some form of
decision making protocol to resolve the question of whether the waters in question meet the nutrient
criteria. There are two approaches to this:
• Establish a decisionmaking rule equating all of the criteria such as the frequency and duration of
exceedences and the critical combination of response variables requisite for action
• Establish an index that accomplishes the same result by inserting the data into an equation that
relates the multiple variables in a nondimensional comprehensive score
• Management Response (Chapter 8)
There are a variety of possible management responses to the overenrichment problem identified by
nutrient criteria. Chapter 8 describes some regulatory and nonregulatory processes that involve the
application of nutrient criteria. It also presents a 10-step process that allows the resource manager to use
these approaches to improve water resource condition. The emphasis is on developing a scientifically
responsible, practical, and cost-effective management plan.
The chapter also describes three basic categories that encompass all management activities: education,
funding, and regulation. It closes with the admonition to always carefully evaluate the success of the
management project, report results, and continue monitoring the status of the water resource.
• Model Applications (Chapter 9)
A variety of empirical and theoretical models are described and discussed, and two specific illustrations
of the application of models to estuarine nutrient management are presented.
• Appendices
A number of appendices supplement the primary text.
It should be noted that completion of each step may not be required of all water quality managers. Many
State or Tribal water quality agencies may have already completed the identification of designated uses,
classified their estuaries and coastal waters, or established monitoring programs and/or databases for
their programs and therefore can bypass those steps. This manual is meant to be comprehensive in the
Nutrient Criteria—Estuarine and Coastal Waters 1-19
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sense that all of the criteria development steps are described; however, the process can be adapted to suit
existing water quality programs.
In any event, a responsible nutrient management plan should meet three conditions. First, the plan and its
component elements must be scientifically defensible; otherwise it might lead to well-intentioned
management actions that are unnecessary or harmful. This is like the admonition to physicians, "above
all do no harm." Second, effective nutrient management must strive to be economically feasible. The
public and local interests are more likely to support approaches that provide meaningful benefit
compared with their cost. Finally, these approaches should be practical and acceptable to the
communities involved. The approaches should address appropriate social and political issues, such as
conflicts that might exist between public agencies and landowners, agricultural or other resource users, or
between commercial fishermen and recreationists and environmental or industrial groups. Any
management plan may fail if these three general elements are not sufficiently addressed, and it is almost
certain to fail if they are all ignored.
1-20 Nutrient Criteria—Estuarine and Coastal Waters
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Controlling the Right Nutrients
Physical Processes, Salinity, Algal Net Primary Production
Nutrient Loads and Concentrations: Interpretation of Effects
Physical-Chemical Processes and Dissolved Oxygen Deficiency
Nutrient Overenrichment Effects and Important Biological
Resources
Scientific Basis for
Estuarine and Coastal
Waters Quantitative
_T . „ . Concluding Statement Regarding Nitrogen and Phosphorus
Nutrient Criteria
2.1 INTRODUCTION
At the turn of the last century nitrogen and phosphorus were prized as the fuel that fed the great engine
of marine production. Today they are seen as lethal pollutants leading to toxic blooms and suffocation.
Just as weeds are fine plants growing in the wrong place, nitrogen and phosphorus are essential
chemicals that can get into the wrong places at the wrong times. We should not lose sight of their
critical role in sustaining production (Nixon 2000).
Purpose and Overview
This chapter describes the scientific basis for development of nutrient criteria for estuarine and coastal
waters. A number of scientific issues are addressed to develop nutrient criteria. Water quality managers
can improve their application of science to nutrient criteria development if they consider these systems'
large latitudinal and climatic range, high ecosystem-based variability, complexity, diversity, and broad
range in land-sea margin human activities. These features suggest a high degree of system individuality,
especially at larger scales. These features occur because estuaries and coastal waters are transitional
ecosystems buffeted by variable landward-based freshwater input volumes and constituents, influences of
oceanic provinces, and human disturbances, including nutrient enrichment, superimposed on these
natural regimes (Figure 2-1). Even in a relatively narrow section of coastline, the ecosystem diversity
and variability may be quite large. These characteristics challenge the investigator to develop useful
predictive schemes. Some progress has been achieved, but areas of important uncertainties are also
noted.
Coastal areas, including estuaries and upwelling regions, account for only 10% of the ocean by area but
at least 25% of the ocean's primary productivity and upwards of 95% of the world's estimated fishery
yield (Walsh 1988). These areas are also an important organic carbon sink of atmospheric CO2 In
addition, coastal counties account for only 17% of the U.S. landmass, but their population exceeds 141
million. Thus, more than half of the Nation's population lives in less than one-fifth of the total area, and
this trend is expected to grow (NRC 2000). These statistics underpin the fact that estuarine and open
coastal areas have, and continue to show, stress from human activities including nutrient pollution, as
noted in Chapter 1. These demographics argue strongly for a scientific understanding of how nutrients
flux through estuarine and nearshore coastal ecosystems and impair water quality use.
Nutrient Criteria—Estuarine and Coastal Waters 2-1
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Coastal Zone
Tidal
Front
1. River
2. Estuary
3. Coastal boundary layer
4. Shelf proper
5. Shelfbreak
Slope Sediments
Figure 2-1. Idealized scheme defining the coastal ocean and the coastal zone, with some key biochemical fluxes
linking land and sea and pelagic and benthic processes. The latter are not to scale. Source: Alongi 1998.
Some Important Nutrient-Related Scientific Issues
A large number of issues with a scientific component may complicate nutrient criteria development in
estuaries and open coastal waters. Some of the more important issues are summarized below and are
discussed in more detail later in this and following chapters. These issues illustrate how science
underpins nutrient criteria development.
Determination of which nutrients are causing the problem is critical. In some cases, this will be known
with considerable assurance, but in others further study is advisable. Without such knowledge, it is
difficult to develop reliable nutrient criteria. It is important to understand at what scale one is discussing
the question of nutrient limitation. The term "nutrient limitation" is often used quite loosely and without
formal definition (Howarth 1988). For phytoplankton, Howarth makes the following points and argues
that it matters a great deal which of the following questions is being addressed:
• Limitation of the growth rate of phytoplankton populations currently in a waterbody
Limitation of the potential rate of net primary production, allowing for possible shifts in the
composition of phytoplankton species
Limitation of net ecosystem production
2-2
Nutrient Criteria—Estuarine and Coastal Waters
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Each of these definitions can be considered "correct," but each addresses different questions. Clearly,
phytoplankton growing in an oligotrophic environment may be adapted to maximize growth rates under
low nutrient conditions, as evidenced by their organic nutrient composition approaching the Redfield
atomic ratio of C:N:P of 106:16:1 (Redfield 1958; Goldman et al. 1979). An increase in nutrient supply
would likely shift species composition to those adapted to the higher nutrient regime, and net primary
production would potentially increase. Thus, it is plausible that potential net primary production can be
nutrient limited even if the growth rate of currently dominant phytoplankton species is not. If a nutrient
is added to a system and net primary production increases, the system is considered to have been nutrient
limited regardless of whether the species composition has shifted. Similarly, when a nutrient criterion is
exceeded, enrichment is presumed to be of concern even if the system's productivity has not responded.
This is the definition used in this manual for addressing effects of nutrient overenrichment.
Why not use net ecosystem production as the preferred definition, as the ecosystem is the level of system
organization that might seem most relevant? For example, the ecosystem was the level of the whole-lake
experiments that contributed to defining P as the primary limiting nutrient for north temperate freshwater
lakes (Schindler 1977). Net ecosystem production equals gross primary production in excess of total
ecosystem respiration. For the biomass of an isolated ecosystem to be maintained, the net ecosystem
organic production must equal or slightly exceed 0. Imports of organic matter can augment the internal
net production. Howarth argues that it is difficult to relate nutrient supplies to net ecosystem production
because the respiration term is sensitive to allochthonous input of organic matter as well as internal net
production. So, for practical reasons, net primary production, which is directly related to algal biomass
production, is the preferred measure of nutrient limitation.
The import of organic matter, especially in estuaries, can lead to water quality problems (e.g., hypoxia).
Organic matter input from sewage was historically a major source of organic carbon that drove aquatic
systems toward dissolved oxygen (DO) deficiency through direct microbial heterotrophic activity
(Capper 1983). However, the input of nutrients, whether in organic form followed by recycling or
inorganic form with direct nutrient uptake, is what stimulates potential phytoplankton biomass
production, and this organic matter may contribute to symptoms of nutrient overenrichment identified in
Chapter 1.
It is frequently difficult to distinguish natural ecosystem variability associated with net primary
production from that induced by anthropogenic stress, especially nutrient enrichment, which often is a
consequence of variability in physical processes. An example is the difficulty, even with a 50-year
record, in distinguishing the effects of freshwater flow of the Susquehanna River and co-linear effects of
nutrient loading on Chesapeake Bay phytoplankton biomass production indicated by chlorophyll a (chl a)
concentrations (Harding and Perry 1997). Such indeterminancy is a condition that water quality
managers must contend with, and argues for broad scientific input.
It is important to understand nutrient load and ecological response relationships because of the need to
conduct load allocations (e.g., total maximum daily loads, TMDLs), and it may be necessary to perform
some management triage when systems are poised along a gradient of risk and there are too many
Nutrient Criteria—Estuarine and Coastal Waters 2-3
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systems to treat in a timely fashion. Also, as explained later, ecological responses to nutrient enrichment
may be quantitatively related to nutrient load rather than complexity in physical transport and mixing.
The relationship between N load and seagrass recovery in Tampa Bay, FL, is an example of where
nutrient load was predictive but concentration of N was not (Greening et al. 1997).
As discussed in Chapter 3, classification of estuaries and coastal shelf systems at large scale (e.g.,
Chesapeake Bay versus Delaware Bay) is in an early state of development with regard to predicting many
nutrient enrichment effects. This is because of the relatively high degree of ecosystem individuality at
the larger scale, where comparability among systems tends to breaks down. The result is that scientific
generalizations are usually circumscribed with consequences that may lead to higher management costs.
Resource managers and environmental scientists should work together to improve predictability of
nutrient enrichment effects because there are too many systems in the Nation to study all estuaries and
coastal systems comprehensively.
These ecosystems exhibit a notable degree of process asymmetry and lag in responses, which means that
a stress at one location and time may show up as a response at another location and time. Additionally,
different mechanisms may result in a similar response (Malone et al. 1999). This type of behavior
enhances the tendency to confound cause-and-effect relationships.
Along the same lines, conceptual models for estuaries (and coastal waters) in particular are still evolving.
These models suggest that systems modulate stresses so that a single stress does not necessarily result in
a single response (Cloern 2001) (Figure 2-2). This fact alone contributes to ecological uncertainty in
load-response relationships. Conceptual models help define expectations of cause-and-effect
relationships and degree of nutrient-caused impairment, and refine hypotheses. Conceptual models
should be a standard tool for water quality managers.
Antecedent conditions are important. This can be understood in terms of whether enough factors are
present at the right place and time to lead to an integrated response, such as a dinoflagellate bloom. Such
conditions resemble nonlinear dynamics, which may be a major constraint to prediction of effects. Also,
estuaries and nearshore coastal waters are subject to episodic events, which injects considerable
uncertainty into predictions (e.g., Tropical Storm Agnes impacted Chesapeake Bay in June 1972: Davis
and Laird 1976). A relatively large database is often required to determine when effects of such major
events have reached a new steady state.
Estuaries and nearshore coastal waters naturally vary in the type, abundance, and geographical coverage
of biological communities at risk to nutrient overenrichment, largely because of habitat differences. This
variability is partially offset by salinity, which tends to "normalize" biotic community distributions
(Kinne 1964). When ambient historical data are unavailable or sediment cores are ineffective in
2-4 Nutrient Criteria—Estuarine and Coastal Waters
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Nutrient Loading \
Filter
Changes In:
Chlorophyll
Primary Production
Macroalgal Biomass
Sedimentation of Organic C
Si:N and N:P Ratios
Toxic-harmful Algal Blooms
Phytoplankton Community
^Responses,.
Reversible
Changes In:
Benthos Biomass
Benthos Community
Vascular Plants
Habitat Quality/Diversity
Water Transparency
Organic C in Sediments
Sediment Biogeochemistry
BottonrMvater Dissolved Oxygen
Seasonal Cycles
Mortality of Fish/Invertebrates
Nutrient Cycling
Food web Structure
Figure 2-2. Schematic representation of the contemporary (Phase II) conceptual model
of coastal eutrophication. Advances in recent decades include explicit recognition of (1)
a complex suite of both direct and indirect responses to change in nutrient inputs; (2)
system attributes that act as a filter to modulate these responses; and (3) the possibility of
ecosystem rehabilitation through appropriate management actions to reduce nutrient
inputs to sensitive coastal ecosystems. Source: Cloern 2001.
characterizing resources lost through nutrient overenrichment, it is often difficult to establish an accurate
historical reference or determine the potential recovery from nutrient stress. Apparently, many estuaries
became moderately to highly enriched before effective monitoring programs provided accurate
descriptive information on biotic community distributions and abundance. When all else fails,
professional judgment should be used to estimate reference conditions.
Finally, water quality managers should anticipate that nutrient enrichment will act with other stressors
and forms of ecosystem disturbance and modify their respective ecological expressions (Breitburg et al.
1999).
These considerations suggest that water quality managers may face a large array of uncertainties
regarding nutrient criteria development and implementation for estuaries and nearshore coastal waters.
This manual attempts to guide application of established scientific principles and to reveal important
uncertainties that bear on nutrient criteria development. This chapter begins with a contextual discussion
of the watershed perspective characterized as the "river-to-ocean continuum."
River-to-Ocean Continuum: Watershed/Nearshore Coastal Management Framework
This section describes the physical relationship of estuaries and nearshore coastal waters to their
respective water and sedimentary boundaries. This description provides a context for understanding
problems of nutrient overenrichment in coastal ecosystems. Estuaries and nearshore coastal systems
share some features, but important differences reflect how nutrients cause problems.
Nutrient Criteria—Estuarine and Coastal Waters
2-5
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Some Important Identifying Features of Estuaries (adapted partly from Cloern 1996)
1. Estuaries are located between freshwater ecosystems (lakes, rivers, and streams; freshwater and
coastal wetlands; and groundwater systems) and coastal shelf systems (Figure 2-2). These
ecological boundary conditions create a transition between contrasting freshwater and open-ocean
ecosystems.
2. Estuaries are relatively shallow; often, on average, only a few meters to a few tens of meters deep.
This promotes a strong benthic-pelagic coupling that influences nutrient cycling through changes in
system nutrient stoichiometry. A well-developed benthic community participates in nutrient
cycling.
3. River-influenced estuaries are quite different from systems. Vertical mixing is regulated primarily
by the seasonal cycle of heat input and thermal stratification that retards vertical mixing. However,
in estuaries vertical mixing is regulated by a larger and more variable source of buoyancy: the
riverine input of freshwater that acts to stabilize the water column. Also, freshwater input
establishes longitudinal and vertical salinity gradients and drives nontidal gravitational circulation,
a major contributor to flushing.
4. Estuaries are particle-rich relative to coastal systems and have physical mechanisms that tend to
retain particles. These suspended particles mediate a number of activities (e.g., absorbing and
scattering light, or absorbing hydroscopic materials such as phosphate and toxic contaminants).
New particles enter with river flow and may be resuspended from the bottom by tidal currents and
wind-wave activity.
5. Many estuaries are naturally nutrient-rich because of inputs from the land surface and geochemical
and biological processes that act as "filters" to retain nutrients within estuaries (Kennedy 1984).
Variability in freshwater discharge is reflected in the estuarine salinity gradient, which has important
consequences for stenohaline organisms, especially nonmotile forms. The salinity gradient of estuaries
has been classified by on the Venice System, and salinity classes approximate the distribution of many
estuarine organisms (Figure 2-3). Changes in salinity (e.g., wet and dry decadal periods) often modify
population distributions and biotic community structure (Carriker 1967). Rivers and lakes process
nutrients and modify nutrient ecological stoichiometry before the material arrives downstream, where
receiving coastal waters further nutrient cycling (Billen et al. 1991). Nutrient cycling occurs along the
continuum; phytoplankton and other algae are key agents of biochemical change (Redfield 1963) (Figure
2-4). Redfield et al. (1958) demonstrated that phytoplankton in active growth phase tend to maintain a
C:N:P ratio close to 106:16:1. Annual rates of net primary production in coastal shelf environments tend
to overlap rates of estuaries, but coastal shelves on average are somewhat lower in magnitude, except in
upwelling areas where rates may, on average, exceed those of estuaries by a factor of two to three (Walsh
1988) (Table 2-1).
2-6 Nutrient Criteria—Estuarine and Coastal Waters
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Fresh Water
Oligohaline
Mesohaline
Polyhaline Euhaline
Venice System
Other Zone
Characterizations
Salinity (ppt)
Nontidal Fresh Tidal Fresh Low Brackish Moderately Brackish
RIVER
ESTUARY
OCEAN
Figure 2-3. Salinity zones. The Venice System is a well-accepted method of characterizing salinity zones and
covers the salinity ranges from riverine regions to the ocean. The freshwater category in the Venice System has been
modified in this atlas to account for the tidal and nontidal regions found in rivers with estuarine portions. Source:
Lippson et al. 1979, Environmental Atlas of the Potomac Estuary, MD Department of Natural Resources.
OCEAN N
RIVER
Reactive
inorganic
substances
Small pelagic
heterotrophs
Particulate
organic matter
Consumption
Regeneration
Consumption
Phytoplankton
X7
Invertabrate
larvae, copepods,
rotifers, etc.
Metazoan
food
chain
Consumption
Sedimentation
Regeneration
Denitrification
I I
Bacteria, crustaceans,
ascidians, polychaetes,
mollusks, nematodes, etc.
Benthic
heterotrophs
Figure 2-4. Schematic illustrating the central role of phytoplankton as agents of biogeochemical change in shallow
coastal ecosystems. Phytoplankton assimilate reactive inorganic substances and incorporate these into paniculate
(POM) and dissolved organic matter (DOM) which support the production of pelagic and benthic heterotrophs.
Arrows indicate some of the material fluxes between these different compartments. Denitrification has been added to
the figure. Source: Cloern 1996.
Nutrient Criteria—Estuarine and Coastal Waters
2-7
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Table 2-1. Categorization of the world's continental shelves based on location, major river,
and primary productivity
Latitude
(°)
Region
Major Primary Production
River (g Cm 2 yr1)
Eastern Boundary Current
0-30
30-60
Ecuador-Chile
Southwest Africa
Northwest Africa
Baja California
Somali coast
Arabian Sea
California- Washington
Portugal-Morocco
-
-
-
-
Juba
Indus
Columbia
Tagus
1000-2000
1000-2000
200-500
600
175
200
150-200
60-290
Western Boundary Currents
0-30
Mesotrophic Systems
30-60
Brazil
Gulf of Guinea
Oman/Persian Gulfs
Bay of Bengal
Andaman Sea
Java/Banda Seas
Timor Sea
Coral Sea
Arafura Sea
Red Sea
Mozambique Channel
South China Sea
Caribbean Sea
Central America
West Florida shelf
South Atlantic Bight
Australian Bight
New Zealand
Amazon
Congo
Tigris
Ganges
Irrawaddy
Brantas
Fitzroy
Fly
Mitchell
Awash
Zambesi
Mekong
Orinoco
Magdalena
Appalachicola
Altamaha
Murray
Waikato
90
130
80
110
50
110
100
20-175
150
35
100-150
215-317
66-139
180
30
130-350
50-70
115
2-8
Nutrient Criteria—Estuarine and Coastal Waters
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Table 2-1. Categorization of the world's continental shelves based on location, major river, and primary
productivity (continued)
Latitude Region
(°)
Argentina-Uraguay
Southern Chile
Southern Mediterranean
Gulf of Alaska
Nova Scotia-Maine
Labrador Sea
Okhotsk Sea
Bering Sea
Phototrophic Systems
60-90 Beaufort Sea
Chukchi Sea
East Siberian Sea
Laptev Sea
Kara Sea
Barents Sea
Greenland-Norwegian Seas
Weddell-Ross Seas
Eutrophic Systems
30-60 Mid-Atlantic Bight
Baltic Sea
East China Sea
Sea of Japan
North-Irish Sea
Adriatic Sea
Caspian Sea
Black Sea
Bay of Biscay
Louisiana/Texas shelf
Major Primary Production
River (g Cm 2 yr1)
Parana
Valdivia
Nile
Fraser
St. Lawrence
Churchill
Amur
Kuskokwim
Mackenzie
Yukon
Kolyma
Lena
Ob
Pechora
Tjorsa
-
Hudson
Vistula
Yangtze
Ishikari
Rhine
Po
Volga
Danube
Loire
Mississippi
70
90
30-45
50
130
24-100
75
170
10-20
40-180
70
70
70
25-96
40-60
12-86
300-380
75-150
170
100-200
100-250
68-85
100
50-150
120
100
Source: Adapted from Walsh, with additional data from Alongi, and Postma and Zijlistra.
Nutrient Criteria—Estuarine and Coastal Waters
2-9
-------
Some Identifying Features ofNearshore Coastal Waters
1. Nearshore coastal waters extend from the coastal baseline at high tide and across the mouths of
estuaries to approximately three nautical miles. Coastal waters are relatively deep compared to
estuaries with depths ranging from a few meters to several hundred meters, depending on coastal
location.
2. Coastal longshore currents are a principal mechanism to exchange water masses.
3. Upwelling of nutrients from the deep ocean can be locally important.
4. Nearshore coastal systems tend to be particle-rich compared to the open ocean, but much less so
than adjoining estuaries.
5. Nearshore coastal systems have a weaker benthic-pelagic coupling than estuaries mainly because
they are deeper.
Coastal environments in the continental United States show only modest levels of upwelling compared to
well-known upwelling areas, such as coastal Ecuador-Chile. The Gulf Stream, which flows
northeastward along the South Atlantic coast from the Florida Straits to North Carolina, lies close enough
to the shoreline to affect water temperature and circulation of nearshore waters. Dynamic core rings that
slide off to the mainland side of the Gulf Stream affect local conditions. The coastal environment is
dynamic in terms of phytoplankton bloom formation and dissipation (Walsh 1988). This has relevance to
characterization of reference conditions and monitoring for nutrient criteria performance because the
systems, though not as physically dynamic at short temporal scales as estuaries, are still difficult to assess
in terms of average conditions. Synoptic survey tools such as aerial surveillance with fixed-wing aircraft
and satellites can provide wide coverage, including short-term phytoplankton dynamics.
2.2 CONTROLLING THE RIGHT NUTRIENTS
Overview
Chapter 1 introduced the geographical extent and magnitude of the overenrichment problem and
suggested the importance of nitrogen (N) versus phosphorus (P) as limiting nutrients. Several recent
review papers (Downing 1997, Smith 1998, Smith et al. 1999, Conley 2000) and the NRC (2000) volume
concluded that the major nutrients causing overenrichment problems (e.g., algal blooms) in estuaries and
nearshore coastal waters are N and P. Silica (Si) may limit diatom production at relatively high levels of
N and P. Iron is a co-limiting nutrient in some ocean areas and may exert some limitation in shelf
waters,but its importance in open coastal waters usually is secondary to N (NRC 2000). Additionally, P
limits primary production in some tropical nearshore habitats, although study of these systems is limited
(Howarth et al. 1995). Often the addition of both N and P will elicit greater phytoplankton biomass
stimulation than the sum of both nutrients added separately (Fisher et al. 1992). There are reported cases
where both N and P are required to elicit a phytoplankton biomass production response in estuaries
(Flemer et al. 1998), suggesting that N and P supply rates were equally limiting. Tropical lagoons, with
2-10 Nutrient Criteria—Estuarine and Coastal Waters
-------
carbonate sands low in P and unaffected by human activity, also are prone to P limitation. For example,
the seagrass Thalassia testudinum was P-limited in Florida Bay (Powell et al. 1989, Fourqurean et al.
1992a,b).
Tidal fresh and brackish waters in many estuaries typically are more light limited than higher saline
waters (Flemer 1970, Sin et al. 1999). As freshwater fluxes seaward, processes operate to modify
nutrient stoichiometry (e.g., sedimentation of P-absorbed particles, denitrification, and differential
microbial decomposition). A number of temperate estuaries exhibit seasonal shifts in nutrient limitation
with winter-spring P limitation and summer-fall N limitation (D'Elia et al. 1986; Fisher et al. 1992,
Malone et al. 1996) (Table 2-2). The Redfield ratio (C:N:P) of marine benthic plants approximates
550:30:1, substantially richer in organic carbon, much of which is structural material, and indicates that
these plants require less N and P than do phytoplankton (Atkinson and Smith 1982). In summary, the
foregoing results suggest that both N and P criteria are needed, depending on season and local ecosystem
conditions (Conley 2000).
Some Empirical Evidence for N Limitation of Net Primary Production
Three case studies provide some of the strongest evidence available that water quality mangers should
focus on N for criteria development and environmental control (see NRC 2000 for details). One study
involves work in large mesocosms by the University of Rhode Island (Marine Ecosystem Research
Laboratory-MERL) on the shore of Narragansett Bay. Experiments showed that P addition was not
stimulatory, but N or N+P caused large increases in the rate of net primary production and phytoplankton
standing crops (Oviatt et al. 1995).
In another study, nutrient releases from a sewage treatment plant were monitored in the Himmerfjarden
Estuary south of Stockholm, Sweden, on the Baltic Sea (Elmgren and Larsson 1997). Throughout a 17-
year field experiment (i.e., whole-ecosystem study), the concentration of total N tended to reflect the N
input from the sewage treatment plant, and both abundances of phytoplankton and water clarity were
clearly related to the total N concentration and not to total P. This experiment involved independent
increases and decreases in N and P over the observation period.
A third whole-ecosystem study involved long-term changes in Laholm Bay, Sweden (Rosenberg et al.
1990). Early signs of overenrichment appeared in the 1950s and 1960s and steadily increased over time
(Figure 2-5). Among the earliest reported signs were changes in the composition of macroalgal species.
Over time the filamentous algae typical of enriched conditions became more prevalent, and harmful algal
blooms (HABs) became more common during the 1980s. These changes correlated best with changes
over the decades in N loads rather than P loads. These field studies are excellent examples of the power
of long-term monitoring of nutrient and biological variables in estuaries (Wolfe et al. 1987). Importantly,
these three ecosystem experiments correlated well with short-term bioassay experiments and ratios of
dissolved inorganic N:P ratios in these ecosystems (NRC 2000). The above whole-system field
experiments and the large preponderance of bioassay data in estuaries and nearshore coastal systems
(Howarth 1988) and generally low inorganic N:P atomic ratios at peak primary production (Boynton et
al. 1982) make a strong case for the widespread importance of N as a controlling nutrient for net coastal
Nutrient Criteria—Estuarine and Coastal Waters 2-11
-------
Table 2-2. Estuaries exhibiting seasonal shifts in nutrient limitation with
spring P limitation and summer N limitation
Estuary
Reference
Baltic Sea
Himmerfjarden Estuary, Sweden
Gulf of Riga, Latvia
Roskilde Fjord, Denmark
Bay of Brest, France2
Chesapeake Bay, USAa
Mainstem
Patuxent River Estuary
York River Estuary
Rhode River Estuary
Delaware Estuary, USA
Neuse River Estuary, USA
Graneli et al. 1990, Elmgren & Larsson 1997
Maestrini et al. 1997
Pedersen&Boruml996
DelAmoetal. 1997
Maloneetal. 1996
D'Eliaetal. 1986
Webb 1988
Gallegos& Jordan 1997
Pennock & Sharp 1994
Mallin&Paerll994
" Systems displaying seasonal dissolved silicate limitation.
Source: Conley 2000.
8000
7000
6000
D Nitrogen
+ Phosphorus
Exceptional
plankton blooms
Early indications
of eutrophicatio
800
700
600
500
400
300
200
100
0
1950
1960
1970
Year
1980
1990
Figure 2-5. Transport of nutrients to Laholm Bay, Sweden. Periods of significant
changes in the marine biota are also indicated (modified from Rosenberg et al. 1990).
Source: NRC2000.
2-12
Nutrient Criteria—Estuarine and Coastal Waters
-------
marine primary production and a major contributor to water quality problems. Interpretation of nutrient
ratios was initially applied in the open ocean by Redfield (1934) and further elaborated on by Redfield
(1958) and Redfield et al. (1963). Boynton et al. suggested that when inorganic N:P ratios for a variety
of estuarine systems are interpreted, atomic ratios less than 10 indicated N limitation and ratios greater
than 20 indicated P limitation (Figure 2-6). Some have suggested that it matters whether the inorganic N
is in the form of ammonium- or nitrate-N. High concentrations of ammonia-N may inhibit nitrate-N
uptake; however, Dortch (1990) reported that this phenomenon is more variable than widely believed.
Figure 2-7 summarizes major factors that determine whether N or P is more limiting in aquatic
ecosystems where one of these macronutrients is limiting net primary production.
Some Threshold Responses to Nitrogen Overenrichment
Kelly (in press) summarized several generalizations that appear to hold for N overenrichment in
estuaries. Over a range of average dissolved inorganic nitrogen (DIN) from <1 to >20 (iM, chlorophyll a
tends to increase at slightly less than 1 (ig/L with every 1 (iM increase in DIN or approximately about
0.75 (ig chl/(iM DIN (e.g., see Figure 3-2b in Chapter 3). Evidence is especially strong thatN
concentrations can reduce or eliminate growth of estuarine submerged aquatic vegetation (SAV) and
higher salinity seagrasses (Sand-Jensen and Borum 1991; Dennison et al. 1993; Duarte 1995) by both
water column shading and epiphytic overgrowth. Estuarine SAV and seagrasses tend to show light
limitation when surface insolation approximates 11% at the surface of the canopy, but this figure varies
between about 5% and 20% depending on species. Stevenson et al. (1993) transplanted plugs ofRuppia
maritima, Potamogeton perfoliatus, and P. pectinatus in different areas of the Choptank Estuary,
Chesapeake Bay, and reported that survival thresholds occurred when total suspended solids were
between ~15 and 20 mg/L, chlorophyll a was 15 (ig/L, DIN was below 10 (iM, and PO4 was below 0.35
(iM. Kelly (in press) reviewed a number of studies and suggested that an approximate threshold for
hypoxia occurred at about 80 (iM TN (Table 2-3) (normalized TN loading for residence time expressed
in years and divided by depth). These relationships document the importance of N as a major cause of
estuarine water quality impairment. Also, these ecological response thresholds are a useful rule of
thumb, but some deviations are to be expected. In data-poor estuaries, such thresholds are a first-order
target until more adequate data can be developed to establish reference conditions.
Although overenrichment from N causes many symptoms of marine water quality impairment, it is the
interaction of biogeochemical, biological, and physical processes that modulate the effects of a particular
N supply (Cloern 2001) (Appendix A). These relationships had their genesis in the late 19th and early
20th Centuries in northern Europe, especially in German and Scandinavian marine research institutes
(Mills 1989). Water quality managers who understand this interplay will assess cause-and-effect
relationships with a deeper insight. Knowledge of algal nutrient physiology is necessary information, but
it alone is insufficient to explain why blooms occur.
Effects of Physical Forcing on Net Primary Production
Each physical forcing (e.g., river inflows, wind velocity, irradiance, water temperature, and tidal
currents) contributes to phytoplankton population variability by influencing rates of vertical mixing,
Nutrient Criteria—Estuarine and Coastal Waters 2-13
-------
0
NitrogerrPhosphorus Ratios
10 20 30
40
RIVER DOMINATED
Pamlico River, North Carolina
Narragansett Bay, Rhode Island
Western Wadden Sea, Netherlands
Eastern Wadden Sea, Netherlands
Mid-Patuxent River, Maryland
Long Island Sound, New York
Lower San Francisco Bay, California
Barataria Bay, Louisiana
Upper San Francisco Bay, California
Victoria Harbor, British Columbia
Mid-Chesapeake Bay, Maryland
Duwamish River, Washington
Upper Patuxent River, Maryland
Hudson River, New York
Apalachicola Bay, Florida
Upper Chesapeake Bay, Maryland
EMBAYMENTS
Roskeeda Bay, Ireland A
Bedford Basin, Nova Scotia
Southeast Kaneohe Bay, Hawaii
St. Margarets Bay, Nova Scotia
Central Kaneohe Bay, Hawaii
Vostok Bay, former USSR
LAGOONS
Beaufort Sound, North Carolina
Chincoteague Bay, Maryland
Peconic Bay, New York
High Venice Lagoon, Italy
FJORDS
Baltic Sea
Loch Etive, Scotland
120
225
c
g
'•^
'co
IS
o o
o -
"CO
3
'V *-
52 100
240
V
48
190
125
0
10
20
30
40
Figure 2-6. Summary of nitrogen:phosphorus ratios in 28 sample estuarine ecosystems.
Horizontal bars indicate the annual ranges in nitrogen:phosphorus ratios; solid triangles
represent the ratio at the time of maximum productivity. Vertical bands represent the
typical range of algal composition ratios (modified from Boynton et al. 1982). Source:
NRC 2000.
2-14
Nutrient Criteria—Estuarine and Coastal Waters
-------
Nitrogen fixation (N2)
Phytoplankton
n itrogen: phosphorus
16:1
16:1
Water column
nitrogen: phosphorus
Recycling
External loading
Water residence time
Sediments
Pore water
N:P
Denitrification (N2)
Burial
Figure 2-7. Factors that determine whether nitrogen or phosphorus is more limiting in aquatic
ecosystems, where one of these macronutrients is limiting to net primary production.
Phytoplankton use nitrogen and phosphorus in the approximate molar ratio of 16:1. The ratio of
available nitrogen in the water column is affected by: (1) ratio of nitrogen:phosphorus in external
inputs to the ecosystem; (2) relative rates of recycling of nitrogen and phosphorus in the water
column, with organic phosphorus usually cycling faster than organic nitrogen; (3) differential
sedimentation of nitrogen in more oligotrophic systems; (4) preferential return of nitrogen or
phosphorus from sediments to the water column due to processes such as denitrification and
phosphorus adsorption and precipitation; and (5) nitrogen fixation (modified from Howarth 1988;
Howarthetal. 1995). Source: NRC2000.
Nutrient Criteria—Estuarine and Coastal Waters
2-15
-------
Table 2-3. DO, nutrient loading, and other characteristics for selected coastal areas and a MERL mesocosm
enrichment experiment (source: Kelly in press)
System
Experimental"
MERL-control
MERL- IX
MERL-2X
MERL-4X
MERL-8X
MERL-16X
MERL-32X
Field"1
Baltic Sea e
Scheldt
Chesapeake Bay f-8
Potomac River f
Guadalupe estuary h
Ochlocknee Bay
Delaware Bay
Narragansett Bay '
Providence River '
Providence Riv.j-k
Boston Harbor '
N. Outer Harbor m
N. Gulf of Mexico °
Area
(m2)
2.63
2.63
2.63
2.63
2.63
2.63
2.63
(km2)
374,600
277
11,542
1,210
551
551
24
1,989
328
24.13
24.13
103
13
20,000
Depth
Avg
(m)
5
5
5
5
5
5
5
55
11.2
6
5.9
1.4
1.4
1
9.7
8.3
3.7
3.7
5.5
10
30
Annual TN
loading
(mmol m 2)
800
1,750
2,950
4,850
9,000
18,500
34,000
217
13,400
938
2,095
548
2,058
5,995
1,900
1,960
13,600
13,600
21,600
107,692
6500
Res. Time
(mo)
0.9
0.9
0.9
0.9
0.9
0.9
0.9
250
3
7.6
5
10
1
0.1
4
0.9
0.083
0.233
0.266
0.03
6°
DO
Status"
OK
OK
OK
OK
-H
H
A
H/A
H/A
A
H/A
9
9
OK
OK
OK
H
H
-H
OK
H/A
Vertical
Mixing
Status
mixed
mixed
mixed
mixed
mixed
mixed
mixed
stratified
??
stratified
stratified
??
??
stratified
weak strat
stratified
stratified
weak strat
mixed
stratified
Normalized
TN Loading
(uM)b
12
26
44
72
133
274
503
81
295
98
146
322
121
49
64
17
25
70
86
27
107
Primary
Production
(gCm'y1)
190(100)
270(115)
305 (243)
515(305)
420(171)
900(601)
1150(901)
-149-170
9
-380 to 520
(361-858)
-290 to 325
9
9
9
-200 to 400
270 to 290
9
9
9
263 to 546
-290 to 320
*H= hypoxia, A= anoxia.
'Volumetric TN loading is normalized for residence time to yield an "expected" or potential concentration. The value is calculated as: Annual
TN Loading * Residence time (expressed in years) divided by Depth. Units are thus mmol/m3, or uM. See Kelly 1997a,b; 1998. The value is
not decremented for denitrification or burial, removal processes that have greater effect on concentrations in longer residence time systems (cf.
Nixon et al. 1996, Kelly 1998).
'See Nixon et al. 1984, Oviatt et al. 1986, Nixon 1992, Nixon et al. 1996. DIN was used to enrich treatment conditions (e.g. 1X...32X) and is
represented in Figures 5, 6, and 7. TN values include input of organic forms with feedwater, which is only a substantial portion of input at the
control and the low end of the enrichment gradient. Production for year 1 of experiment was extrapolated using empirical model of Keller 1988,
which did not include measurements of primary production above 600 g C m"2/1 (Nixon 1992). These values are used in Figures 6 and 7.
Parenthetical production values for year 2 are from Keller 1988. Hypoxic and anoxic events were periodic, not chronic.
""Except for Providence River, Boston Harbor and Gulf of Mexico, loading is TN as reported by Nixon et al. 1996. With noted exceptions for
individual systems below, see Nixon (1992, 1997) for productivity references.
'Also see Elmgren 1989, Cederwall and Elmgren 1990, Rosenberg et al. 1990. Table value for TN loading from Nixon et al. 1996 is lower than
DIN input in Nixon 1997 plot, which included N input across the halocline. Lower value is labeled in
Figure 6.
fAlso see Boynton et al. 1995, Boynton and Kemp 2000; historical Chesapeake production range (parenthetical) is from Boynton et al. 1982.
2-16
Nutrient Criteria—Estuarine and Coastal Waters
-------
Table 2-3. DO, nutrient loading, and other characteristics for selected coastal areas and a MERL mesocosm
enrichment experiment (continued)
8Mainstem stratification, increasing anoxic extent; Officer et al. 1984, Boynton and Kemp 2000.
"Top line is for dry flow, bottom line is for wet flow.
'Only strongly stratified by freshwater at head of Bay in Providence River area, see notes j, k below. Production range is from Nixon 1997 (does
not include historical presettlement estimate of 120-130 g C m"2/1).
jOviatt et al. 1984, Doering et al. 1990, Asselin and Spaulding 1993; TN loading from seaward and landward inputs, avg residence time (2.5 d),
low DO in 13-15 m channel.
kUses longer 7-d residence time during very low flow conditions, Asselin and Spauling 1993.
'TN budget indues direct estimate of ocean loading as well as land loading. Nixon et al. 1996 gave a preliminary budget; table shows improved
budget of Kelly 1998. Freshwater stratification and near hypoxia/occasional hypoxia only occur in inner harbor. See Signell and Butman 1992
for flushing estimate of whole harbor.
"Northern harbor section, Kelly 1998. Harbor station production of Kelly and Doering 1997.
"Area represents greatest measured extent of hypoxic zone. Higher production is for immediate plume (Rabalais et al. 2000). TN loading is to a
20,000-km2 hypoxic zone only (and thus is a maximal rate) based on Mississippi/Atchafalaya input of 130 x 10' moles y"1 (Howarth et al. 1996;
Turner and Rabalais 1991). Rate is consistent with long-term average (1980-1996) estimated by CENR 2000 of 1,567,900 metric tons y1.
"Assumed a 6-mo residence time (-seasonal turnover) for illustration only; if longer, then normalized concentration would increase accordingly.
Nutrient Criteria—Estuarine and Coastal Waters 2-17
-------
sedimentation, horizontal transport, production, and grazing. Each forcing has its characteristic timescale
of variability (e.g., 12.4-hr tidal period, the diel 24-hr light cycle, several days to weeks-long storm
events of enhanced river flow and wind stress, and seasonal cycles of irradiance and temperature; Cloern
1996).
Phytoplankton growth depends on nutrient supplies, as expected, but growth is significantly modulated
by complex physical processes that operate at virtually every physical scale (Giller et al. 1994). For this
reason, it is desirable for RTAGs and State water quality managers to have ready access to individuals
with a specialty in physical oceanography.
In estuaries, bottom topography and bathymetry form the basin in which tidal currents, freshwater inflow,
and wind vectors act as principal drivers of estuarine and coastal physical processes and contribute to
variability in mixing and circulation of waters (Cloern 1996) (Figure 2-8). Physical processes can
attenuate or exacerbate nutrient enrichment effects depending on the form of interaction. For example,
the Delaware River Estuary receives TN and TP loads somewhat larger than does the mainstem
Chesapeake Bay, yet the Delaware Estuary has lower phytoplankton production and does not have a
hypoxia problem, largely because of its relatively strong vertical mixing (i.e., a weak vertical density
stratification) and horizontal water exchange with the open ocean system (Pennock 1985).
Freshwater inflow is the "master driver" that defines the ecological character of river-dominated
estuaries. Boynton and Kemp (2000) proposed a simple conceptual model to explain effects of river flow
on Chesapeake Bay ecological processes associated with nutrient inputs (Figure 2-9). These authors
stated:
The importance of freshwater inputs is obvious; it is a central feature in the definition of estuarine
systems, it influences physical dynamics (Boicourt 1992), is well correlated with nutrient inputs
(Summers 1993), and has been implicated in regulating either directly or indirectly estuarine
processes ranging from primary production (Boynton et al. 1982; Cloern et al. 1983) to benthic
secondary production (Flint 1985) to fish recruitment (Stevens 1977) and catch (Sutcliffe 1973;
Sutcliffe etal. 1977; Ennis 1986).
Boynton and Kemp applied regression techniques to datasets from mid-Chesapeake Bay, a mesohaline
area, to test the ideas represented in Figure 2-9. They showed that Susquehanna River flow was
significantly related to annual average primary production, annual average surface chlorophyll a, spring
deposition of total chlorophyll a per square meter, and total chlorophyll a deposition rate (meter squared
per day). They also showed that the decline in dissolved oxygen concentrations in deep water during the
spring bloom period was also related to flow (Figure 2-10). Although this relationship could be driven
by riverflow effects on stratification, which in turn regulates dissolved oxygen depletion, they argue that
river inputs of nutrients are of primary concern. This is because years of high and low stratification did
not correlate well to years of high and low rates of oxygen decline. The implication is that nutrient
enrichment played a key role in deep-water hypoxia.
2-18 Nutrient Criteria—Estuarine and Coastal Waters
-------
'
.,
t-Nd M »: T gf ,=34 ^, - l^^^J^-Ss -O' -
Figure 2-8. Cartoon diagrams of three physical forcings that operate at the
interface between SCEs and the coastal ocean (tides), watershed (river inflow), and
atmosphere (wind). Each physical forcing influences the growth rate of the
resident phytoplankton population through, for example, its influence on the
distribution of suspended sediments and turbidity. Each forcing also influences the
rate of vertical mixing, with riverine inputs of freshwater as a source of buoyancy
to stratify the water column and the tide and wind as sources of kinetic energy to
mix the water column. Each forcing is also a mechanism of water circulation that
transports phytoplankton horizontally. Much of the variability of phytoplankton
biomass during blooms can be understood as responses to fluctuations in these
interfacial forcings. Source: Cloern 1996.
Nutrient Criteria—Estuarine and Coastal Waters
2-19
-------
Geographic
Positioning of
Water Column
re: Benthos
1
Nutrient
Pools
Figure 2-9. Simple schematic diagram showing the influences of river flow on ecosystem
stocks and processes examined in this study. The mechanistic relationships between river
flow and the stocks and processes shown in the diagram are explained in the text. Source:
Boynton and Kemp 2000.
0.20
0.16
0.12
0.08
0.04
0.00
| y = 0.030 + 0.0096 X r2 = 0.76; p < O.oT]
1987
Hypoxia -15 May_
1985
1986$
H990
19891
1991
'1992
'Hypoxia - 7 July
1988
I
I
I
I
0
16
4 8 12
Total Chlorophyll a Deposition Rate
mg mj day"1
Figure 2-10. Scatter diagram showing the relationship between the rate of
decline in dissolved-oxygen concentrations in deep water (dDO dfl) and
average deposition rates of total chlorophyll a during the spring-bloom period.
Data are from the 1985-1992 period and were collected at the R-64 site. The
date on which hypoxia (DO concentration <1 mg I"1) was first encountered
during highest (1987) and lowest (1992) deposition years is also indicated.
20
2-20
Nutrient Criteria—Estuarine and Coastal Waters
-------
Freshwater inflow plays a major role in the degree of stratification (Figure 2-1 la-d) and nontidal flushing
(Figure 2-12) of estuaries. Density stratification influences the depth of vertical mixing relative to the
euphotic zone depth and the tendency toward hypoxia formation, that is, the effect of sealing off bottom
waters from reaeration. On a seasonal basis, stratification greatly influences the degree of hypoxia, but
seems to have a lesser role on an interannual scale (see above paragraph). Tidal displacement also
contributes to flushing (Figure 2-13). Numerous studies have documented the role of freshwater inflow
regulation of primary production through interaction with other estuarine processes via different
mechanisms (Pennock and Sharp 1994, Harding and Perry 1997, Cloern 1996, Sin et al. 1999). Freshets
deliver substantial quantities of nutrients to an estuary and lead to blooms (Mallin et al. 1993, Rudek et
al. 1991). Effects of rainfall operating on hydrographic processes have been shown to influence trophic
organization (Livingston 1997). A significant effect of episodic freshwater inflow is determining the
appropriate averaging period for reference conditions applicable to nutrient criteria development. The
issue applies to decadal wet and dry cycles as well. Water quality managers should anticipate that even
in estuaries relatively free of anthropogenic nutrient enrichment, some level of hypoxia may occur during
wet weather cycles. This "natural" condition, should it be observed will need to be factored into nutrient
criteria development.
Other Physical Factors
Other physical factors (e.g., salinity, temperature, and light) influence the expression of nutrient
enrichment effects and are extensively reported in standard textbooks. For example, salinity can
influence enrichment effects and can also influence biotic distributions (e.g., grazing populations),
primarily through the osmotic capabilities of resident organisms (Kinne 1964). Temperature and light
availability to photosynthetic organisms is obviously important. Temperature regulates, within certain
limits, the metabolic rates of organisms, especially poilkilotherms, and influences the distribution of
many species. Light also influences the feeding behavior of many planktonic animal forms, especially
crustacean filter feeders, which has relevance to algal grazing. Climatic factors influence phytoplankton
biomass production in estuaries (Lehman 2000). Additional information on the roles of temperature and
light as limiting factors to net primary production and effects of nutrient overenrichment is provided in
Appendix B.
2.3 NUTRIENT LOADS AND CONCENTRATIONS: INTERPRETATION
OF EFFECTS
The issue of whether or not to focus on nutrient concentration versus loading criteria has been a
contentious one among both scientists and managers. Whether or not to use concentrations or
loading as criteria largely depends on the spatial and temporal scales of assessing ecosystem
responses to nutrient inputs (H. Paerl, personal communication).
Nutrient Criteria—Estuarine and Coastal Waters 2-21
-------
Head of
Estuary,
Ocean-
B — Highly Stratified
C — Partially Mixed
Figure 2-lla-d. Schematic diagram of coastal plain estuary types, indicating direction
and degree of mixing. Arrows show direction of net mass transports of water, and the
arrow size indicates the relative magnitudes of the transports. Source: Lippson et al.
1979.
2-22
Nutrient Criteria—Estuarine and Coastal Waters
-------
I I = units of net downstream nontidal transport
I I = units of net upstream nontidal transport
Figure 2-12. Net transports in estuaries resulting from estuarine flows and mixing. At
any one point along an estuary, the difference between upstream- and downstream-
directed transports is equal to the freshwater input to that point. In this example with no
tributaries, the difference is equal to the input at the head of the estuary. Source: Lippson
etal. 1979.
Net Downstream Displacement
END
END
Flood Direction
Net Upstream Displacement
Figure 2-13. Net movement of a particle in each layer of a two-layered flow system.
Source: Lippson etal. 1979.
Nutrient Criteria—Estuarine and Coastal Waters
2-23
-------
Conceptual Framework
Nutrient concentrations are what phytoplankton (and other plants) respond to instantaneously or on very
short time scales. The dissolved inorganic and, to some extent, organic nutrient concentrations that
remain in a water parcel after a short period of phytoplankton growth are largely what is left over or
unused. (Note: Some dinoflagellates can obtain nutrients from particulate materials and exhibit other
complex forms of nutrition.) Nutrient uptake, including any luxuriant uptake, will be mostly converted
into organic form, given a suitable short period for growth. Thus, total concentration is a measure of the
nutrient in living form as well as any unused organic and inorganic forms. If concentrations of nutrients
are to be used as criteria, the total concentration is most likely to reflect the short-term phytoplankton
growth potential (Boynton and Kemp 2000).
Recycling is an important aspect of phytoplankton biomass production. If nutrients in a water parcel are
all converted into algal biomass, then maintaining the algal biomass requires rapid recycling or additional
supplies to the water parcel. With loss of phytoplankton from the water column through sedimentation,
grazing and conversion of phytoplankton to animal biomass, dispersion, and advection, maintenance and
any further net primary production require new supplies of nutrients. These processes all involve longer
time scales that include seasonal and interannual considerations of ecosystem water quality (i.e., use
impairments) and habitat response.
Examples
Some examples of regression relationships between nutrient load and concentration and response
variables are instructive because nutrient concentration often does not provide a useful relationship.
There is a range in the lag time between nutrient load and coastal water ecosystem responses. Such lags
have been reported for a number of estuaries, including the Patuxent (Kemp and Boynton 1984),
mainstem of the Chesapeake Bay (Malone et al. 1988), mesohaline York River estuary (Sin et al. 1999),
and Logan River and Moreton Bay, Australia (O'Donohue and Dennison 1997). Nixon et al. (1996)
developed a number of regressions between residence time and response variables (e.g., percent total N,
percent P exported, percent N retained from land and atmosphere, and percent N denitrified) from a
number of estuaries and coastal marine systems. Dettmann (in press) developed relationships somewhat
similar to those of Nixon et al. that included some different estuaries and coastal waters employing a
modified algebraic expression for residence time (e.g., Figure 2-14). The temporal scale of these
regressions typically ranges from months to annual averages. These regressions help frame causal
relationships but usually are not adequate by themselves to establish nutrient criteria. For example, the
Delaware Bay lies between the northern Adriatic Sea and Chesapeake Bay in terms of the fraction of N
exported, but the Delaware Bay has few symptoms of nutrient overenrichment.
For a number of coastal embayments in Virginia and Maryland, chlorophyll a concentration regressed on
a TN loading rate that was scaled to a unit area loading rate of the receiving waterbody surface area,
resulting in a relatively high R2 (Boynton et al. 1996). Peak chlorophyll a concentrations in the Potomac
Estuary regressed against peak TN load showed the highest chlorophyll a concentrations occurred under
average flow conditions (Boynton 1997). Maximum freshwater inflows resulted in a very strong density
stratification, but the nutrients were advected into the lower Chesapeake Bay, and thus no bloom formed
2-24 Nutrient Criteria—Estuarine and Coastal Waters
-------
Norsminde Fjord
oston Hbr.
» Observations
—Model
a = 0.30 mo-
r2 = 0.94
chlock-
onee Bay
uadalupe Estuary ('87)
Narragansett Bay 4
Westerschelde
. Adriatic Sea
Delaware Estuary
Potomac Estuary
• Guadalupe Estuary ('84)
Chesapeake Bay
Q1
1 10 100
Freshwater Residence Time (months)
1000
Figure 2-14. The fraction of landside nitrogen input exported from 11 North
American and European estuaries versus freshwater residence time (linear time
scale). Baltic Sea not shown. Source: Dettmann (in press).
in the lower Potomac estuary. Low freshwater inflows resulted in much weaker vertical density
stratification and apparently a low nutrient supply that limited phytoplankton bloom potential (Figure
2-15).
Using an interannual time scale, Harding (1994) summarized the historical (1950-1994) nutrient and
chlorophyll a trends for the mainstem of the Chesapeake Bay. Nitrogen, P, and chlorophyll a
concentrations increased considerably over the period of record. Harding and Perry (1997) applied a
statistical time series model and determined that confounding effects of freshwater inflow did not explain
the chlorophyll a increase in the lower bay. The DIN:DIP ratios suggested a greater influence of DIN as
a limiting nutrient to biomass production. Variation in the flow of the Susquehanna River over the
period of record tends to cloud the empirical relationships, especially in the oligohaline region and
brackish zone.
By inference, nutrients were hypothesized to be the principal causative agent. Since the 1970s, the
winter-spring freshet has been associated with a strong diatom bloom, and in 1989 a drought delayed
delivery of DIN and Si to the mesohaline reach of the bay until late spring, thus leading to a late-season
phytoplankton biomass increase composed primarily of flagellates.
Phytoplankton growth and biomass accumulation appear to be directly related to riverborne nutrient
inputs in the Chesapeake Bay (Boynton et al. 1982, Malone et al. 1988). Typically, years with higher
river flow (within limits) are marked by greater algal biomass, which supports elevated respiration and
more rapid depletion of bottom water DO in deep, stratified estuaries (Boicourt 1992). However, this
relationship is confounded by interannual variations in salinity stratification because stratification is
directly related to river flow (Seliger and Boggs 1988, Officer et al. 1984). Distinguishing between the
Nutrient Criteria—Estuarine and Coastal Waters
2-25
-------
^<
LU
Q_
Q_
P
o
LU
1°
o
60
50
40
30
20
10
0
(a)
1990
100
200
300
TN Load, kg/day x 1000
(1-2 month lag)
60
50
40
30
20
10
(b)
1990
small winter peak
100
200
300
400
500
PEAK TN LOAD, kg/day x 1000
Figure 2-15. Scatter plots of water column averaged chlorophyll a at a mesohaline station (MLE 2.2) versus
several different functions of total nitrogen (TN) loading rate measured at the fall line of the Potomac River
estuary. Source: Boynton 1997.
2-26
Nutrient Criteria—Estuarine and Coastal Waters
-------
effects of physical and biological processes on interannual variations in anoxia/hypoxia is only now
beginning on the basis of mathematical modeling and long-term empirical monitoring data. Stratification
from freshwater inflow from the Susquehanna River apparently is insufficient by itself to explain the
increased hypoxic volumes in the Chesapeake Bay from the early 1980s to 1999 (J. Hagy, personal
communication). In shallow estuaries the hypoxic volume, if present, is likely to be highly variable
spatially owing to the influence of variable freshwater inputs and estuarine in situ physical factors that
cause wide excursions and mixing of water masses (e.g., Neuse River estuary, H. Paerl, personal
communication).
A detailed study of nutrient and phytoplankton relationships in the mesohaline region of the mainstem of
the Chesapeake Bay demonstrated that "despite high inputs of DIN and dissolved silicate relative to DIP
(molar ratios of N:P and Si:P > 100), seasonal accumulations of phytoplankton biomass within the salt-
intruded reach of the bay appear to be limited by DIN supply while the magnitude of the spring diatom
bloom is governed by the dissolved Si supply" (Malone et al. 1996, Conley and Malone 1992). The
maximum chlorophyll-specific productivity occurred in the late summer, the maximum biomass occurred
in the spring, and volumetric-based productivity occurred in midsummer (see their Figure 4). This
temporal asymmetry leads to difficulties in ascribing simple empirical relationships between
phytoplankton biomass and nutrient concentrations.
2.4 PHYSICAL-CHEMICAL PROCESSES AND DISSOLVED OXYGEN DEFICIENCY
Dissolved oxygen deficiency, or hypoxia, is of critical importance to the health of aquatic life. The role
of physical processes, especially mixing and physical circulation of estuarine waters, has been widely
reported in the literature (Smith et al 1992). "There is no other environmental variable of such ecological
importance to coastal marine ecosystems that has changed so drastically in such period of time as
dissolved oxygen" (Diaz and Rosenberg (1995). One of the earliest studies to measure DO in a U.S.
estuary occurred in the Chesapeake Bay and Potomac River in 1912 (Sale and Skinner 1917),
approximately two decades after Winkler developed his now legendary method for determining the
concentration of DO in aquatic systems. Hypoxia was already present in the bottom waters of the lower
Potomac River estuary at this early date because a measurement indicated only a DO < 2.0 ml/L, or 35%
saturation.
Individual species exhibit a range in adaptability to relatively low DO concentrations (e.g., see "EPA
822-D-99-002 Draft Ambient Water Quality Criteria for Dissolved Oxygen [Saltwater]: Cape Cod to
Cape Hatteras"). Hypoxia and H2S apparently cause synergetic effects that make marine benthic animals
more sensitive to hypoxia when H2S is present (Diaz and Rosenberg 1995). These authors suggest that
the occurrence of hypoxia in shallow coastal and estuarine areas appears to be increasing, and evidence
suggests that the increase has global dimensions and seems most likely to be accelerated by human
activities (Nixon 1995, Bricker et al. 1999). Although hypoxia has undesirable consequences, when
bottom waters go anoxic wholesale biogeochemical changes occur. These changes can include release of
phosphate from sediments, emergence of highly toxic hydrogen sulfide, elimination of nearly all
Nutrient Criteria—Estuarine and Coastal Waters 2-27
-------
multicellular animals from sediment habitats, reduction in the coupled nitrification-denitrification, and
changes in metal solubilities, with many metals becoming toxic.
Diaz and Rosenberg (1995) concluded that should DO concentrations become slightly lower, catastrophic
events may overcome the systems and alter the productivity base that leads to economically important
fisheries and amenities. Aquatic biota exposed to low DO concentrations may be more susceptible to the
adverse effects of other stressors such as disease, toxic chemicals, and habitat modification (Holland
1977). Low DO conditions can increase the vulnerability of the benthos to predation, as the infaunal
animals extend above the sediment surface to obtain more oxygen (Holland et al. 1987). Dissolved
organic carbon apparently is a major carbon and energy source for bacteria (i.e., microbial loop; Azam et
al. 1983), whose metabolism is a major cause of hypoxia. Hypoxia and anoxia indicate that a coastal
ecosystem is severely stressed by nutrient overenrichment and should receive immediate attention by
water quality managers.
2.5 NUTRIENT OVERENRICHMENT EFFECTS AND IMPORTANT
BIOLOGICAL RESOURCES
Benthic Vascular Plant Responses to Nutrients
A major lesson learned over the past 25 years is that nutrient overenrichment has had a devastating effect
on SAV, whether estuarine species or higher salinity seagrasses. This conclusion is based on work
conducted mostly on the U.S. Gulf of Mexico and Atlantic Coasts (Tomasko et al. 1996, Tomasko and
LaPointe 1991, Kemp et al. 1983, Orth and Moore 1983, Burkholder et al. 1992, Taylor et al. 1995, Short
et al. 1995). Dennison et al. (1993) reported the following habitat criteria for SAV: DIN of 10.7 (iM,
DIP of 0.33 (iM; N:P (atomic) of 32; and chlorophyll a of 15 (ig/L. These criteria are being re-analyzed
by the EPA Chesapeake Bay Program.
The relationship between N load and concentration and chlorophyll a is not limited to phytoplankton.
Predictive regression relationships between N and chlorophyll a, water column light attenuation, and
seagrass recovery in Tampa Bay were found for N loading, not ambient N concentrations (Janicki and
Wade 1996, Greening et al. 1997). Tomasko et al. (1996) detected a negative correlation between N
loads and turtle grass (Thalassia testudinum) biomass and productivity in Sarasota Bay, FL.
Moore and Wetzel (2000) determined experimentally that eelgrass (Zostera marina) in the York River
estuary, lower Chesapeake Bay, is exposed to N concentrations adequate to stimulate enough epiphytic
growth to shade out this vascular plant. In mesocosms containing a complex of species characteristic of
shallow marine coastal lagoons along the Narragansett Bay coast, Taylor et al. (1995) showed that N
alone—but not P alone—caused an increase in water column concentrations of chlorophyll a and
particulate N, increased daytime net production, and increased growth of juvenile winter flounder.
Eelgrass beds and drift algae apparently were shaded out by phytoplankton at high nutrient levels.
Experiments conducted by Neundorfer and Kemp (1993) on the submersed plant Potamogeton
perfoliatus in microcosms using lower Choptank Estuary water demonstrated that effects of N and P on
algal densities were synergistic in that responses to N addition were greatest at high P loading and vice
2-28 Nutrient Criteria—Estuarine and Coastal Waters
-------
versa. Also, combined amendments (N+P) at highest treatment rates resulted in epiphytes and
phytoplankton increasing more than when these nutrients were added individually. On the basis of
microcosm studies and the literature, Sturgis and Murray (1997) suggested that there may be a more
complex relationship between nutrient enrichment and SAV growth and survival. For example, the
relationship may depend on the form, delivery frequency, and loading rate of nutrients.
There now appears to be enough scientific data and knowledge to establish nutrient regimes that will
protect temperate and subtropical seagrass ecosystems.
Other Examples of Important Biotic Effects of Nutrient Overenrichment
It is difficult to find recent quantitative relationships between nutrient loading and fishery impacts for
coastal systems. One explanation is that the large marine vertebrate species which are mostly extinct or
severely over-fished help determine the nutrient assimilative capacity of marine ecosystems including
estuaries and coastal waters (Jackson et al. 2001). For economically important fisheries, variable fishing
pressure may cloud the analysis and other factors may vary to obscure nutrient-related patterns. Often,
one is left with mostly anecdotal insights as to potential negative effects of Overenrichment on higher
trophic levels focusing on data and insights only from recent decades. There is a plausible and positive
relationship between marine fisheries yield and nitrogen supply, with a wide range in estuarine and
coastal marine habitats represented (Nixon 1992). This approximately natural response is analogous to
what mariculturists attempt to achieve when they fertilize fish enclosures, but these enclosures, whether
on land or in the marine environment, are known to cause local water quality problems. The relationship
Nixon reported on involved a two-step function: a positive relationship between primary production (g C
m-2 y-1) and DIN input (moles m-2 y-1) and between fisheries yield (kg ha-1 y-1) and primary production
(Figure 2-16a-c). In contrast to the foregoing positive relationship, a pelagic-demersal ratio from fishery
landings from 14 study areas in European coastal waters appeared to be a proxy for the differential
impact of nutrients on pelagic and benthic systems mediated by nutrient enrichment, resulting in hypoxia
(de Leiva Moreno et al. 2000). A general model suggests that Overenrichment can lead to decreased
fisheries productivity (Figure 2-17).
Oysters are ecosystem engineers that create biogenic reef habitat important to estuarine biodiversity,
benthic-pelagic coupling, and fishery production (Lenihan and Peterson 1998). These authors conducted
an analysis of habitat degradation (i.e., oyster dredging) through fishery disturbance that enhanced
impacts of hypoxia on oyster (Crassostrea virginica) reefs in North Carolina. This is a fairly
complicated story but the conclusions from the analysis seem inescapable. Dredging lowered the oyster
reef into the hypoxic zone where the reef and associated organisms died from DO depletion. Another
example of effects of nutrient Overenrichment causing impacts on oysters was reported by Ryther (1954)
for Long Island, New York duck farms where nutrient enrichment caused phytoplankton to grow that
were indigestible for oysters.
Hypoxia is known to kill other benthic organisms. Diaz and Rosenberg (1995) cited many studies where
hypoxia resulted in the deaths of benthic communities. A related cause with hypoxia is that polychaetes
may extend themselves out of their sediment burrows and become easier prey to fish predators. Another
Nutrient Criteria—Estuarine and Coastal Waters 2-29
-------
1000
z
g
o
Q ' >.
O^ 100
n E
£ »
D_ O)
Q_
•in
— IMA
"U; ^3
......... 15a l.5b A
n?a «14a
- 1 2b .14b
— LJ W
- 2a 2 i
;
i i i linil i i i I mil i i i I mil i i 1 1 ii ill i i i Imil
0.1
10
102
103
104
DIN INPUT mmol my
Figure 2-16a. Primary production by phytoplankton (14C uptake) as a function of the estimated annual
input of dissolved inorganic nitrogen per unit volume of a wide range of marine ecosystems. Source:
Nixon (1992).
O
O^
Q '>.
Q_ ^
>: O
E
Q_
1000
100
10
I ....I
I....I
0.01
0.1
1
10
100
DIN INPUT moles m "2 y "1
Figure 2-16b. Primary production by phytoplankton (14C uptake) as a function of the annual
input of dissolved inorganic nitrogen per unit area of a wide range of marine ecosystems.
Source: Nixon (1992).
2-30
Nutrient Criteria—Estuarine and Coastal Waters
-------
1000
03
D)
H.
d
_i
LJJ
>-
CO
LJJ
01
LJJ
CO
100
10
• ESTUARIES
» SHELF
o UPWELLING
A OTHER
13
A
20
A—
10
20
-A
I I I I I I I I I
10
1.0
CM
o
D)
0.1
100
1000
PRIMARY PRODUCTION g C m"2 y"1
Figure 2-16c. Fisheries yield per unit area as a function of primary production in a wide range of estuarine
and marine systems. Modifed from Nixon (1988) to include a revised primary production estimate for the
Peru Upwelling from Guillen and Calienes (1981). Systems identified and data sources in Nixon (1982)
and Nixon etal. (1986). Source: Nixon (1992).
Nutrient Criteria—Estuarine and Coastal Waters
2-31
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Although higher nutrient concentrations initially increase the productivity of fisheries, ecological systems
worldwide show negative effects as nutrient loading increases and hypoxic or anoxic conditions develop. Each
generic curve in the lower half of the figure represents the reaction of a species guild to increasing nutrient
supplies. The top half of the figure illustrates trends in various marine systems around the world. Reversals show
that trends toward overenrichment have been turned around in several areas.
Oligotrophic
(low productivity)
rophic Eutrophic Dystrophic
m productivity) (high productivity) (no productivity)
SE North Sea
N. Gulf of Mexico
N. Adriatic Sea
Great Lakes
Kattegat
Baltic Sea
Seto Inland Sea
Yellow Sea
NW Black Sea
Chesapeake Bay
Seasonal—*-4 Permanent bottom anoxia
Figure 2-17. Comparative evaluation of fishery response to nutrients. Although higher
nutrient concentrations initially increase the productivity of fisheries, ecological systems
worldwide show negative effects as nutrient loading increases and hypoxic or anoxic
conditions develop. Each generic curve in the lower half of the figure represents the
reaction of a species guild to increasing nutrient supplies. The top half of the figure
illustrates trends in various marine systems around the world. Reversals show that trends
toward overenrichment have been turned around in several areas. Source: CENR 2000.
effect of hypoxia on the biota is the loss of sufficient bottom habitat. This is often difficult to
quantitatively relate to economically important species but the negative effect may still be real. If
endangered species are present, this hypoxic effect is one of direct societal and legal concern.
2.6 CONCLUDING STATEMENT ON NITROGEN AND PHOSPHORUS CONTROLS
It is important to note that in estuaries and nearshore coastal marine waters, the fact that nitrogen often
limits algal biomass production does not mean that managers should be unconcerned about phosphorus
enrichment. In river-dominated temperate estuaries, the upper reaches of estuaries, such as lakes and
rivers, are often phosphorus limited. The manager who therefore concentrates on phosphorus
management alone risks letting an undue amount of nitrogen proceed downstream to exacerbate problems
2-32
Nutrient Criteria—Estuarine and Coastal Waters
-------
where an abundance of P allows the excess N to drive trophic conditions to unacceptable levels of
nutrient enrichment.
Similarly, any reductions achieved in P loadings and concentrations at the coastal margin will limit
potential eutrophy/hypertrophy even in the face of abundant nitrogen. Consequently, the prudent
management strategy is to limit both phosphorus and nitrogen. Emphasis on one or the other as an
element of symptomatic management in fresh or saline waters may be appropriate in some cases, but the
manager must always be concerned about the downstream consequences and the net enrichment effects
to the larger system.
In summary, attempting to understand the nutrient overenrichment problem in estuaries and coastal
ecosystems primarily from a bottom-up perspective provides a limited perspective. This manual has
included references to the historical past that reported on potential positive effects of top-down controls
on nutrient overenrichment. It is likely that the most scientifically robust nutrient criteria will need to
take into account the effects of past overfishing and its consequences for marine eutrophication (Jackson
et al. 2001). Thus, higher trophic levels are more than just a thermodynamic response to nutrient
enrichment because they help modulate many of the negative consequences of overenrichment.
Ecological feedback mechanisms that involve higher trophic levels can be a positive tool in nutrient
management.
Nutrient Criteria—Estuarine and Coastal Waters 2-33
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CHAPTER 3
Classification of Estuarine and
Coastal Waters
Major Factors Influencing Estuarine Susceptibility
to Nutrient Overenrichment
Examples of Coastal Classification
Coastal Waters Seaward of Estuaries
3.1 INTRODUCTION
Purpose and Background
Classification is an important step in addressing the problem of degradation, especially because of
nutrient overenrichment. There are too many nutrient-degraded estuaries in the United States for the
Nation to conduct comprehensive ecosystem studies of all those affected by overenrichment. Where
possible, similar estuaries, tributaries, or coastal reaches should be equated through physical
classification to reduce the magnitude of the criteria development problem and to enhance predictability
of management responses. To be useful, classification should reduce variability of ecosystem-related
measures (e.g., water quality factors) within identified classes and maximize interclass variability. This
is important because managers need to understand how different types of estuaries and coastal waters, as
well as important habitat differences within these systems, respond to nutrient overenrichment in order to
plan effective management strategies.
The ecosystem processes that regulate nutrient dynamics, discussed in Chapter 2, should provide the
elements for initial development of a useful classification system. Although predicting susceptibility of
estuarine and coastal waters to nutrient overenrichment is in a primitive state, several approaches are
reviewed because they have some utility even if they are only marginally adequate for prediction of
nutrient effects. The general approach is also appropriate for coastal systems.
General trends relate N loading with chlorophyll and primary productivity; however, these trends are
seldom usefully predictive for individual systems or for all classes of coastal systems (Kelly in press).
Progress has been especially slow in predicting many of the secondary, but societally important, effects
of nutrient overenrichment, e.g., bottom water dissolved oxygen (DO) deficiency, harmful algal blooms
(HABs) or species-specific HABs, formation of macroalgal mats, fisheries productivity, and species
composition. For many cross-system comparisons, N loading and SAV decline have become more
predictive than for other indirect effects (Duarte 1995; Dennison et al. 1993), but even here the
predictions may be confounded by highly variable ecosystem factors. A major impedance to effective
understanding is limited comparative studies designed to test hypotheses regarding estuarine
susceptibility to nutrient enrichment (Turner 2001). Post hoc comparative approaches and assessment of
disparate studies have been useful but clearly inadequate (Livingston 200 Ib).
Post hoc statistical approaches have helped explain some of the variability in eutrophication, but have not
captured the actual mechanisms and their interactions controlling eutrophication across estuaries and
Nutrient Criteria—Estuarine and Coastal Waters
3-1
-------
coastal waters (NRC 2000). The ability to explain the mechanisms in a predictive manner is clearly a
critical national and global research need, as nutrient overenrichment of coastal ecosystems extends far
beyond the shores of the United States. For site-specific criteria, several approaches are available,
including empirical regression and mechanistic simulation models. The large effort typically required to
calibrate and verify mechanistic models is an indicator of the difficulty in understanding the many
potential confounding factors of ecosystem-level prediction. A basic premise of this manual is that
knowledge of the physical setting and the minimally disturbed ecosystem reference condition must
underpin monitoring and management efforts to protect and restore coastal systems impaired by nutrient
overenrichment.
Only in approximately the past two decades have comparative studies of nutrient dynamics among two or
more relatively large estuaries been published (e.g., Fisher et al. 1988; Malone et al. 1999; Pennock et al.
1994). Comparative analysis of the Delaware and Chesapeake Bays has provided insights regarding
processes that control expression of nutrient enrichment (Chapter 2). For example, both systems are
drowned river mouth coastal plain estuaries and are located adjacent to each other along the coast, but
have very different responses to nutrient loading. Delaware Bay has a somewhat larger nutrient load than
the Chesapeake, but has few of the nutrient enrichment symptoms well chronicled for the Chesapeake
Bay (Flemer et al. 1983, Sharp et al. 1994, Chesapeake Bay Program Periodic Status Reports). Similar
insights have been provided by comparing nutrient processes between Delaware and Mobile Bays
(Pennock et al. 1994). Susceptibility appears to be largely explained by differences in the physics of
flushing, including bathymetry and related physical habitat differences.
Defining the Resource of Concern
As a first step in classification, defining the resource of concern is important. Resources of concern are
estuaries and coastal waters located in the contiguous States or within authorized Tribal lands. Managers
must decide which waterbodies to include in the population to which criteria will be applicable. A lake
classification may exempt small ponds that might be excluded because of their size and man-made
nature, whereas tidal creeks, although small, still have a functional connection to the larger estuary and
might not be excluded because of size. Many estuaries and coastal waters share multiple political
boundaries, and for the sake of consistency all involved jurisdictions should jointly decide on the scale of
inclusiveness. For open coastal waters, a State or authorized Tribe's legal authority may extend for a
relatively short distance on the continental shelf, e.g., 3 nautical miles. However, coastal oceanographic
processes seaward of the statutory limit likely influence nutrient overenrichment processes and
exacerbate the difficulty of diagnosing the anthropogenic contribution to nutrient problems.
3.2 MAJOR FACTORS INFLUENCING ESTUARINE SUSCEPTIBILITY TO
NUTRIENT OVERENRICHMENT
The NRC (2000) publication summarized approximately a dozen factors deemed important to
characterize the susceptibility of estuaries to nutrient loading. A short list is provided; however, it is
expected that the following list will be modified and refined as more is learned about the subject:
3-2 Nutrient Criteria—Estuarine and Coastal Waters
-------
1. System dilution and water residence time or flushing rate
2. Ratio of nutrient load per unit area of estuary
3. Vertical mixing and stratification
4. Algal biomass (e.g., chlorophyll a, and chlorophyll a corrected for nonchlorophyll a light
attenuation over seagrass/SAV beds and macroalgal biomass as AFDW)
5. Wave exposure (especially relevant to seagrass potential habitat)
6. Depth distribution (bathymetry and hypsographic profiles)
7. Ratio of side embayment (s) volume to open estuary volume or other measures of embayment
influence on flushing.
Several terms listed above are briefly discussed because their significance often is not adequately
appreciated.
Dilution
The volume of an estuary affects its ability to dilute inflowing nutrients. Thus, the loading rate of
nutrient per unit volume of the estuary is a better indicator of the potential for exceeding the assimilative
capacity of the estuary as a whole than is the absolute loading rate. This ratio may not express the
potential for local effects near the point of entry into the estuary, as nutrients there are diluted by only a
fraction of the total estuary volume. The potential for such local effects is reduced if mixing into the
main body of the estuary is rapid.
Water Residence Time
Estuaries that flush rapidly (i.e., have a short residence time) will export nutrients more rapidly than
those that flush more slowly, resulting in lower nutrient concentrations in the estuary. Dettmann (in
press) has derived a theoretical relationship between the mean residence time of freshwater in an estuary
and the increase in the average annual concentration of total nitrogen in the estuary as a result of inputs
from the watershed and atmosphere. In addition, estuaries with residence times shorter than the doubling
time of algal cells will inhibit formation of algal blooms. Residence time or flushing rate is discussed in
more detail in Appendix C.
Stratification
Highly stratified systems are more prone to hypoxiathan are vertically mixed systems. Stratification not
only limits downward transport of oxygen from atmospheric reaeration, it also retains nutrients in the
photic zone, making them more available to phytoplankton. In stratified systems, it may be more
appropriate to estimate the dilution potential of the estuary using the volume above the pycnocline rather
than the entire volume of the estuary.
It is expected that the shortened list will be revised and modified as more is learned about factors
important in estuarine and coastal waters classification. Some of these factors will apply to estuaries and
others to coastal waters.
Nutrient Criteria—Estuarine and Coastal Waters 3-3
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3.3 EXAMPLES OF COASTAL CLASSIFICATION
Scientists and resource managers have used various classification schemes for many years to organize
information about ecological systems. As discussed earlier, estuaries and coastal water systems are
characterized by a suite of factors (e.g., river flow, tidal range, basin morphology, circulation, and
biological productivity) that are ultimately controlled largely by geology and climate. A review of
Chapter 6 in the NRC (2000) publication provides useful descriptions of the various approaches to
estuarine classification, and some pertinent features are highlighted below.
Geomorphic Classification
Geomorphic classification schemes provide some insight into the circulation structure and are a first-
order estimate of water residence time or flushing characteristics. Such classifications may not in
themselves be predictive of susceptibility to nutrient enrichment, but they are a useful place to begin a
first-order assessment of susceptibility. Knowledge of deep channels, however, identifies potential areas
subject to hypoxia, and the extent of shallow waters and associated factors (e.g., wind fetch) often
provides insights into potential seagrass habitat.
Estuaries can be divided geomorphically into four main groups (Pritchard 1955, 1967; Dyer 1973): (1)
coastal plain estuaries, (2) lagoonal or bar-built estuaries, (3) fjords, and (4) tectonically caused estuaries.
This classification frequently appears in textbooks, and only some important features relative to nutrient
susceptibility are described.
Coastal Plain Estuaries: Classical and Salt Marsh
Both subclasses are characterized by well-developed longitudinal salinity gradients that influence
development of biological communities. Examples of the classical type include the Chesapeake Bay (the
largest estuary of this type), Delaware Bay, and Charleston Harbor, SC. Vertically stratified systems
with relatively long residence times (e.g., Chesapeake Bay) tend to be susceptible to hypoxia formation.
Pritchard (1955) further classified drowned river valley estuaries into four types (A-D) depending on the
advection-diffusion equation for salt (Table 3-1). Type C estuaries are less sensitive to algal bloom
formation and hypoxia because of mixing features.
The salt marsh estuary lacks a major river source and is characterized by a well-defined tidal drainage
network, dendritically intersecting the extensive coastal salt marshes (Day et al. 1989). Exchange with
the ocean occurs through narrow tidal inlets, which are subject to closure and migration following major
storms (e.g., Outer Banks, NC). Consequently, salt marsh estuarine circulation is dominated by
freshwater inflow, especially groundwater, and tides. The drainage channels, which seldom exceed a
depth of 10 m, usually constitute less than 20% of the estuary, with the majority consisting of subaerial
and intertidal salt marsh. These systems are a common feature of the Atlantic coast, particularly between
Cape Fear, NC, and Cape Canaveral, FL. Mangrove estuaries occur from around Cape Canaveral south
on Florida's east coast and on Florida's west coast from around Tarpon Springs south. Nutrient
dynamics, primary production, and system respiration that occur within emergent marshes may greatly
affect water quality in the estuarine channels (Cai et al. 1999).
3-4 Nutrient Criteria—Estuarine and Coastal Waters
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Table 3-1. General drowned river valley estuarine characteristics
Estuarine
type3
A
B
C
D
Dominant
mixing force
River flow
River flow, tide
Tide, wind
Tide, wind
Mixing
energy
Low
Moderate
High
V. high
Width/depth
ratio
Low
Moderate
High
V. High
Salinity
gradient
Longitudinal
vertical
Longitudinal
vertical
Longitudinal
lateral
Longitudinal
Mixin
g
indexb
>1
<1/10
<1/20
7
Turbidity
V. high
Moderate
High
High
Bottom
stability
Poor
Good
Fair
Poor
Biological
productivity
Low
V. high
High
Moderate
Example
Southwest Pass
Mississippi River
Chesapeake Bay
Delaware Bay
?
CD
r-c
O
^
r-h"
CD
m
Tollows Pritchard's advection-diffusion classification scheme.
bFollows Schubel's definition: MI = equation here- (vol. freshwater discharge on 1A tidal period) / (vol. tidal prism).
Source: Neilson and Cronin, 1981.
CD
Q>
D
Q.
O
O
CD
00
Ul
-------
Lagoons
Lagoons are characterized by narrow tidal inlets and are uniformly shallow (i.e., less than 2 m deep)
open-water areas. The shallow nature enhances sediment-water nutrient cycling. Flushing is typically
of long duration. Most lagoonal estuaries are primarily wind-dominated and have a subaqueous drainage
channel network that is not as well drained as the salt marsh estuary. Lagoons fringe the coast of the
Gulf of Mexico and include the mid-Atlantic back bays; Pamlico Sound, NC; and Indian River Lagoon,
FL. Although these systems are typically shallow, they may have pockets of hypoxic water subject to
spatial variability because of freshwater pulsing and wind effects. Some lagoonal systems have relatively
strong vertical stratifications near the freshwater river mouth and may be subject to hypoxia formation
(e.g., Perdido Bay, AL/FL; Livingston 200la).
Fjords and Fjordlike Estuaries
Classical fjords typically are several hundred meters deep and have a sill at their mouth that greatly
impedes flushing. Hypoxia/anoxia is often a natural feature but anthropogenic nutrient loading can
severely exacerbate the problem. Examples of classical fjords on the North American continent can be
found in Alaska and Washington State (Puget Sound). Some other estuaries were also formed by glacial
scouring of the coast, but in regions with less spectacular continental relief and more extensive
continental shelves. Examples of these much shallower, fjordlike estuaries can be found along the Maine
coast.
Tectonically Caused Estuaries
Tectonically caused estuaries were created by faulting, graben formation (i.e., bottom block-faults
downward), landslide, or volcanic eruption. They are highly variable and may resemble coastal plain
estuaries, lagoons, or fjords. San Francisco Bay is the most studied estuary of this type (Cloern 1996).
Man-Made Estuaries
Especially around the Gulf of Mexico, dredged bayous, canals, and salt water impoundments with weirs
function as estuaries but do not fit well any of the other types presented. As a special case, especially in
the Gulf of Mexico, the passes of some estuaries periodically were closed off by storms and historically
remained closed until a natural event reopened them (e.g., Perdido Bay, AL/FL; R. Livingston, personal
communication). In recent years, these systems typically are maintained in an open condition by
dredging. Dredged inlets such as at Ocean City, MD, also fit this classification.
Physical/Hydrodynamic Factor-Based Classifications
Classification Using Stratification, Mixing, and Circulation Parameters
Estuarine circulation was a dominant consideration used in an earlier classification of the Chesapeake
Bay, a coastal plain system, and major tributaries, and is largely utilized today with some modifications
(Flemer et al. 1983). Coastal plain estuaries are sometimes classified by mixing type: highly stratified,
partially mixed (moderately stratified), or well mixed (vertically homogeneous). The flow ratio of these
estuaries (the ratio of the volume of freshwater entering the estuary during a tidal cycle to its tidal prism)
is a useful index of the mixing type. If this ratio is approximately 1.0 or greater, the estuary is normally
3-6 Nutrient Criteria—Estuarine and Coastal Waters
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highly stratified; for values near 0.25 the estuary is normally partially mixed; and for ratios substantially
less than 0.1, it is normally well mixed (Biggs and Cronin 1981).
Stratification/Circulation Parameters
Hansen and Rattray (1966) developed a two-parameter classification scheme based on circulation and
stratification of estuaries. Circulation is described by the nondimensional parameter U/Uf, where Us is
the net (time-averaged) longitudinal surface current and Uf is the cross-sectional average longitudinal
velocity. Stratification is represented by the nondimensional parameter 5S/S0, where 5S is the top-to-
bottom difference in salinity and S0 is the mean salinity. Jay et al. (2000) review alternative two-
parameter classification schemes involving parameters such as the ratio of tidal amplitude to mean depth,
along-estuary and vertical density differences and vertical tidal excursion of isopycnals, or other factors
that take into account effects of tidal flats and provide additional discussion to which the reader is
referred for additional insights. They argue that the merit of the approach is its simplicity of parameters
employed and the predictive ability with regard to salt transport needed to maintain salt balance in
modeling.
Classification Using Water Residence Time
Water residence time, the average length of time that a parcel of water remains in an estuary, influences a
wide range of biological responses to nutrient loading. The residence time of water directly affects the
residence time of nutrients in estuaries, and therefore the nutrient concentration for a given loading rate,
the amount of nutrient that is lost to internal processes (e.g., burial in sediments and denitrification), and
the amount exported to downstream receiving waters (Dettmann in press, Nixon et al. 1996). Residence
times shorter than the doubling time of algae will inhibit bloom formation because algal blooms are
exported from the system before growing to significant numbers. Residence time can also influence the
degree of recruitment of species reproducing within the estuary (Jay et al. 2000).
There are a number of definitions of water residence time, including freshwater residence time and
estuarine residence time (Hagy et al. 2000; Miller and McPherson 1991), each with its own interpretation
and utility. Freshwater residence time is the mean amount of time required for freshwater entering the
estuary to exit the seaward boundary, whereas estuarine residence time is the average residence time in
the estuary for all water, regardless of its origins. Because nutrient loading is generally associated with
freshwater inputs, freshwater residence time is generally the most useful measure in considering estuary
sensitivity to nutrient loading. Freshwater residence time of a given estuary is influenced by numerous
factors, including freshwater loading rate (Pilson 1985; Asselin and Spaulding 1993; Hagy et al. 2000),
tidal range, and wind forcing (Geyer 1997), and therefore varies over a range of time scales.
Residence time and volume together may be used to scale nitrogen loading to estuaries to permit
calculation of nitrogen concentrations and perform cross-system comparisons. Dettmann (in press) uses
a model that includes mechanistic representations of nitrogen export and loss within estuaries to show
that [Nu ], the contribution to the annual average concentration of total nitrogen in an estuary from upland
sources (watershed, direct discharges, and atmosphere), may be calculated as
Nutrient Criteria—Estuarine and Coastal Waters 3-7
-------
where I/is the annual average loading rate (mass/month) of total nitrogen from all upland sources
(watershed and atmosphere), Jfw is the freshwater residence time in months, Vis the estuary volume, and
" is a parameter (value = 0.3 month"1) related to losses of nitrogen to processes such as denitrification and
burial in sediments within the estuary.
Definitions of residence time, the methods used to measure or calculate them, variability of residence
time, and other estimators of residence time are described further in Appendix C.
River Flow, Tides, and Waves
Dronkers (1988) proposed an estuarine classification that distinguished various types of estuarine
ecosystems based on water exchange processes (e.g., river flow, tides, and waves) that greatly affect
energy and material fluxes including mixing (Table 3-2). This classification suggests that river flow in
partially mixed estuaries is essentially neutral, but its variation relative to hydrodynamic residence time
can be important in interpreting property-salinity diagrams (Cifuentes et al. 1990) (Figure 3-1). River
flow in the partially mixed mainstem of the Chesapeake Bay is seasonally important.
Tidal Amplitude—A Dominant Physical Factor
Tidal amplitude provides a means to broadly classify estuaries relative to their sensitivity to nutrient
supplies. Monbet (1992) analyzed phytoplankton biomass in 40 estuaries and concluded that macrotidal
estuaries (mean tidal range >2 m) generally exhibit a tolerance to nitrogen pollution despite high loadings
originating from freshwater outflows (Figures 3-2a, b). These systems generally exhibit lower
concentrations of chlorophyll a than do systems with lower tidal energy, even when they have
comparable concentrations of nitrogen compounds. Estuaries with mean annual tidal ranges <2 m seem
more sensitive to dissolved nitrogen, although some overlap occurs with macrotidal estuaries.
NOAA Scheme for Determining Estuarine Susceptibility
NOAA (Bricker et al. 1999) developed a categorical approach based on surveys and decision rules that
led to a classification of estuarine nutrient export potential (e.g., dilution potential and flushing
potential). From this information a susceptibility matrix was constructed. The low, moderate, and high
susceptibility indices were combined with low, moderate, and high human levels of nutrient input,
resulting in a final matrix of overall human influence (see Appendix D for details).
Comparative Systems Empirical Modeling Approach
The empirical regression method can be used to determine the response of estuarine systems to nutrient
loading. This approach requires that the response factor be common to all systems in the analysis and
assumes that any graded response among systems is due to a common form of disturbance, e.g., nutrients.
The space-for-time paradigm (Pickett 1988) posits that relationships between nutrient inputs and
3-8 Nutrient Criteria—Estuarine and Coastal Waters
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Table 3-2. Classification of coastal systems based on relative importance of river flow, tides, and
waves to mixing
Type
River flow
Tide
Waves
Description
I
II
III
IV 0
V
VI
VII
River delta
River delta (plus barriers)
Tidal river delta
Coastal plain estuary
Tidal lagoon
Bay
Coastal lagoon
Plus and minus designations indicate relative impacts; e.g., - means that river discharge is very small relative to tidal
and wave energy.
Source: Adapted from Dronkers 1988.
ecologically meaningful estuarine responses, using multiple systems, have predictive capability, at least
for the systems used in the model development. This allows for a wide range in nutrient loading and
estuarine types to be included. The comparative-systems empirical approach has been used to determine,
for example, relationships between nutrient inputs and fish yields (Lee and Jones 1981, Nixon 1992),
benthic biomass, production and abundances (Josefson and Rasmussen 2000), summer ammonia flux
(Boynton et al. 1995), chlorophyll a concentration (Boynton et al. 1996, Boynton and Kemp 2000,
Monbet 1992), primary productivity (Nixon et al. 1996), and the dominant source of primary productivity
(Nixon et al. in press). In many of these cases, important environmental factors such as flushing time and
depth are used to normalize the nutrient loading in a similar way as Vollenweider (Vollenweider 1976)
Did for lakes to yield more precise relationships. Appendix E provides additional details.
Other Considerations
Habitat Type
The presence and extent of different habitat/community types may help distinguish one or more estuaries
within a region. These types may include seagrasses, mangroves, mudflats, deep channels, oyster reefs,
dominance of sand versus mud bottoms, extensive emergent marshes (typically coastal plain systems),
and the presence of unconsolidated versus rocky shorelines. Some of these categories may be
subclassified by salinity ranges (e.g., oligohaline, mesohaline, and polyhaline). Although related more to
water quality, blackwater versus turbid versus relatively clear estuaries defines a group representative of
estuaries around the Gulf of Mexico.
Theoretical Considerations
Coastal zone managers may wish to consider more theoretical approaches to classification as ecosystem
science develops a more in-depth understanding of ecosystem processes for estuaries under their
purview. Several different approaches are described in Appendix F.
Nutrient Criteria—Estuarine and Coastal Waters
3-9
-------
DC
LU
O
o
O
I-
LLJ
DC
O
GC
O
1 RIVER '
~
A
CONSERVATIVE
OCEAN
OCEAN -
SALINITY
RIVER
\
D
PRONOUNCED SINK
OCEAN -
Figure 3-1. Idealized micronutrient-salinity relations showing concentration and mixing
of nutrient-rich river water with nutrient-poor seawater. Source: Peterson et al. 1975. A.
Expected concentration-salinity distribution of a substance behaving in a conservative
manner (e.g., chloride) in an estuary. B. Expected concentration-salinity distribution of a
substance for which the estuary is a source (e.g., paniculate carbon). C. Expected
concentration-salinity distribution of a substance for which the estuary is a sink (e.g.,
phosphorus). D. Expected concentration-salinity distribution of a substance for which
the estuary is a pronounced sink, that is, where the concentration of the substance in the
estuary is lower than the river and the ocean (e.g., Si). Source: Biggs and Cronin 1981.
3-10
Nutrient Criteria—Estuarine and Coastal Waters
-------
Mean tidal range
20-
I1
o
u
10
0
<2m
>2m
0
10 100
D.I.N. loadings g nfy"1
1000
Figure 3-2a. Relationship between the mean annual loadings of dissolved inorganic nitrogen
(DIN) and the mean annual concentration of chlorophyll a in microtidal and macrotidal estuaries.
Summary
Various ways are available to classify estuaries regarding their vulnerability to nutrient enrichment.
None appear to provide all the information a resource manager may want for decisionmaking. The
NOAA estuarine export potential (EXP) appears to have the current greatest utility for predictive
purposes for large systems, but even this approach embodies considerable variability (e.g., see Figures 6-
5 and 6-6 in NRC 2000). For embayments within a larger estuary, the comparative empirical modeling
approach has been demonstrated to have considerable utility. The more theoretical models eventually
may provide greater predictive power, especially as to biological sensitivities to nutrient enrichment.
They are data intensive and may become more useful at a future time.
3.4 COASTAL WATERS SEAWARD OF ESTUARIES
Several approaches are available to classify coastal waters. The geomorphic focus is a good place to
begin, hydrographic considerations should follow, and finally habitat and community features should be
considered. Although functional considerations and theoretical indices are not described for coastal
waters, they have as much relevance for these waters as they do for estuaries. Even though much of the
concern for coastal waters will be within 20 nautical miles of shore, and most of that within the 3-mile
limit, elements of the following large-scale classification scheme will have value to the manager and
investigator.
Nutrient Criteria—Estuarine and Coastal Waters
3-11
-------
Mean tidal range + >22™
37
37+34 32++++36
32 32+ 33 33
+33
100
1000
D.I.N. mM
Figure 3-2b. Relationship between the mean annual concentrations of dissolved inorganic nitrogen (DIN)
and chlorophyll a in microtidal and macrotidal estuaries. Source: Monbet 1992.
Geomorphic Classification
The flow of energy and nutrients through coastal food webs differs greatly among continental shelves,
and is driven largely by differences in the form and amount of primary production (e.g., seagrasses are
important in the Big Bend area of Florida and kelp forests are important habitats along much of the U.S.
Pacific Coast and sections of coastal Maine). These differences in turn ultimately are determined by
differences in local and ocean-scale patterns of climate (e.g., light and temperature effects), water
circulation, chemistry, and shelf geomorphology (Alongi 1998). The spring bloom, especially along the
U.S. Atlantic Coast, generally progresses from low to higher latitudes but with sharper seasonal peaks
toward higher latitudes. Variability in the progression should be considered in any classification scheme.
Because near-coastal shelf oceanographic processes usually are not limited by the jurisdiction of a single
State, it is important that a similar classification approach be shared among coastal States, where that
oceanography determines the sensitivity of the ecosystem to nutrient enrichment. The geographic extent
of the shelf in which a State has jurisdiction is a useful place to begin classification. Here one should
consider whether the shelf is wide or narrow (e.g., mid-Atlantic versus Pacific Coast). The Texas coastal
shelf is very wide, with a gentle slope compared with much of the northern Gulf of Mexico. The
steepness of the slope is another useful factor, as it may influence bottom sediment stability and
3-12
Nutrient Criteria—Estuarine and Coastal Waters
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upwelling. The degree of bottom roughness or sculpture may influence vertical mixing, which may in
turn influence water column stability and depth of the euphotic zone versus mixing depth.
Nongeomorphic Classification
Walsh characterized the world's continental shelves on the basis of their location, major rivers, and rates
of primary production, and included some U.S. coastal waters for comparison. The shelf proper is where
oceanic and estuarine boundaries often intermingle. At the shelf edge, cold, nutrient-rich water and
associated materials intrude onto the shelf. There, exchanges often are rapid, promoting conditions
favorable for higher fertility than in the open ocean. Higher primary productivity is the main reason why
approximately 90% of the world's fish catch is harvested on the continental shelves versus the open sea
(Alongi 1998).
Water Quality Trend Detection on the Shelf
Physical mixing and advective processes may add considerable variability to water column measures.
Therefore, it is important to consider detection of trends in nutrient concentrations and measures such as
chlorophyll a based on comparisons at a reference salinity (e.g., 30 psu). Otherwise, classification
schemes may incur extraneous variability. A common approach is to use "mixing diagrams" to compare
measured changes in an ambient constituent among sampling periods. At mid- and higher latitudes,
winter measures of DIN and DIP may provide insight into long-term trends of changes in nutrient
concentrations available to drive the spring bloom. At low latitudes winter values will likely have less
applicability, as primary production has a smaller seasonal signal.
Presence of Large Rivers
Although large rivers are included in Walsh's characterization of shelf systems, it seems useful to
distinguish shelf areas based primarily on large rivers, such as the Mississippi River in the Gulf of
Mexico and the Columbia River off the Washington-Oregon coast. Large rivers on the shelf dominate
local ecological relationships.
Hydrographic Features
Vertical salinity differences tend to decrease toward the open ocean boundary. The principal reason is
summertime thermal stratification. The thermocline tends to be deeper toward the open sea margin,
except where buoyancy effects are associated with large rivers that flow onto the shelf. Coastal waters
contain a variety of biotic communities, including a diverse assemblage of macroepifauna and -infauna,
kelp forests, coral reefs, bottom and pelagic fishes, marine mammals, and seabirds. The relationship of
these communities to physical-ordering factors can assist in classification.
Temperate and subtropical coastal waters also experience a seasonal sea-level fluctuation, whereby
summer levels rise approximately 0.2 m by upper-ocean heat expansion, producing what is known as
thermosteric effects (Pattulo et al. 1955, Bell and Goring 1998). This nontidal process operates in
conjunction with other factors affecting apparent mean sea level (e.g., near and far-field wind effects and
barometric pressure). Depending on local conditions, water levels overlying the continental shelf and in
estuaries can rise from 0.1 to 0.2 m. Such a rise may seem nominal but can have a significant impact in
Nutrient Criteria—Estuarine and Coastal Waters 3-13
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wetlands and other low-lying areas that potentially exchange nutrients and suspended sediments with the
coastal ocean.
Coastal ocean waters range from quite cold (e.g., Gulf of Maine) to quite warm (e.g., Gulf of Mexico).
Where large rivers enter coastal waters, such as the Mississippi River Plume and Chesapeake Bay Plume,
visual discoloration can be observed because suspended material from land runoff and relatively high
plankton concentrations contrast with the predominantly blue color of the open ocean. The Columbia
River and the Mississippi River form an "estuary" mostly at sea, as very little of the diluted seawater is
bounded by land.
Physical gradients are dynamic and change at multiple scales. Seasonal or wet and dry periods
frequently differ depending on the various shelf gradients associated with estuarine, riverine, and
ocean/shelf break processes. Regional geomorphology and physical mixing processes play a pivotal role
in energy flow and material cycles. For example, the Loop Current in the eastern Gulf of Mexico may
show seasonal reversals and vary seasonally in its penetration onto the shelf. Along-shore drift inside the
north-flowing Gulf Stream off the Mid-Atlantic Bight tends to transport materials southward toward the
North Carolina coast. Further south, the Gulf Stream forms a seaward boundary that tends to
significantly isolate in-shore waters from those beyond the shelf break. Local current maps are available
from the National Ocean Service of NOAA (www.noaa.gov; then click on nos).
Many different types of boundaries or fronts occur in coastal seas, but no formal classification exists.
Alongi (1998) lists five categories:
• Shelf-sea (tidal) fronts
Estuarine fronts or plumes
• Shelf-break fronts
• Upwelling fronts
Island wakes and fronts caused by other land features
Fronts provide increased physical stability at local scales, which may positively influence primary
production and energy flow to higher trophic levels (see Chapter 2).
Habitat/Community Differences
Presence of Mangrove/Seagrass and Coral Communities
Along the southeastern Florida Atlantic coast exists a combination of mangrove, seagrass, and coral reef
ecosystems. In some localities, each community type may dominate the others, but often they co-occur.
Seagrass communities may dominate certain shelf areas along the west coast of Florida (e.g., Big Bend
region). The Flower Gardens, a disjunct coral community, exist off the southern coast of Texas. Alongi
(1998) devotes chapters to coral reefs and mangrove ecosystems including factors regulating primary
productivity (e.g., N and P). The role of N and P enrichment versus grazing in coral reef ecosystems is
still strongly debated in the scientific literature (e.g., Miller et al. 1999). A paper by Chen and Twilley
(1999) discusses soil nutrient relationships and productivity in a Florida Everglades mangrove ecosystem
along an estuarine gradient (see the references cited above for the most recent perspective). These
3-14 Nutrient Criteria—Estuarine and Coastal Waters
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distinctive ecosystems provide a basis for local coastal waters classification. Mangrove communities
also occur along the lower Texas coast, and seagrasses are a dominant community in Laguna Madre, TX.
Presence of Seaweed
Seaweeds are common algal communities in rocky intertidal zones (e.g., Fucusspp.), attaching
themselves by means of a holdfast. Seaweeds belong to three marine algal classes: Chlorophyceae
(green algae), Rhodophyceae (red algae), and Phaeophyceae (brown algae). The kelps (Laminariales),
members of the brown algae, live subtidally but in relatively shallow waters and can form large forests
along the cooler north Atlantic and Pacific coasts. These communities also may occur in the higher
salinity reaches of estuaries. Alongi (1998) provides a discussion of primary production, factors limiting
growth, nutrient cycling, and grazing in these communities.
Nutrient Criteria—Estuarine and Coastal Waters 3-15
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CHAPTER 4
Variables and Measurement
Methods To Assess and Monitor
Estuarine/Marine Eutrophic Conditions
Causal and Response Variables
Field and Laboratory Methods
Nutrient Enrichment and Ammonia Toxicity
4.1 INTRODUCTION
This chapter provides an overview of several measurable trophic state variables that can be used to
establish nutrient criteria for estuaries and nearshore coastal waters. Trophic state variables are those
variables that can be used to evaluate or predict the trophic status or degree of nutrient enrichment of
estuaries and nearshore coastal waters, especially when compared with reference conditions. The primary
variables include two causal variables (TN and TP) and two response variables including a measure of
algal biomass (e.g., chlorophyll a for phytoplankton or macroalgal biomass (AFDW) and water clarity,
e.g., Secchi depth or electronic photometer), and the addition of dissolved oxygen, as appropriate. These
variables are relevant at the national scale to practically all estuaries and are potentially relevant to
nearshore coastal waters.
Several variables are important indicators of nutrient overenrichment for a large number of estuaries, but
in many cases the data and supporting science are inadequate for most systems (e.g., algal species
composition). Important secondary variables include seagrass and estuarine submerged aquatic
vegetation (SAV) distribution and abundance, macroinfaunal community structure, phytoplankton
species composition, and organic carbon concentrations, respectively. Seagrasses and SAV typically
provide important shallow water habitat information, and hypoxia/anoxia are measures of loss of bottom
habitat often associated with deeper waters. Organic carbon (total, particulate, and dissolved) is also
included as a secondary variable because this variable is consistent with Nixon's (1995) definition of
eutrophication. Changes in benthic macroinfaunal community structure often correlate with organic
carbon enrichment and degree of hypoxia and anoxia (Diaz and Rosenberg 1995). The importance of
algal species composition has implications for food webs (Roelke 2000). These variables are discussed
in Chapter 2.
As indicated in Chapter 2, the concentration of the primary nutrient variables may not correlate well with
one or more response variables in estuaries, especially hypoxia or anoxia and measures of phytoplankton
biomass. In this case, predictive relationships should be attempted with nutrient loads using first
empirical regression models or other statistical approaches if necessary to account for ecosystem-based
nonlinearities. Application of mechanistic computer models is another approach (see Chapter 9).
Interpretation of nutrient enrichment indicators, especially for estuaries, is complicated by the interaction
with measures of mixing and flushing as discussed in Chapters 2 and 3. Salinity gradients are associated
with flushing but also play an important role in the type of biological communities exposed to nutrient
Nutrient Criteria—Estuarine and Coastal Waters 4-1
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enrichment. These physical considerations must always play a part in nutrient enrichment predictions
including establishment of reference conditions as discussed in Chapters 2 and 6.
4.2 CAUSAL AND RESPONSE INDICATOR VARIABLES
Nutrients as Causal Variables
Nitrogen
Nitrogen is one of the most important limiting nutrients of autotrophic assemblages (e.g., phytoplankton
and periphyton) incorporated into estuarine and nearshore coastal marine bioassessments. In those
estuaries where N has been demonstrated to limit algal biomass production, it typically does so at higher
salinities along the salinity gradient (Chapter 2). Most research has focused on the role of inorganic-N as
a stimulant to algal biomass production (Stepanauskas et al. 1999). However, about 70% of the dissolved
N transported by rivers worldwide (1012g yr"1) is dissolved organic N (DON) (Meybeck 1982). In
contrast to P, control of N sources is more difficult because diffuse gaseous sources of N (N2) can be
assimilated directly from the atmosphere by N fixation, a process conducted by a variety of bacteria and
cyanobacteria (blue-green algae). Also, dissolved inorganic N forms, especially nitrite and nitrate, are
highly soluble and do not precipitate easily or sediment out when freshwater enters the brackish zone of
estuaries as inorganic P is likely to do.
Total N measured as a water quality indicator consists of organic and inorganic forms. Although some
dissolved organic N may be used for algal growth, especially if remineralized by bacterioplankton
(Carlson and Graneli 1993; Seitzinger and Sanders 1999), it and particulate organic forms participate in
algal biomass production through recycling processes (Chapter 2). In systems with hypoxic or anoxic
conditions, the rate of decomposition is reduced. Although still an open question, apparently relatively
little of the DON is directly utilized by phytoplankton, except for urea and free amino acids (Antia et al.
1991; Paerl et al. 1999). Dissolved organic N in rainwater (synthetic addition of urea and other
constituents in bioassays) was shown experimentally to stimulate bacterioplankton and phytoplankton
growth; however, the DON resulted in the dominance of diatoms and dinoflagellates whereas
ammonium-N stimulated production more of small monads (Seitzinger and Sanders 1999). Further work
is required to test whether this response is widely applicable. Thus, the source of DON can influence the
degree of DON utilization by the microbial community. Inorganic N consists of ammonia, nitrite, and
nitrate N. Ammonia N is a primary product of microbial degradation of organic N, and, if not used
directly by autotrophic algae and vascular macrophytes and microbial heterotrophs for growth, it may be
oxidized through nitrification to nitrite and nitrate. Varying proportions of organic N may be relatively
refractive and contribute very little to N overenrichment problems. However, the readily recyclable
component may contribute to N enrichment problems locally and further seaward. Some experimental or
model analysis (e.g., box model) of the utilization of DON and in some cases particulate organic N for
each coastal system is usually warranted.
In estuaries, N concentrations, especially the inorganic forms, typically vary widely seasonally,
interannually, and along salinity gradients. In temperate river-dominated estuaries, nitrate concentrations
may reach very high concentrations (e.g., >100 \\M) in tidal fresh to brackish reaches (see Appendix G;
4-2 Nutrient Criteria—Estuarine and Coastal Waters
-------
Neilson and Cronin 1981) due to wash-off associated with various land use activities including point and
nonpoint sources (e.g., agricultural cropland). By late spring to early summer, the nitrate concentration
may be below analytical detection limits. Nitrite concentrations seldom reach high levels in surface
waters due to plant utilization and conversion to nitrate through nitrification. The principal bacteria
genera that mediate nitrification include Nitrosomonas, but species ofNitrosococcus, Nitrobacter, and
Nitrospina are also important (Sharma and Ahlert 1977, Watson et al. 1981). If dissolved oxygen is
limiting nitrification, then nitrite may accumulate (Helder and de Vries 1983). Ammonia concentrations
in open estuarine and nearshore coastal waters located away from point sources typically vary from
below detection limits to approximately 1.0 to 5 (iM, depending on growing season and rates of organic
N decomposition. Much higher values may occur for relatively short periods. The ionized form of
ammonia/ammonium is the most abundant reduced form and represents approximately 97% of the total
(Sillen and Martell 1964). The equilibration between the ionized and un-ionized fractions is controlled
by temperature, salinity, and pH, resulting in a range of un-ionized ammonia of 1% to 5% of the total at
typical salinities, pH, and temperature (Emerson et al. 1975). Ammonia may be toxic to marine larvae,
not just a stimulus to algal growth. Unionized ammonia concentrations in the range of 1.0|iM
approximate those that are known to be toxic to marine larvae, especially molluscs (U.S. EPA 1989).
Denitrification may remove from a few to approximately 50% of the TN load entering temperate
estuaries annually (Seitzinger 1988, Cornwell et al. 1999) depending largely on residence time of the
water, sediment biogeochemical conditions (macroinfauna present to maintain irrigation, oxic conditions
in the overlying bottom water), and water column depth. This process helps to modulate extreme DIN
concentrations (Chapter 2). Typical values for dissolved inorganic N (DIN) and a few TN concentrations
in estuaries and coastal nearshore waters are presented in Appendix G as a basis to help establish
expectations for various coastal systems. It should be noted that N concentrations vary widely in space
and time and the values in Appendix G are only intended to be rough guides. Specifics of analytical
techniques to measure the various forms of N are included at the end of this chapter (Field Sampling and
Laboratory Analytical Methods).
In open coastal waters of the North Atlantic Ocean at temperate latitudes, there is a typical seasonal
progression in DIN and DIP concentrations associated with phytoplankton blooms. The spring bloom
reduces these inorganic forms while phytoplankton biomass accumulates. This progression begins at
lower latitudes and moves to higher latitudes. The spring bloom typically crashes in late spring, and
summer biomass levels often are nutrient limited. Often a small bloom occurs in the fall following the
fall thermocline breakdown that allows mixing and replenishment of nutrients from deeper waters into
the upper surface layers, where a short burst of production occurs before light becomes limiting.
Accumulation of deepwater nutrients during the winter has been used to assess the potential for spring-
summer overenrichment in coastal seas based on trends in "salinity-nutrient mixing diagrams" (European
Union Northern Marine Eutrophication Criteria Program, Ulrich Claussen, Germany, personal
communication). Seasonal nutrient patterns in estuaries are quite variable. In some estuarine systems, a
winter buildup of N and P has been observed (e.g., Patuxent River Estuary), especially when freshwater
flows remained low and point sources dominated the nutrient supply (e.g., Flemer et al. 1970). Mixing
diagrams also help interpret nutrient behavior in estuaries; however, some precautions are important to
recognize (e.g., see Sharp et al. 1986).
Nutrient Criteria—Estuarine and Coastal Waters 4-3
-------
At the interface between fresh and marine waters, a process occurs that results in an apparent increase in
the ionized ammonia concentration. This process is apparently driven by the increased electrolyte
solution of the salts, which has a significant impact on the production and nitrification process, thus
yielding higher ionized ammonia levels (Rysgaard et al. 1999). Ionized ammonia adsorption to particles
was decreased, especially in the 0 to 10% salinity range, as were the nitrification and denitrification
processes. Further evaluation showed that the reduction in nitrification and denitrification processes was
due not only to the displacement of bacteria and ionized ammonia from particles, but also to decreased
bacterial activity. The projections from these studies were that ionized ammonia would be produced at a
rate of 1 (iM/g of sediment in the water. The changes in N dynamics that affect adsorption of suspended
solids may need to be included when considering acceptable levels in fresh water sources to estuaries.
Phosphorus
Phosphorus is an important plant nutrient that may limit algal biomass production in tidal fresh to
brackish zones of estuaries and some subtemperate to tropical marine coastal systems (Chapter 2). There
are no common stable gaseous forms of phosphorus, so the phosphorus cycle is endogenic, without an
atmospheric component (Manahan 1997). The main natural reservoirs of phosphorus are poorly soluble
minerals (e.g., hydroxyapatite) in the geosphere. Erosion of these materials from terrestrial sources and
their transport to the sea are important sources of new phosphorus in seawater. The phosphorus entering
the sea is mostly orthophosphate, PO4"3 (Kennish 1989). In previous decades, prior to widespread
phosphate bans in detergents, estuaries received a considerable portion of P from detergents. The ban
resulted for many estuarine systems in an elevated DIN:DIP ratio. In estuaries and nearshore coastal
waters, phosphorus is present in dissolved inorganic form as well as dissolved and particulate organic
form. Some fraction of P may be strongly embedded in a mineral matrix, and this renders that fraction
relatively inert to biological utilization. For this reason, often measures of TP may represent some
component that is not biologically available and managers should consider this in developing P criteria.
Plants directly take up the phosphates as essential nutrients during photosynthesis. Some algae have the
capability to break down dissolved organic P (DOP) with alkaline phosphatase (algal and free
phosphatases) and utilize the phosphate as inorganic phosphate (Huang and Hong 1999). Alkaline
phosphatase apparently is located on phytoplankton cell membranes, which makes it difficult to
determine whether the uptake is direct for DOP or the DOP undergoes enzymatic hydrolysis on the cell
membrane. Malone et al. (1996) suggested by inference that Chesapeake Bay phytoplankton may utilize
organic sources of P, in part, because the DIN:DIP thresholds approach 160, which is considerably
greater than the N:P ratio reported by Redfield et al. (1963). Orthophosphates are typically preferred by
autotrophic phytoplankton, although some assimilation of organic phosphorus may occur, especially
during periods of P deficiencies (Boney 1975). When plants die, or are eaten, the organic phosphorus is
rapidly converted to Orthophosphates through the action of phosphorylases within fecal material,
phosphatases in the plant cells, and finally by bacteria (Riley and Chester 1971).
To summarize, phosphorus occurs in natural waters and in wastewaters almost solely as phosphates.
These are classified as Orthophosphates, condensed phosphates, and organically bound phosphates
(common analytes are total phosphorus [TP] and dissolved or particulate organic phosphorus [DOP,
POP]). These compounds may be soluble, in particulates or detritus, or incorporated as organic P in
4-4 Nutrient Criteria—Estuarine and Coastal Waters
-------
organisms. Phosphorus is essential to the growth of organisms and can limit phytoplankton biomass
production, which is most commonly observed in freshwater systems (Hecky and Kilham 1988) and
some estuaries and coastal marine systems (Chapter 2). In instances where phosphate is limiting, the
discharge of raw or untreated wastewater, agricultural drainage, or certain industrial wastes may
stimulate the growth of algae. Appendix G provides examples of P concentrations in several forms.
Silica
Silica, as an important algal nutrient, has received much less attention in estuarine nutrient
overenrichment studies than N and P based on the limited volume of literature citations (e.g., see Malone
et al. 1996) and recent reviews of estuarine eutrophication (Chapter 1). Silica limitation of diatom
production, a major algal group that requires Si (and silicoflagellates), often is a measure of N or P
overenrichment (D'Elia et al. 1983; Conley and Malone 1992). Dissolved Si is a product of weathering
and erosion of rocks on land with subsequent transport to the sea (Conley and Malone 1992). Because Si
has essentially no human sources, except possibly from erodible soils under human influence, it is not a
strong candidate for regulation. In some parts of the ocean, organisms (such as diatoms and radiolarians)
abound that have produced skeletons of a noncrystalline form of hydrated silica-opal. As these skeletons
settle to the sea floor they slowly dissolve, releasing silica. Officer and Ryther (1980) predicted that
increases in N and P to estuaries and coastal waters from human activities, coupled with the reduction in
silicates to the sea from construction of artificial lakes, would alter the N:Si and P:Si ratios. These
alterations were postulated to alter phytoplankton populations to reduce the relative abundance of
diatoms and enhance the relative abundance of flagellates. Egge and Aksnes (1992) showed that diatoms
always numerically dominated the phytoplankton community when concentrations of silica were in
excess of 2.2 (JVI. Dominance by diatoms ceased or became more variable when concentrations of Si
were less than this value.
Ryther and Officer (1981) reinterpreted the relationship of N pollution in Long Island Inlets during the
1950s. Nitrogen may have limited the nuisance Nannochloris, blooms but they hypothesized that the
bloom persisted because diatoms had been eliminated by Si depletion. Also, the degree of Si limitation
of spring diatom blooms in Chesapeake Bay that fuel summer anoxia has direct ecological implications
(Conley and Malone 1992, Malone et al. 1996). Freshwater sources of Si dominate estuarine supplies
(Fisher et al. 1988). Typically, Si limitation can be potentially deduced from ambient ratios relative to
the nutrient-sufficient N:Si:P biomass ratios of 16:16:1 (Redfield et al. 1963; Conley et al. 1993). In
Chesapeake Bay, the dissolved Si:DIP ratio often approximates 100-300 (Malone et al. 1996), suggesting
strong Si limitation. Significant increases in Mississippi River N and P concentrations and loading and
decreases in silicate have occurred during the 20th century (Rabalais et al. 1996). The increased P
loading and associated increased diatom production and eventual burial in river sediments, as predicted
by Officer and Ryther (1980), has resulted in a reduced Si supply to the coastal environment. The
consequence is that diatom production, generality a preferred phytoplankton group to support higher
trophic levels, is now more Si limited than in previous decades. The N:P:Si ratios on coastal Louisiana
and Texas now suggest the possibility of a joint nutrient limitation of phytoplankton production.
Nutrient Criteria—Estuarine and Coastal Waters 4-5
-------
Silica concentrations for the Coastal Texas/Louisiana coast averaged approximately 5.3 (iM in the late
1980s but averaged about 9.0 (iM during the early 1960s. Silicate concentrations in the Chesapeake and
Delaware Bays and the Hudson River Estuary ranged from about 90 to near detection levels, 30 to near
detection limits, and 30 to 3 (iM, respectively (Fisher et al. 1988). Eyre and Balls (1999) reported that Si
was less likely to limit diatom production in tropical estuaries than in temperate ones because
concentrations tend to be much higher in tropical estuaries.
The role of silica may be more important to diatom species composition and food quality as future
research may document. More attention in the future should be given to the measurement and assessment
of the role of Si in estuarine and nearshore coastal primary productivity and food web dynamics and as a
basis for controlling co-limiting N and/or P.
Response Variables
Chlorophyll a andMacroalgal Biomass
Chlorophyll a is the molecule mediating photosynthesis in most all green plants (except prochlorophytes,
which contain divinyl chlorophyll), including phytoplankton; it is relatively easy to measure either
spectrophotometrically or by fluorescence and is commonly used to indicate phytoplankton biomass.
However, the amount of chlorophyll per cell can vary widely. Conversion factors from weight of
chlorophyll to weight of carbon (a desired biomass unit) can vary by a factor of 10. Adaptation to light
levels is the primary reason for observed variability; photoadaptation can cause the chlorophyll per cell
to vary widely. The technology for measuring chlorophyll has greatly improved over the decades. The
Welschmeyer (1994) fluorometric analysis reduces the interference due to chlorophyll b and
phaeopigments. The HPLC procedure is capable of detecting and quantifying various pigments
characteristic of different algal groups (e.g., diatoms, cyanophyta, chlorophyta, and dinoflagellates)
(Jeffery et al. 1997).
Rapid proliferation or blooming of phytoplankton, as reflected in chlorophyll a measurements, occurs
throughout the ocean but is most often associated with temperate coastal and estuarine waters and at
higher latitudes. In winter months, growth of phytoplankton populations is generally minimal because of
insufficient light and also because a turbulent and unstable upper water column carries the phytoplankton
cells below the euphotic zone (where light is not sufficient) before they can divide.
Chlorophyll a concentrations vary widely as a function of nutrient supply, water column stability,
euphotic zone depth (light availability), sinking, grazing, disease organisms (e.g., viruses), and
flushing/mixing (Chapter 2). Values in excess of 12 to 15 (ig/L are likely to cause severe shading of
seagrasses (Kelley in press). Concentrations in estuaries during summer optimum growing conditions
may exceed 50 to 80 (ig/L when nutrient loading is high (Monbet 1992). Summer values in the range of
20 to 40 (ig/L are frequently observed in enriched estuaries. In contrast, concentrations in overenriched
temperate U.S. estuaries during the winter may decrease to 1 to 5 (ig/L Nearshore coastal areas removed
from high nutrient loads may experience chlorophyll concentrations in the range of approximately 1 to 3
(ig/L (Appendix G). Very high values may occur during the summer under conditions of high levels of
4-6 Nutrient Criteria—Estuarine and Coastal Waters
-------
nutrient enrichment (e.g., the Mississippi River Plume on the Texas/Louisiana Shelf [Rabalais et al.
1996]).
Macroalgal biomass, especially benthic unattached forms (i.e., Ulva spp.), often becomes abundant in
relatively shallow estuaries that experience nutrient overenrichment. In estuaries that receive most of
their nutrient load from groundwater (e.g., Waquoit Bay, Cape Cod, MA; see Chapter 2) benthic
macroalgae may shade out seagrasses. Continued enrichment typically leads to reduction of macroalgae
as phytoplankton predominate in the water column. Macroalgae are difficult to adequately sample for
chlorophyll a, and thick mats often contain sheets of algal material that has begun to degrade. The most
common method to sample benthic macroalgae is to collect samples and express the biomass on a dry
weight basis.
Measures of Water Clarity
Light Attenuation Coefficient
The Secchi disc has been a mainstay as a tool in estimating water clarity; however, this simple and
inexpensive tool does not provide all of the information required to distinguish the light attenuation
effects of living phytoplankton pigments (i.e., traditionally estimated by chlorophyll a) from other factors
(e.g., inorganic suspended sediments, organic nonchlorophyll-based detritus, and humic-like materials)
that reduce water clarity. EPA's Chesapeake Bay Program (Chapter 2) has developed an analytical
approach that partitions the effect of chlorophyll a from total suspended solids that contribute to
reduction in water clarity. This approach has been used successfully in estimating the combined factor
contribution to light attenuation over submerged aquatic vegetation beds (Dennison et al. 1993). In
turbid coastal waters, the analyst should be aware of lower values for the constant 1.7 to estimate the
light attenuation coefficient (see Giesen et al. 1990 and references in Chapter 2). More precise estimates
of the light attenuation coefficient can be made with electronic submersible light meters including PAR
meters (photosynthetic active radiation) and submersible spectral radiometers. These meters are now in
widespread use, and their use should be encouraged because they give a direct measure of light
attenuation, especially in shallow water where depth may limit use of the Secchi disc.
Attenuation of light in the sea in nonalgal bloom areas is determined principally by the amount of
suspended matter present, but in estuaries and nearshore coastal waters, color from humic-like materials
may significantly compete with particulate material in light attenuation. In moderately turbid coastal
waters, 1% of the surface visible light energy may penetrate to a depth of only 10 to 20 m, but in shallow
estuaries depths often are from 10 cm to 3 m or so. There typically is a strong seasonal variability in
water clarity in temperate estuaries between the active growing season and the winter, and in
subtemperate to tropical estuaries water clarity often is a function of the wet season. In the Atlantic
temperate open coastal areas with the coming of spring, the depth of the euphotic zone often increases
and the depth of the mixed layer decreases because of the development of the seasonal thermocline. This
allows a spring bloom to develop. The thermocline tends to confine the algal cells to the euphotic zone,
which becomes rich with nutrients as a result of winter mixing. In estuaries, the pycnocline may also
have this effect. In partially mixed estuaries where light is adequate at depth, diatoms may grow below
the pycnocline (Malone et al. 1996). If the necessary growth-promoting factors are also present,
Nutrient Criteria—Estuarine and Coastal Waters 4-7
-------
conditions are optimal for proliferation of phytoplankton from seed stock, which may be either the
plankton cells themselves or their resting stages (Riley and Chester 1971).
Seech i Depth
The Secchi disc is a useful tool to estimate water clarity (Holmes 1970). Secchi disc measurements often
have a longer historical record than electronic measurements, which facilitates assessment of trends in
water clarity. Secchi depth measurements are obtained with a 40 cm plastic or metal Secchi disk that is
either white or is divided into black and white quadrants on a nonstretchable line that is calibrated in
decimeters. The disc should be weighted to maintain a level position, especially under strong current
conditions. The disk is lowered into the water until it disappears from view and the depth is recorded.
The disk is then slowly raised to the point where it reappears and the depth is recorded again. The mean
of these two measurements is the Secchi depth. Observations are made from the shady side of the vessel
to reduce problems of glare; however, when a small boat is used for field work a "viewing tube" allows
readings under full sunlight conditions. Measurement should be made without sunglasses.
Dissolved Oxygen
Dissolved oxygen (DO) is an integrative measure of ecosystem health and habitat function. As a first-
order estimate, the percent saturation of surface and bottom waters is an index of the
production/respiration ratio. Dissolved oxygen in bottom waters serves as a measure of habitat
availability for benthic animals and pelagic animals that feed on the bottom. EPA has developed
saltwater DO criteria for coastal waters between Cape Cod and Cape Hatteras (see
www.epa.gov/ost/standards/dissolved). Profiles of DO are indicative of oxygen depletion conditions
such as hypoxia and anoxia. Lack of oxygen in bottom waters causes sediment to release dissolved
nutrients including orthophosphorus, ammonia, and in addition, toxic hydrogen sulfide.
Carbon Compounds
Organic matter content is typically measured as total organic carbon (TOC) and dissolved organic carbon
and is an essential component of the carbon cycle. The rate of organic carbon production and
decomposition and the resulting microbial biomass are at the heart of the eutrophication problem.
Evaluation of the carbon-containing compounds in an aquatic ecosystem can indicate its organic
character. The larger the carbon or organic content, the greater the growth of microorganisms that can
contribute to the depletion of oxygen supplies. TOC is a more convenient and direct expression of
organic carbon content than are the biochemical oxygen demand (BOD), assimilable organic carbon
(AOC), or chemical oxygen demand (COD) methods. TOC is independent of the oxidation state of the
organic matter and does not measure other organically bound elements, such as N and hydrogen, or
inorganics that can contribute to the oxygen demand measured by BOD and COD. In spite of its
versatility, TOC does not provide the same kind of information as BOD, AOC, or COD, and should not
be used to replace these methods.
At the surface of the sea, the concentrations of particulate and dissolved organic carbon range up to 12.5
(iM and between 75 and 150 (iM, respectively. In coastal environments, concentrations of dissolved and
particulate organic carbon are greater by factors of ~7-fold. Concentrations of dissolved and particulate
4-8 Nutrient Criteria—Estuarine and Coastal Waters
-------
organic carbon in surface waters are equivalent to 150 to 1,800 (ig C/L (Millero 1996). Organic carbon
represents approximately 50% of the dissolved and particulate organic material in seawater (Millero
1996). However, the major form of carbon in seawater is associated with inorganic carbonate systems.
Benthic Macroinfauna
Benthic macroinfauna are an important biological component of estuarine and nearshore coastal marine
ecosystems. These communities contribute to benthic food webs, contribute to nutrient cycling and
system productivity through benthic-pelagic coupling of nutrient recycling, help stabilize bottom habitats,
and contribute to marine biodiversity. Benthic infaunal communities are quite diverse within an estuary
or coastal region. Diversity is a function of salinity, with higher diversities associated with higher
salinities (Carriker 1967). Sediment irrigation provided by benthic infauna enhances denitrification by
increasing the flux of ammonium into oxic microenvironments where nitrification can occur and the flux
of nitrite and nitrate into the anoxic sediment zone where denitrification becomes possible (Chapter 2).
4.3 FIELD SAMPLING AND LABORATORY ANALYTICAL METHODS
The following sections provide additional information on field sampling and laboratory methods for
selected variables. A list of suggested methodologies for analysis of biochemical parameters is provided
in Table 4-1. These methods have been summarized from nationally or regionally recognized reference
compendiums (APHA 1998, ASTM 1976, U.S. EPA 1979, Spotte 1992) and provide acceptable methods
for determining the concentrations of nutrients as well as acceptable methods for measuring the effects of
those nutrients in estuarine and marine waters.
Field Sampling Methods
Nutrients, Hydrography, and Sediments
Physiochemical profiles should be recorded for each field sampling station. Important parameters to be
measured include water temperature, pH, dissolved oxygen, salinity, light attenuation, surface radiation,
and total depth. Generally, a multiparameter water quality instrument CTD is used. Sampling depth will
vary depending on specific objectives; however, enough vertical depth reading should be taken to
characterize the physical structure of the water column. For example, CDT measurements might be taken
at frequent intervals in the vicinity of the pycnocline (e.g., every 0.1 m in highly stratified estuaries).
Overall current dynamics can be mapped with oceanographic tools such as current meters, drift cards,
and acoustic Doppler sounders.
Field sampling of discrete water samples for laboratory analysis can be performed using standard
nonmetallic plastic water bottles. Samples are drawn into prelabeled bottles and fixatives are applied as
appropriate to the subsequent analysis. Nutrient and organics samples are stored on ice until reaching a
shoreside sample handling location. Nutrient samples are filtered using graduated syringes and then
frozen. Samples for total TN and TP are filtered or unfiltered as appropriate, and 20 mL of sample is
frozen for analysis. See Chapter 5 for additional sampling protocols.
Nutrient Criteria—Estuarine and Coastal Waters 4-9
-------
Table 4-1. Suggested methods for analyses and monitoring of eutrophic conditions of coastal and marine
environments (* = primary EPA preferred causal and response variables)
Eutrophication indicators
Field
*Water clarity
PH
Dissolved oxygen
Salinity
Light attenuation
Temperature
Laboratory analyses
*Total phosphorus
(including
orthophosphate,
POP, and DOP)
Dissolved
orthophosphate
Particulate phosphorus
*Total N, incl. DON,
DIN, and PON1
Total KjeldahlN
Ammonia/ammonium
Nitrate
Nitrite
Suggested methods
Secchi depth
CTD probe
CTD probe
Salinometer
Sensor
CTD probe
SM 4500P-E
SM 4500P-E
EPA 365.2
CBP IV.D.2
CBPIV.D.3
CBP IV.D.4
SM 4500N-C
ASTMD3867
EPA
EPA-AERP18
CBPIV.D.8
SM 4500org-C with
SM 4500NH3-H
EPA351.3/.1
(mod.)
SM 4500NH3-B/H
EPA 350.1
CBP IV.D.7
SM 4500NO3-F
EPA 353.2
SM 4500NO3-F
EPA 353.2
SM 4500NO2-B
EPA 354.1
Detection limit
or range
O.lm
0.01 pH
0.02 mg DO/L
0.1 psu
0.05%@100%light
0.1 °C
0.3 uM
0.32 uM
—
0.03 uM
0.02 uM
0.04 uM
0.36 uM
0.7-143 uM
—
—
1.9uM
—
1.4-1429 uM
—
1.4-1429 uM
0.7-1429 uM
0.3 uM
35.7-714 uM
—
35.7-714 uM
—
0.7-71 35.7-714 uM
—
Comments
—
—
or Winkller Azide Mod.
—
e.g.,LI-COR-LI-192SA
sensor
—
Ascorbic acid method
Auto, persulfate method
—
Auto, persulfate method
Ascorbic acid method
Ascorbic acid method
Persulfate method
Persulfate method
Persulfate method
—
Auto, persulfate method
Semi-micro-Kjeldahl method
Auto, phenate method
Colorimetric/titration
Auto, phenate method
Colorimetric phenate
Auto, phenate method
Auto, cadmium reduction
—
Auto, cadmium reduction
—
Colorimetric method
—
References
EPA 903-R-96-006
—
—
—
—
—
APHA 1998
APHA 1998
EPA 600/4-79-020
EPA 903-R-96-006
EPA 903-R-96-006
EPA 903-R-96-006
APHA 1998
ASTM1976
EPA 903-R-96-006
EPA 600/4-87-026
EPA 903-R-96-006
APHA 1998
APHA 1998
EPA 600/4-79-020
APHA 1998
EPA 600/4-79-020
EPA 903-R-96-006
APHA 1998
EPA 600/4-79-020
APHA 1998
EPA 600/4-79-020
APHA 1998
EPA 600/4-79-020
4-10
Nutrient Criteria—Estuarine and Coastal Waters
-------
Table 4-1. Suggested methods for analyses and monitoring of eutrophic conditions of coastal and marine
environments (* = primary EPA preferred causal and response variables) (continued)
Eutrophication indicators
Nitrate + nitrite
Particulate N
Total organic carbon
Dissolved organic
carbon
Particulate carbon
Total silicates
Total suspended solids
Total volatile solids
BOD
COD
Biological measures
Phytoplankton biomass
Zooplankton biomass
Chlorophyll a°
Suggested methods
CBPIV.D.5
SM 4500NO3-F
EPA 353.2
EPA 4.1.4
CBP IV.D.6
CBPIV.D.8.10
SM5310TOC-D
SM5310TOC-C
EPA 415.1
SM5310TOC-C
EPA 415.1
ASTMD2574-79
CBPIV.D.10
CBP.IV.D.9
SM 4500SiO2-D
ASTMD859-68
CBP-IV-15
EPA 370.1
CBPIV.D.15
SM 2540-D
CBCIV.D.13
SM 2540-E
Estuarine
SM5210-B
EPA 405.1
CBPIV.D.ll
SM 5220-D
EPA 410.4
—
SM 10200-H
Detection limit
or range
0.01 iiM
35.7-714 iiM
—
0.7-143 iiM
0.01 iiM
1.36 iiM
>0.1 mgC/L
>0.01 mg TOC/L
—
>0.01 mg TOC/L
—
—
0.5 mg/L
0.097 mg/L
0.33-0.83 iiM
—
0.17-23. 3 iiM
—
0.22 iiM
2-20,000 mg/L
2.0 mg/L
—
—
—
—
—
—
—
—
0.01 mg/M3
Comments
Auto, colorimetric method
Auto, cadmium reduction
—
Technicon autoanalyzer
Auto, colorimetric method
Filtration/combustion
Wet oxidation method
Persulfate method
—
Persulfate method
—
—
Catalytic combustion
Filtration/combustion
Heteropoly blue method
—
—
Molybdosilicate method
Dried at 103-105°C
Filtration/heat
—
—
5 -day method
—
5 -day method
—
—
—
Fluorometric, HPLC, Spectro.
References
EPA 903-R-96-006
APHA 1998
EPA 600/4-79-020
EPA 503/2-89/001
EPA 903-R-96-006
EPA 903-R-96-006
APHA 1998
APHA 1998
EPA 600/4-79-020
APHA,1998
EPA 600/4-79-020
ASTM1976
EPA 903-R-96-006
EPA 903-R-96-006
APHA 1998
ASTM1976
EPA 903-R-96-006
EPA 600/4-79-020
EPA 903-R-96-006
APHA 1998
EPA 903-R-96-006
APHA 1998
EPA 430/9-86-004
APHA 1998
EPA 600/4-79-020
EPA 903-R-96-006
APHA 1998
EPA 600/4-79-020
—
APHA 1998
Nutrient Criteria—Estuarine and Coastal Waters
4-11
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Table 4-1. Suggested methods for analyses and monitoring of eutrophic conditions of coastal and marine
environments (* = primary EPA preferred causal and response variables) (continued)
Eutrophication indicators Suggested methods
EPA AERP12
ASTMD3731-79
CBPIV.D.12
Phaeophytin SM 10200-H
EPA AERP12
ASTMD3731-79
CBPIV.D.12
Dinoflagellate density —
Diatom density —
Dinoflagellate/diatom —
Perennial plant density —
Ephemeral plant density —
Epiphytic growth —
Phytoplankton blooms —
Fish kills —
Detection limit _ _ „
Comments References
or range
— — EPA 600/4-87-026
Spectrophotometer ASTM1976
l.Ong/L Spectrophotometer EPA 903-R-96-006
0.01 mg/M3 Fluorometric, HPLC, Spectro. APHA 1998
— — EPA 600/4-87-026
— Spectrophotometer ASTM1976
l.Ong/L Spectrophotometer EPA 903-R-96-006
— — —
— — —
— — —
— — —
— — —
— — —
— — —
— — —
a DON, dissolved organic N, DIN, dissolved inorganic N; PON, particulate organic N.
b Phytoplankton segments: The HPLC procedure is capable of detecting and quantifying various pigments characteristic of
different algal groups (e.g., diatoms, cyanophyta, chlorophyta, and dimoflagellates) (Jefferey et al. 1997).
4-12
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Laboratory Analytical Methods
Detailed methods and references are given in Table 4-1. Some general considerations are presented in
the following sections.
Water Column Nutrients
Nitrogen Compounds
Several methods have been used to determine the concentration of N species in the marine environment.
Methods presented in this document are relatively easy to use, do not require extensive instrumentation,
provide detection limits below those expected in marine environments, and are in general use by many
investigators. The most common forms of N in eutrophication evaluation in order of decreasing
oxidation state are nitrate, nitrite, ammonia, and organic N. The sum of these is expressed as TN and is
not to be confused with total Kjeldahl N (TKN), which is the sum of organic N and ammonia. Total N
can be determined through oxidative digestion of all digestible N forms to nitrate, followed by
quantitation of the nitrate. Nitrite is an intermediate oxidation state of N, both in the oxidation of
ammonia to nitrate and in the reduction of nitrate. Such oxidation and reduction may occur in
wastewater treatment plants, water distribution systems, or natural waters. Ammonia is produced largely
by deamination of organic N-containing compounds and by hydrolysis of urea. The two major factors
that influence selection of the method to determine ammonia are concentration and presence/absence of
interferences (e.g., high concentrations of colored organic substances such as humic-like materials or
paper mill effluents).
Total N is measured by the persulfate method, which digests all organic and inorganic - containing
compounds. All N-containing materials (except nitrogen gas) are measured after sample digestion has
occurred. Various organizations have adjusted sample volume or automated the process and produced
different ranges of detection. The lowest detectable concentration is ~ 0.7 (iM of TN. This is in the
range of the measured available N (0.7 to 5.0 (iM TN) for studies performed off the continental shelf in
the North Atlantic from 1956 to 1958 (Kennish 1989). Kjeldahl N minus the ammonia concentration is
the surrogate measurement for all organic N-containing compounds.
Ammonia/ammonium is measured by the indophenol blue (= phenate) or specific ion electrode methods
after conversion of ammonia and ammonium to ammonia. This is done by raising the pH of the sample
above 11. This method has some essential features (e.g., minimal interference from waters highly stained
with humic materials and paper mill effluents); however, the level of detection is relatively high (e.g., 2.0
(iM NH3-N) but adequate for ammonia-rich waters (Flemer et al. 1998). Ammonia electrodes do not
work directly in seawater. In the spectrophotometric methods, the ammonia is reduced to
monochloramine and then reacted with phenol to form a blue color. In the specific ion electrode method,
the ammonium is converted to ammonia using a strong basic solution and partial pressure of ammonia
gas (i.e., free ammonia) in solution, which is related to the dissolved ammonia concentrations by Henry's
Law.
Nitrates and nitrites are measured in combination using the cadmium reduction procedure of Wood et
al. (1967). This colorimetric method determines the concentration of these two materials after reaction
Nutrient Criteria—Estuarine and Coastal Waters 4-13
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of nitrites to produce an azo dye, the color of which is proportional to the concentration of the combined
nitrates and nitrites. Total nitrate is determined by subtracting the concentration of nitrite from the
combination of the two. The process for measurement of nitrite produces the same azo dye as the
combined measure, but without the Cd reduction. The difference in these two measures is the nitrate
concentration.
Phosphorus
The target detection limit for measurement of P in seawater is ~0.3 (iM. The procedures for the
measurement of total particulate and dissolved P as well as orthophosphate in seawater provide detection
limits that are less than this value (U.S. EPA 1996). These procedures convert the phosphorus-
containing compounds to orthophosphate through the digestion of the sample with alkaline persulfate.
This treatment is then reacted with ammonium molybdate and antimony potassium tartrate in acidic
solution to produce an intense blue complex with ascorbic acid. Interferences with elevated
concentrations of Si can be avoided by maintaining an acid concentration in the reagents and analyzing
the material at elevated temperatures of ~37°C. The resulting phosphomolybdic acid reduction produces
a purple-blue complex that is measured at 885 nm on a spectrophotometer. This method of measuring
reactive silicate is recommended in Millero (1996).
Silica
The target detection limit for measurement of Si in seawater is ~0.7 (iM. Producing pigmented
silicomolybdate complex by procedures contained in U.S. EPA (1996) provides adequate sensitivity after
the samples are filtered (0.45 \\m GF/F filter) to remove interfering particles and turbidity, and after the
interferences of phosphates and arsenates are removed with oxalic acid. The resultant filtrate is treated
with a solution containing metol-sulfate (p-methyl-amino-phenol sulfate) to produce a blue color that is
evaluated more efficiently than the yellow color recommended for evaluation in U.S. EPA (1996), with a
spectrophotometer at 812 nm (Strickland and Parsons 1968). This method of measuring reactive silicate
is also recommended in Millero (1996).
Carbon
Total carbon consists of inorganic and organic forms that are in particulate and dissolved size classes.
The distinction between total and organic carbon is based on acidifying samples to remove the inorganic
forms and filtering through 0.45 (im GF/F filters to remove the particulate forms. Total carbon is
measured by burning the sample to release the particles contained on the glass fiber filter. This converts
the carbon to CO2, which is then transported to a thermal conductivity detector for measurement. The
carbon left behind in the filtrate is catalytically combusted using a platinum catalyst at ~680°C that is
then transported to a nondispersive infrared detector. The EPA methods (U.S. EPA 1996) will provide
adequate detection of both dissolved and particulate carbon in the total and organic phases. The
difference in total carbon and organic carbon represents the inorganic fractions that are primarily CaCO3
shells.
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Sediment Analyses
Bulk Sediment
Cores are collected from field sites to help determine the historical record and sedimentation rate. Short
cores, the upper 30 cm of the substrate, can be obtained with a HAPPS core, designed to collect a
relatively undisturbed core of surficial sediment (Kannerworff and Nicolaisen 1973) and used to profile
sedimentary particulate organic carbon and N. Carbon-N analyses follow the method of Hedges and
Stern (1984); samples for dissolved constituents in pore water are extracted either by whole-core
squeezing or by centrifugation (Devol et al. 1997, Brandes and Devol 1995, Lambourn et al. 1991). Deep
coring devices are used to collect continuous sediment core samples 2 to 3 m into the sediment bed.
These deeper cores are used for analysis of 210Pb, carbon and N, sulfide, and biogenic silica in order to
determine burial rates of 210Pb and 210Ra.
The sedimentation rate is estimated based on the change in activity of naturally occurring 210Pb
radionuclide produced at a constant rate from the decay of 210Ra, using the excess 210Pb inventory method
of Anderson et al. (1987). Excess 210Pb is determined from the difference between total 210Pb activity in
the sediment and the activity of the background 210Pb being produced from 210Ra. To collect samples for
measurement of 210Pb and 210Ra activity at depth with the sediment, cores are sectioned and each section
is then homogenized and placed in a precleaned 16 oz jar, with a small subsample removed and placed
into a glass vial for particulate C and N analysis (Evans-Hamilton, Inc. 1998).
The excess 210Pb inventory method yields accumulation rates (g/(cm2/yr)), which are converted to a
sedimentation rate (cm/yr) using the bulk sediment density g/cm3. For evaluation of seasonal trends, the
upper cm is subsampled at 0.25 cm intervals, and in 1 cm intervals below the first cm, following the
assumption that any seasonal storage of N or carbon would manifest almost entirely at the surface of the
sediment.
Pore Water Profiles
Pore water profiles of manganese, iron, nitrate, and oxygen demonstrate that oxidation of iron and
magnesium yields less energy than does oxidation of carbon by oxygen or nitrate. Consequently,
concentration peaks of these species are located below the depletion depths of oxygen and nitrate. In
anaerobic environments, after the supplies of oxygen, nitrate, manganese, and iron are exhausted, sulfate
reduction is the dominant mode of organic matter oxidation and nutrient remineralization.
Sulfate reduction rate can be measured with the radiotracer method of Christensen et al. (1987). A
significant fraction of the oxygen flux may be consumed by the reoxidation of sulfide produced during
sulfate reduction (Canfield 1993).
Sediment traps are used to measure the quantity and composition of the flux of materials settling through
the water column to the sediment. There are four materials of interest: chlorophyll as an indicator of
planktonic algal remains, pheaopigments as an indicator of degraded plankton that has been consumed by
zooplankton, particulate organic carbon (POC), and particulate organic N (PON). Total sedimentation
rate is corrected for resuspension materials in order to derive the net flux to sediment. Samples are
Nutrient Criteria—Estuarine and Coastal Waters 4-15
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collected by in situ benthic flux chambers, and measurements of oxygen, silicate, nitrate, ammonium,
phosphate, and N gas are made (Evans-Hamilton, Inc. 1998).
Determination of Primary Productivity
Primary productivity refers to the growth rate of the phytoplankton community and is commonly
measured using trace amounts of radioactive carbon (as bicarbonate) that label the photosynthetic
reaction. Additional variables are measured to support these data: biomass (as estimated by chlorophyll
a), incoming solar radiation, and nutrient concentrations at depth. Primary productivity, P, is defined as
P = (i x B
where (i is the specific growth rate (growth normalized per cell) and B is the biomass of the
phytoplankton population (amount of cells). These variables are > 'compound' = as they in turn depend
on other variables. Growth rate depends on light (solar radiation), dissolved nutrients in the water
column, and water temperature. The phytoplankton biomass is determined by the net result of growth
and loss (grazing, mixing, sinking) processes and reflects enrichment conditions.
To estimate primary productivity, samples are collected at varying depths corresponding to
predetermined light levels. Fresh samples at each light level are collected for analysis of chlorophyll a,
nutrients, and primary productivity in two sets of two clear bottles and one dark bottle; each set is filled
for ambient treatment and nutrient spike treatment. Nutrient spiking consists of adding an initial
concentration of 10 (iM N (NH4CL) and 1 (iM phosphorus (KH2PO4) to seawater. Nutrients are
monitored from additional samples collected and tested for nitrate, nitrite, ammonium, orthophosphate,
and silicate. Samples are inoculated with 14C-labeled sodium bicarbonate and, if appropriate, the nutrient
spike, and placed in a screened bag to simulate the light level from which they were collected. Samples
are incubated at in situ conditions for 24 hours and then transported to the laboratory for filtration using
glass fiber filter paper (Whatman GF/F, nominal pore size 0.7 (im or smaller pore size). The filters are
placed into vials containing EcoLume scintillation cocktail. The specific activity of the filtered
particulates is measured in a scintillation counter. Primary production is calculated as mg C/(m3/day)
using the basic equations found in Parsons et al. (1984) (Evans-Hamilton, Inc. 1998).
In productive coastal waters, measurements using the light and dark bottle technique with changes in
dissolved oxygen often can be used in place of the 14C method (Strickland and Parsons 1968). In some
cases, free water gas-based (e.g., DO) methods are possible to measure ecosystem metabolism (Odum
1956; Odum et al. 1959; Kemp and Boynton 1980).
Phytoplankton Species Composition
Samples collected from the field are analyzed to identify and enumerate autotrophic phytoplankton, as
well as heterotrophic dinoflagellates and microzooplankton species. From 20 to 50 mL aliquots of
samples are settled in separable counting chambers for at least 24 hours before examination under
phase-contrast optics with an inverted microscope following the classic Utermohl technique (Lund et al.
1958). A single transect across the center of the chamber is counted at 390 x magnification for
4-16 Nutrient Criteria—Estuarine and Coastal Waters
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flagellates; 150x magnification is used for other organisms. From 25% to 100% of the chamber bottom
is examined, depending on cell concentrations in the sample. Appropriate multipliers are used to convert
all counts to common units of cells/L (Sournia 1978). Organisms are identified to the lowest taxonomic
category possible. Even quite small changes in the physical and chemical parameters and availability of
micronutrients can have a significant effect on the growth constants of algae. A difference in doubling
time of 25% between two fast-growing organisms can lead to one outnumbering the other by 15 to 1 in a
week and quickly lead to alterations in species assemblages (Riley and Chester 1971).
There are numerous algal species in estuarine and open coastal waters that are considered to be harmful
(e.g., see Dortch et al. 1998, Anderson and Garrison 1997, Anderson 2000). This is a rapidly changing
area of marine ecology and experts should be consulted for specific taxonomic identifications.
Macrobenthos, Macroalgae, and Seagrasses and SAV
Macroinfauna are typically sampled with coring devices or bottom grab samplers and wet-sieved through
0.5 (iM mesh sieves to separate the animals from very fine sediments. Stacked sieves can be used to
remove larger shell fragments and sand particles. A relaxant (e.g., 0.3% propylene phenoxytol) is
applied prior to addition of formalin. Samples are usually preserved in 10% buffered formalin for several
weeks and then transferred to 60%-70% isopropanol (Diaz and Rosenberg 1995).
Macroalgae are typically sampled by collecting algal material by hand from a known surface area of the
habitat. Various devices may be used (e.g., 0.5 m stainless or plastic hoop).
Both above- and below-ground seagrass and SAV biomass can be collected from a known area of the
bed. Various techniques have been used. An often-used method is to shove metal strips along the
sediment surface in a square meter pattern and anchor the strips at all four corners by pushing a sharp
spike through holes drilled at each end of the strips. Then, the plant material separated to species can be
clear-cut with sharp shears and taken to the laboratory and dried in a heated cabinet at 60°C to constant
dry weight. A sharp spade is required to collect below-ground roots and rhizomes. This material should
be identified and dried to constant weight.
Nutrient Criteria—Estuarine and Coastal Waters 4-17
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CHAPTER 5
Databases, Sampling Design,
and Data Analysis
Developing National/Regional Databases
Sampling Design
Quality Assurance/Quality Control
Statistical Analyses
5.1 INTRODUCTION
Development of national regional numeric nutrient criteria requires that an extensive amount of data
from across the country be evaluated. This information can be an invaluable tool to States and Tribes as
they develop nutrient criteria. Both existing and historical data may provide considerable information
that is specific to the region where criteria are to be set. First the data must be located, then the
suitability of the data (type and quality) ascertained before they can be used for analysis of water quality
parameters. It is also important to determine how the data were collected to make future monitoring
efforts compatible with earlier approaches. Descriptive data that characterize the waterbody are
invaluable.
Data may come from existing sources or can be collected from new sampling programs. Nutrient-related
data for estuaries and coastal waters, collected by various agencies for many different purposes, exist in
numerous databases and have the potential to provide the basis for development of nutrient criteria on a
regional level. This chapter presents an overview of existing databases and a general discussion
concerning the evaluation of such datasets in terms of their use in the nutrient criteria development
process. The list of databases is not all-inclusive—many other data sources exist—but the list provided
is intended to represent the kind of information that is available. This chapter also provides a description
of existing data resources (e.g., U.S. EPA Legacy STORET and ODES) and how these data may be used
to generate preliminary nutrient criteria on regional levels. In addition to discussing the use of existing
data, the chapter discusses new data collection, including consideration for sampling design and the types
of sampling to be considered as part of data collection activities. The chapter ends with a general
discussion of data management, quality assurance, and quality control issues that are integral in the
overall discussion of data storage, accessibility, and utilization.
5.2 DEVELOPING REGIONAL AND NATIONAL DATABASES FOR ESTUARIES
AND COASTAL WATERS
A database is a collection of information related to a particular subject or purpose. Databases are
arranged so that they divide data into separate electronic repositories in tabular format. Data in tables
can be viewed and edited, and new data can be added. A single datum is stored in only one table but can
be viewed from multiple locations. Updating one view of a datum will update it in all the various
viewable forms. Each table should contain a specific type of information. Data from different tables can
be viewed simultaneously according to the user-defined table relationships. That is, the relationship
among data in different tables can be defined so that more than one table can be queried or reported and
accessed in a single view. Data stored in tables can be located and retrieved using queries. A query
Nutrient Criteria—Estuarine and Coastal Waters
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allows the user to find and retrieve only the data that meet user-specified conditions. Queries also can be
used to update or delete multiple records simultaneously and to perform built-in or custom calculations of
data. Data in tables can be analyzed and printed in specific layouts for reports.
To facilitate data manipulation and calculations, it is highly recommended that historical and present-day
data be transferred to a relational database. A relational database is a collection of data items organized
as a set of formally described tables from which data can be accessed or reassembled in many different
ways without having to reorganize the database tables. Each table contains one or more data categories
in columns. Each row contains a unique instance of data for the categories defined by the columns. The
organization of data into relational tables is known as the logical view of the database. Relational
databases are powerful tools for data manipulation and initial data reduction. They allow selection of
data by specific and multiple criteria and definition and redefinition of linkages among data components.
Geographic information systems (GIS) are geo-referenced databases that have a geographic component
(i.e., spatial platform) in the user interface. Spatial platforms associated with a database allow
geographic display of sets of sorted data and make mapping easier. These types of databases with spatial
platforms are becoming more common. The system is based on the premises that "a picture is worth a
thousand words" and that most data can be related to a map or other easily understood graphic. GIS
platforms such as Arc View, Arclnfo, and Maplnfo are frequently used to integrate spatial data with
monitoring data for watershed analysis.
The EPA National Nutrient Criteria Program initiated the development of a national database application
that will be used to store and analyze nutrient data. The ultimate use of these data will be to derive
ecoregion- and waterbody-type specific numeric nutrient criteria. Initially, EPA developed a Microsoft
Access application that was populated with STORET Legacy data, U.S. Geological Survey (USGS)
(NAWQA, NASQAN, and Benchmark) data, and other relevant nutrient data from universities,
States/Tribes, and additional data-rich entities. To serve the general public more effectively and
efficiently, EPA also developed and maintains a web-accessible nutrient database application in an
Oracle™ environment that allows for easy web accessibility, geo-referencing/GIS compatibility, and data
analysis on a State/Tribal, regional, and national basis. The total amount of existing nutrient data
nationally is large (>20 gigabytes), and it is anticipated that more data will be entered into the system.
The Oracle™ application can easily manage large quantities of data and provides ample room for
expansion as more data are collected. The Oracle™ database application is being designed for
compatibility with EPA's modernized STORET. A key feature of the database design will prevent
duplication of effort for users of STORET and the nutrients database application, especially for data
updating. Considerable efforts are also being made to ensure compatibility with other database systems
(e.g., WQS and RAD) currently being developed by EPA's Office of Water. The Oracle™ application
has been online since the fall of 2000.
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Data Sources
Potential sources of data include water quality monitoring data from Federal, State, Tribal, and local
water quality agencies; university studies; and volunteer monitoring programs. However, the data
sources described in this section do not encompass the full extent of available data sources. The data
available in the nutrient database can be used to identify reference areas to begin development of
potential nutrient criteria. The nutrient data sources for estuaries and coastal waters that will be useful
for developing criteria are discussed below. These data sources contain extensive water quality data,
however, data; collection should not be limited to these sources. Collection of scientifically sound water
quality data from any reliable source is encouraged.
Many of the water quality programs listed here include rivers and streams data or mixed freshwater,
estuarine, and coastal water systems data. The rivers and streams information is included in this
document because it gives relevant data about nutrient loading from fluvial systems, which is important
to estuaries and coastal waters. Generally, in estuaries that have been impaired by nutrients, a database
exists, and in less impaired estuaries, the database is often insufficient for comparisons. Nutrient
loading information from fluvial systems may provide a basis for comparison between systems if they
share important geophysical conditions. Such comparisons would assist in developing trends and
extrapolating where insufficient data exist.
EPA Water Quality Data
EPA has many programs of national scope that focus on collection and analysis of water quality data.
The following information on several of the databases and national programs may be useful to water
quality managers as they compile data for criteria development.
STOrage and RETrieval System (STORET)
STORET is EPA's national database for water quality and biological data. EPA's original STORET
system, called the STORET Legacy Data Center (LDC) and operated continuously since the 1960s, was
historically the largest repository of water quality data in the Nation. This legacy mainframe-based
system was the repository of all data held in EPA's original STORET system as of the end of 1998. This
Legacy STORET ceased to exist in the year 2000. In its place, EPA is supporting a modernized database,
simply called STORET, designed as a replacement for the original STORET System. While STORET
will serve as the major repository for more current data, the nutrient criteria database application will
offer major improvements in database content and capabilities that will enable more detailed data
analysis.
Interested parties may view both databases on the World Wide Web. For the nutrient database,
capabilities exist to produce printed reports and download data files. Queries for data via the web will be
designed for use by the general public and will require no special training or software.
STORET is a compendium of data supplied by Federal, State, and local organizations used to evaluate
environmental conditions in the field. The data in STORET are organized by both geographic location
and data ownership. Every field study site is identified by at least one latitude/longitude and, where
Nutrient Criteria—Estuarine and Coastal Waters 5-3
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appropriate, also by State/Province, county, drainage basin, and stream reach. Monitoring activities
recorded include field measurements, habitat assessments, water and sediment samples, and biological
population surveys. Records cover the complete spectrum of physical properties, concentrations of
substances, and abundance and distribution of species observed during biological monitoring. STORET
is designed for maximum compatibility with commercial software, including GISs such as the ESRI
ArcView package, and statistical packages such as PC SAS. STORET downloaded files import easily
into all standard spreadsheet packages. Further information about STORET may be obtained by e-
mailing STORET@epa.gov, or telephoning toll-free at 1-800-424-9067.
Environmental Monitoring and Assessment Program (EMAP)
EMAP is an EPA research program designed to develop the tools necessary to monitor and assess the
status and trends of national ecological resources (see EMAP Research Strategy on the EMAP website:
www.epa.gov/emap). EMAP's goal is to develop the scientific understanding for translating
environmental monitoring data from multiple spatial and temporal scales into assessments of ecological
condition and forecasts of future risks to the sustainability of the Nation's natural resources. EMAP's
research supports the National Environmental Monitoring Initiative of the Committee on Environment
and Natural Resources (CENR). Data from EMAP can be downloaded directly from the EMAP website.
The EMAP Data Directory contains information on available datasets, including data and metadata
(language that describes the nature and content of data). The status of the Data Directory as well as
composite data and metadata files also are available on the EMAP website. EMAP-estuaries data is one
of several areas addressed by the program. Most of the estuaries data were collected during a summer
index period.
Ecological Data Application System (EDAS)
EDAS is EPA's program-specific counterpart to STORET. EDAS was developed by EPA's Office of
Water to manipulate data obtained from biological monitoring and assessment and to assist States/Tribes
in developing biocriteria. It contains built-in data reduction and recalculation queries that are used in
biological assessment. The EDAS database is designed to enable the user to easily manage, aggregate,
integrate, and analyze data to make informed decisions regarding the condition of a water resource.
Biological assessment and monitoring programs require aggregation of raw biological data (lists and
enumeration of taxa in a sample) into informative indicators. EDAS is designed to facilitate data
analysis, particularly the calculation of biological metrics and indexes. Predesigned queries that
calculate a wide selection of biological metrics are included with EDAS. Future versions of EDAS will
include the capability to upload data to, and download data from, the distributed version of modernized
STORET. EDAS is not a final data warehouse, but it is a program or project-specific customized data
application for manipulating and processing data to meet user requirements. The EDAS application is
currently under development; more information will be available through the EPA website.
Ocean Data Evaluation System (ODES)
ODES is used for storing and analyzing water quality and biological data from marine, estuarine, and
some freshwater environments. The system supports Federal, State, and local decisionmakers associated
with marine monitoring programs and managers and analysts who must meet regulatory objectives
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through the evaluation of marine monitoring information. ODES contains data from the National Estuary
Program, the Great Lakes National Program Office, the Ocean Disposal Program, the 301(h) Sewage
Discharge Program, the National Pollutant Discharge Elimination System Program, and the 403 (c)
Program. Records pertain to water quality, fish abundance, bioaccumulation, benthic infauna, fish
histopathology, bioassay, and sediment physical/chemical characteristics. Users can examine both
spatial and temporal relationships among variables. A quality assurance report describing analytical
methods and procedures for each dataset is stored with each dataset.
Chesapeake Bay Program (CBP)
CBP, a cooperative effort between the Federal Government, the States, the District of Columbia, and
local governments in the Chesapeake Bay watershed, provides funds to the States of Maryland and
Virginia for the routine monitoring of 19 directly measured water quality parameters at 49 stations within
the bay watershed. The Water Quality Monitoring Program began in June 1984 with stations sampled
once each month during the colder late fall and winter months and twice each month in the warmer
months. A refinement in 1995 reduced the number of monitoring cruises to 14 per year. Data are
available on the internet at www.chesapeakebay.net/data/.
National Estuarine Programs (NEPs)
Many NEPs have nutrient and related data that could be used for characterization purposes. Presently,
there is no national repository of NEP data, but, in the development of regional nutrient criteria, the
NEPs may serve as an excellent source for information. Some of these programs have electronic
databases and some hard copy data that could be acquired. EPA is attempting to acquire the available
NEP data and eventually enter them into the National Nutrient Criteria Program database. A list of NEP
estuarine systems can be found online at www.epa.gov/nep.
National Oceanographic and Atmospheric Administration (NOAA)
Water Quality Data in the National Oceanographic Data Center (NODC)
NODC is one of three national environmental data centers operated by NOAA and serves as a national
repository and dissemination facility for global Oceanographic data. Its primary mission is to ensure that
global Oceanographic data collected at great cost are maintained in a permanent archive easily accessible
to the world science community and to other users. NODC holds physical, chemical, and biological
Oceanographic data collected by U.S. Federal agencies, including the Department of Defense (primarily
the U.S. Navy); State and local government agencies; universities and research institutions; and private
industry. NODC does not conduct any data collection programs of its own; it serves solely as a
repository and dissemination facility for data collected by others (see website at www.nodc.noaa.gov).
NODC provides data management support for major ocean science projects such as Tropical
Ocean-Global Atmosphere (TOGA), World Ocean Circulation Experiment (WOCE), and Joint Global
Ocean Flux Study (JGOFS). NODC's global holdings of physical, chemical, and biological
Oceanographic data include substantial amounts of data from coastal ocean areas. For example, the
NODC Oceanographic Profile Database holds primarily coastal data (www.nodc.noaa.gov/cgi-
bin/JOPI/jopi).
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National Estuarine Research Reserve System (NERR)
The NERR Systemwide Monitoring Program was designed to identify and track short-term variability
and long-term changes in the integrity and biodiversity of representative estuarine ecosystems and coastal
watersheds for the purposes of contributing to effective national, regional, and site-specific coastal zone
management. The program has two major goals: (1) to support State-specific nonpoint source pollution
control programs by establishing local networks of continuous water quality monitoring stations in
representative protected estuarine ecosystems; and (2) to develop a nationwide database on baseline
environmental conditions in the NERR system of estuaries. Water quality data collected from phase 1 of
the NERR Systemwide Monitoring Program provides data necessary for site and intersite baseline
studies, trend analysis, and impact assessment. Data are available for each of the participating NERR
systems at http://inlet.geol.sc.edu/cbmoweb/30_minute_data.html.
Rivers and Streams Water Quality Data
Rivers and streams water quality data are potentially useful for estuaries and coastal waters. Because
much of the nutrient load to estuaries comes from rivers and streams, it is critical to define nutrient
concentrations landward of tidal influence and to calculate fluvial-based nutrient loads to estuaries and
potentially to coastal waters. EPA STORET, which was discussed in detail previously, includes data
from rivers and streams from across the Nation. Another comprehensive Federal source of river and
stream water quality data is USGS. USGS maintains databases on water quantity and quality for
waterbodies across the Nation. Many of the data for rivers and streams are available through the
National Water Information System (NWIS). The most convenient method of accessing the local
databases is through the USGS State representative. Every State office can be reached through the USGS
home page on the Internet at URL http://www.usgs.gov/wrd002.html. The USGS data from several
national water quality programs covering large regions offer highly controlled and consistently collected
data that may be particularly useful for nutrient criteria analysis. Two programs, the Hydrologic
Benchmark Network (HBN) and the National Stream Quality Accounting Network (NASQAN), include
routine monitoring of rivers and streams during the past 30 years. The USGS National Water Quality
Assessment (NAWQA) Program is building a third national database of stream quality data collected and
analyzed for more than 50 river basins and aquifer systems across the Nation. More information and data
from each of these studies can be found on the USGS website (www.usgs.gov). For additional data
sources, the Rivers and Streams Nutrient Criteria document presents an extensive list of related
freshwater nutrient-related information (Nutrient Criteria Technical Guidance Manual—Rivers and
Streams, 2000, EPA-822-B-00-002).
USGS San Francisco Bay Program
Since 1968, USGS has sustained a research program to understand how coastal ecosystems function and
how those functions are altered by human disturbances. One component of this program is directed to
following and understanding changes in the water quality of San Francisco Bay. The program includes
regular measurements of water quality along a 145-kilometer transect spanning the length of the entire
estuarine system, from the South Bay to the Sacramento River. The program studies many different
aspects of San Francisco Bay, such as changing land use, hydrology, water currents, nutrients, toxic
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contaminants, geological structure, and biological communities. The results of water quality
measurements, and eventually the full dataset, can be accessed at http://sfbay.wr.usgs.gov/access/index/
wqdata.html.
State/Tribal Monitoring Programs
Most States monitor some estuaries and coastal waters within their borders for algal and nutrient
variables. Data collected by State/Tribal water quality monitoring programs can be used for nutrient
criteria development. These data should be available from the agencies responsible for monitoring.
Sanitation Districts
Massachusetts Water Resources Authority (MWRA)
MWRA has conducted a comprehensive monitoring program in Boston Harbor, Massachusetts Bay, and
Cape Cod Bay from 1992 to the present. The program was established to understand baseline conditions
and monitor the effects of effluent discharges into Boston Harbor and Massachusetts Bay. This
multifaceted monitoring program focuses on water quality, benthic ecosystem health, effluent
characterization, and public health issues related to metal, organic, and microbiological contaminants.
All data are stored in an Oracle™ Relational Database Management System to support the monitoring
program.
New York City-Department of Environmental Protection (NYCDEP)
NYCDEP has conducted extensive monitoring to evaluate viable treatment options for sewage effluents
to mitigate conditions that promote eutrophication. The city's monitoring programs have included point
source, water column, sediment, hydrodynamic and atmospheric studies. All data are stored in the
NYCDEP databases and have been used in the development and application of a System-Wide
Eutrophication Model (SWEM) to enhance the city's ability to evaluate the effectiveness of various
treatment options in mitigating conditions that promote eutrophication.
Southern California
The major southern California dischargers of treated sewage effluents into marine waters have conducted
applied research and monitoring programs for more than 30 years. These dischargers include the cities of
San Diego and Los Angeles and the counties of Ventura/Oxnard, Los Angeles, Orange, and San Diego.
The programs are designed to monitor the concentrations and mass emission rates of effluent materials in
the treated effluent; the transport and fate of these materials in the receiving waters; the exposure of the
contaminants to organisms in the receiving waters; and the effects of that exposure to individuals,
populations, and communities of subtidal, intertidal, and water column organisms. Some of this
monitoring is performed to comply with NPDES monitoring efforts, and other monitoring addresses
specific issues of interest to the districts. The data are retained on a number of local databases, but they
are also maintained on EPA's ODES. In addition to the localized databases managed by the sanitation
districts, a research organization (Southern California Coastal Water Research Project) has performed
parallel and specialized monitoring and applied research on the effects of treated sewage effluents in this
region since the early 1970s. Their data are managed onsite and are provided to national data inventories
(e.g., ODES, NODC, STORET).
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California and Oregon
Similar monitoring requirements are established for other locations in California and Oregon. The
districts that have effluents that are discharged into rivers, streams, or oceans are required, through the
NPDES permits, to monitor their treated effluents and the receiving waters. The data are retained
locally, but must also be filed with ODES. The State and Regional Water Resources Control Boards of
California and the Departments of Ecology in Washington and Oregon administer these permits.
Puget Sound, Washington
The cities and counties on the Puget Sound watershed monitor the treated sewage effluents and the
receiving waters in compliance with State water quality parameters. These data are provided to the State
in electronic format and are retained on database systems administered by the Department of Ecology
that are available to ODES. The data in the receiving water environments are collected in methods that
are historically similar to work that has been performed in Puget Sound since the middle 1960s.
Academic and Literature Sources
Many research studies conducted by academic institutions may provide data useful for developing
nutrient criteria. Academic research tends to be site specific and span a limited number of years,
although data for some systems may span 20 years or more. Academic research data should be available
from researchers. However, the scientific literature is likely to be a major source of estuarine and coastal
waters data.
Volunteer Monitoring Programs
Many States have volunteer water quality monitoring programs. Some programs are State sponsored,
while others are independently organized. Citizens in many areas donate their time, money, or
experience to aid State, Tribal, and local governments in collecting water quality data. Volunteers
analyze water samples for dissolved oxygen, nutrients, pH, temperature, and a host of other water
constituents; evaluate the health of stream habitats and aquatic biological communities; note shore zone
conditions and land uses that may affect water quality; catalogue and collect beach debris; and restore
degraded habitats.
State and local agencies may use volunteer data to screen for water quality problems, establish trends in
waters that would otherwise be unmonitored, and make planning decisions. Volunteers benefit from
learning more about their local water resources and identifying what conditions or activities might
contribute to pollution problems. As a result, volunteers frequently work with clubs, environmental
groups, and State/Tribal or local governments to gather information and address problem areas. As with
any other data source, whether student, State, Federal, academic, or volunteer based, documented quality
assurance procedures are an important consideration.
EPA supports volunteer monitoring and local involvement in protecting our water resources. EPA
support takes many forms, including sponsoring national and regional conferences to encourage
information exchange among volunteer groups, government agencies, businesses, and educators;
publishing sampling methods manuals for volunteers; and providing technical assistance (primarily on
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quality control and laboratory methods) and regional coordination through the 10 EPA Regional Offices.
EPA also produces a Nationwide Directory of Volunteer Monitoring Programs, which is available online
at http://yosemite.epa.gov/water/volmon.nsf This directory lists volunteer organizations around the
country engaged in monitoring rivers, lakes, estuaries, beaches, wetlands, and groundwater, as well as
surrounding lands. EPA volunteer monitoring activities are coordinated in part through a website that
lists many resources at http://www.epa.gov/owow/monitoring/volunteer.
Quality of Historical Data
The value of older historical datasets is a recurrent problem because data quality is often unknown.
Knowledge of data quality is also problematic for long-term data repositories such as STORET and long-
term State databases, where objectives, methods, and investigators may have changed many times over
the years. The most reliable data tend to be those collected by a single agency, using the same protocol,
for a limited number of years. Supporting documentation should be examined to determine the
consistency of sampling and analysis protocols. Investigators must determine the acceptability of data
contained in large, heterogeneous data repositories. Considerations and requirements for acceptance of
these data are described below.
Location Data
STORET and USGS data are geo-referenced with latitude, longitude, and up to Reach File 3 (RF1 &
RF3) codes. Geo-reference data can be used to select specific locations or specific USGS hydrologic
units. In addition, STORET often contains a site description. Knowledge of the rationale and methods
of site selection from the original investigators may supply valuable information. Metadata of this type,
when known, are frequently stored within large long-term databases.
Variables and Analytical Methods
Thousands of variables are recorded in database records. Each separate analytical method yields a
unique variable. For example, five ways of measuring total nitrogen (TN) results in five unique
variables. We do not recommend mixing analytical methods in sample analyses because methods differ
in accuracy, precision, and detection limits. Sample analyses should concentrate on a single analytical
method for each parameter of interest. Selection of a particular "best" method may result in too few
observations, in which case it may be more fruitful to select the most frequently used analytical method
in the database. Data may have been recorded using analytical methods under separate synonymous
names, or analytical methods incorrectly entered when data were first added to the database. Review of
recorded data and analytical methods recorded by knowledgeable personnel is necessary to correct these
problems.
Laboratory Quality Control
Laboratory quality control of data (blanks, spikes, replicates, known standards, etc.) where available
should be reported. Such information may have been infrequently reported in larger data repositories and
needs to be identified and coded. Records of general laboratory quality control protocols and specific
quality control procedures associated with specific datasets are valuable in evaluating data quality.
However, premature elimination of lower quality data can be counterproductive because the increase in
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variance caused by analytical laboratory error may be negligible compared with natural variability or
sampling error, especially for nutrients and related water quality parameters. However, data of uncertain
quality should not be accepted unless no other data are available.
Data Collecting Agencies
Selecting data from particular agencies with known, consistent sampling and analytical methods will
reduce variability caused by unknown quality problems. Requesting data review for quality assurance
from the collecting agency will reduce uncertainty about data quality.
Time Period
Long-term records are critically important for establishing trends. Determining if trends exist in the time
series database is also important for characterizing reference conditions for nutrient criteria. Length of
time series data needed for analyzing nutrient data trends is discussed in the Sampling Design section
(Section 5.3).
Index Period
The index period for estimating average concentrations can be established if nutrient and water quality
variables were measured through seasonal cycles. The index period may be the entire year or the
summer season. The best index period is determined by considering water quality characteristics for the
region, the quality and quantity of data available, and estimates of temporal variability (if available).
Additional information and considerations for establishing an index period are discussed in Section 5.3.
Representativeness
Data may have been collected for specific purposes. Data collected for toxicity analyses, effluent limit
determinations, or other pollution problems may not be useful for developing nutrient criteria.
Furthermore, data collected for specific purposes may not be representative of the spatial scale of
interest. The investigator must determine if the spatial scale for the data included in the database is
representative of the area to be characterized. If a sufficient amount of data for the appropriate scale
cannot be found, then new surveys will be necessary (see Section 5.3).
Gathering New Data
New data should be gathered following the sampling design protocols discussed below. New data
collection activities for developing nutrient criteria should focus on filling in gaps where data are
particularly needed for high-risk systems. Data gathered under new monitoring programs should be
imported into databases or spreadsheets and merged with the existing nutrient database for criteria
development.
5.3 SAMPLING DESIGN
This section discusses issues surrounding sampling nutrients, response variables, and related
environmental variables in estuaries and coastal waters. Where appropriate data are unavailable or
insufficient to derive numerical nutrient criteria, efforts must be made to collect new data to fill those
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gaps. New sampling programs should be scientifically based and statistically rigorous while maximizing
available management resources. Such programs are used to better define nutrient and algal relationships
within an ecosystem framework. At the broadest level, sampling efforts should detect or contribute to the
following objectives:
• Identify the reference condition, that is, existing, most natural, least culturally impacted locations
and their relative enrichment status
• Identify whether nutrient concentrations or loads are increasing, decreasing, or remaining the same
• Characterize seasonal patterns in nutrient levels and their relationship to primary productivity
• Help assess the assimilative capacity of the system, that is, contribute to the determination of how
much nutrient loading can be assimilated without causing unacceptable changes in water quality or
algal biomass and composition. (Note: In estuaries and coastal waters, this objective will likely
require application of a computer-based hydrodynamic and nutrient-coupled water quality model.
The intention here is to recognize particularly susceptible waters, not to lower expectations relative
to historical antecedents and the reference condition.)
Some sampling programs may be poorly and inconsistently funded or are improperly designed and
carried out, making it difficult to collect a sufficient number of samples over time and space to identify
changes in water quality or to estimate average conditions with statistical rigor. This section provides a
procedural approach for assessing water quality condition and identifying impairment by nutrients and
algae in estuaries and coastal waters. The approaches described below present sampling designs that
allow one to obtain a significant amount of information while attempting to minimize overall effort (and
cost). Probabilistic and stratified random sampling begin with large-scale random monitoring designs
that are reduced as nutrient and algal conditions are characterized. The tiered approach to sampling
begins with coarse screening and proceeds to more detailed protocols as impaired and high-risk systems
are identified and targeted for further investigation.
Sampling Protocol
Success of nutrient criteria development requires that consideration be given to sampling design.
Initially, the relationship between critical response variables and nutrient concentrations, or in some
cases, nutrient loads, needs to be established. Next, reference sites should be sampled, if feasible, in an
attempt to establish reference conditions within classes of systems or subsystems. Classification should
be linked to the reference condition activity. Nutrient concentrations/load and algal biomass
relationships should help define the ecological state that can be attained if impaired systems are restored.
As discussed in the following sections, this is not a straightforward exercise; it is very difficult to predict
water transport/mixing in estuaries and coastal waters. The physics of these waters plays a major role in
determining the observed patterns in nutrient and chlorophyll a concentrations, turbidity, and bottom
water dissolved oxygen deficiency as well as transport. Variability in time and space further complicates
empirical analysis as pointed out in Chapter 2.
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Nutrient concentrations, chlorophyll a production, and system respiration represented as biological
oxygen demand (BOD) are biochemical processes. How these processes are expressed in terms of
dissolved oxygen, especially in bottom waters, has a great deal to do with the variability in freshwater
flows, density stratification, advection, and mixing. The forcing factors include wind setups, changes in
barometric pressure gradients, freshwater gravitational circulation, and the added complexity of bottom
bathymetry. Interactions of these factors may create "flow jets," flow reversals, three-layered circulation,
and other physical complexities that suggest any monitoring scheme planned for estuaries and coastal
waters would be advised to have a physical oceanographer as part of the team.
Sampling Technique
A reasonable and representative method is to profile the general physical character of the site by a CTD
hydrocast. Water samples then may be collected from the surface 1 meter, mid-depth, and bottom 1
meter of the water column. Sample station activities should be coordinated as much as possible with the
same tidal and current phase each time data are collected. If turbidity is measured by Secchi depth, the
disc should be lowered from the shaded side of the vessel and depth determined from as close to the
water level as practicable. Secchi depth measurements should be made only during periods of full
daylight. The data for each station and sampling event should be recorded for each depth interval. This
permits assessment for surface as well as bottom conditions. Where satisfactory, the results for each
sampling episode can be combined into a mean or median measure representing all depths at that site.
Temporal and spatial medians of the sites then can be determined to establish the representative values
for that reference site.
During sampling visits, the candidate reference stations also should be examined to confirm whether they
actually meet the reference site requirements. This may include looking for nearby discharges into the
waters or tributaries and a quick survey of the shoreline to determine if new modifications may have
changed the site. If an area appears to have been significantly impacted, measurements should be made
for nutrient concentrations and biological response variables such as chlorophyll concentration and fish,
macroinvertebrate, macrophyte, and planktonic community variables. Sites that do not meet the physical
reference requirements should be excluded from the reference dataset. However, a high nutrient
concentration present in an otherwise minimally developed area is not justification alone for exclusion.
This may be part of the natural background level to be identified by the reference condition process.
Initial Considerations
Variability is inherent in sampling, which means that accuracy (how well the measure reflects actual
conditions) and precision (how consistent the measurement is) must be assessed. Precision in ecological
samples and measurements is more easily characterized than accuracy. Replicate samples from an
experimental unit provide the basis for precision analysis. Standard statistical textbooks focus on
precision (i.e., various ways to assess the nature of variability) and especially inferences regarding null
hypotheses. In analytical chemical analyses, accuracy can be assessed by including samples of known
purity and/or amount. Accuracy often refers to systematic errors in a method, whereas precision refers
more or less to random errors. Outliers can be detected with statistical methods, but the so-called outlier
may actually prove to be more accurate than the remainder of the data.
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Some key questions to consider are the following: Does the method reflect more on how the assessor or
analyst uses the method if the technician is imprecise, and Does the method actually do what it is
purported to do? For example, is the arsenic concentration high enough that it interferes importantly
with the phosphate analysis? Does the Loran system accurately place one on station, especially when
electronic interferences are highly probable? Does the temperature setting on the dissolved organic
carbon analyzer allow for accurate measurement or a biased one, or does the chlorophyll a method
accurately correct for interferences from other pigments? For these examples, the precision may be very
high (low variance) but the measurement may be in substantial error so the results are inaccurate. The
following discussion expands on these ideas and provides additional information on statistical concepts
and procedures relative to sample design. It is worth remembering that environmental data may not
conform to the assumptions of normality required for statistical inferences; adjustments often can satisfy
parametric assumption, but when they do not, the analysts must resort to distribution-free methods.
Specifying the Population and Sample Unit
Sampling is statistically expressed as a sample from a population of objects. Finite populations may be
sampled with corresponding natural sample units, but often the sample unit (say, an estuary) is too large
to measure in its entirety, and it must be characterized with one or more second-stage samples of the
sampling gear (bottles, benthic grabs, quadrats, etc.). Each sample unit is assumed to be independent of
other sample units. The objective of sampling is to best characterize individual sample units in order to
estimate some attributes (e.g., nutrient concentrations or dissolved oxygen) and their statistical
parameters (e.g., mean, median, variance and percentiles) of a population of sample units. The objective
of the analysis is to be able to say something (estimate) about the population. It is critical to distinguish
between making an inference about a population of many estuaries (e.g., "lagoonal estuaries around
northern Gulf of Mexico are shallow and mesotrophic") versus an inference about a single estuary or
coastal water (e.g., "estuary ABC has fewer fish species than unimpaired reference estuaries or salinity
zones within estuaries"). These two kinds of inferences require different sampling designs: the first
requires independent observations of many waterbodies and does not require repeated observations
within sample units (pseudoreplication; Hurlbert 1984), while the second often does require repeated
observations within a waterbody. Examples of sample units include:
• A point in an estuary or coastal water (may be characterized by single or multiple sample device
deployments). The population then would be all points in the waterbody, an infinite population.
• A constant area (e.g., square meter, hectare). The population could be all square meters of a coastal
water surface area in a State or region.
• An estuary or a definable subbasin or salinity zone of an estuary as a single unit. Because salinity
often specifies population distributions in estuaries, these zones most often are discrete
environments, at least in the short term, and this is likely to be the most common sample unit. The
population would be all salinity zones in a State or region, a finite population.
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Specifying the Reporting Unit
It is also necessary to specify the units for which results will be reported. Usually, this is the population
(e.g., all estuaries), but it also can be subpopulations (e.g., estuaries within a given nutrient ecoregion)
and even individual locations (e.g., estuaries or coastal waters of special interest). To help develop the
sampling plan, it is useful to create hypothetical statements of results in the way that they will be
reported, for example:
• Status of a place: "The estuary ABC is degraded."
• Status of a region: "An estimated 20% of the estuary area in State XYZ has an elevated trophic
state, above reference expectations"; "Approximately 20% of estuaries in State XYZ have an
elevated trophic state."
• Trends at a place: "Nutrient concentrations in estuary ABC have decreased by 20% since 1980."
• Trends of a region: "Average estuary trophic state in State XYZ has increased by 20% since
1980"; "Average trophic state index values in 20% of estuaries of State XYZ have increased by
15% or more since 1980."
• Relationships among variables: "A 50% increase of N loading above natural background is
associated with decline in taxa richness of benthic macroinvertebrates, below reference
expectations."; "Coastal waters receiving runoff from large nonpoint sources have 50% greater
probability of elevated trophic state above reference conditions than coastal waters not receiving
such runoff."
Sources of Variability
Variability of measurements has many possible sources, and the intent of many sampling designs is to
minimize the variability due to uncontrolled or random effects, and conversely to be able to characterize
the variability caused by experimental or class effects. For example, estuaries may be classified by soil
phosporus content of their surrounding watersheds so that estuaries within a class are likely to have
similar water column concentrations in current or historical reference areas. The population of estuaries
is stratified so that observations (sample units) from the same stratum will be more similar to each other
than to sample units in other strata.
Environmental measures vary across different scales of space and time, and sampling design must
consider the scales of variation. In coastal waters, measurements of some variables such as total nitrogen
or chlorophyll concentrations are taken at single points in space and time (center of the deep depression,
20 m depth, 10 a.m. on 2 July). If the same measurement is taken at a different place (littoral zone, 1 m),
or coastal waters, or time (30 January), the measured value may be different. A third component of
variability is the ability to accurately measure the quantity of interest, which can be affected by sampling
gear, instrumentation, errors in proper adherence to field and laboratory protocols, and the choice of
methods used in making determinations.
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The basic rule of efficient sampling and measurement is to sample so as to minimize measurement errors,
to maximize the components of variability that have influence on the central questions and reporting
units, and to control other sources of variability that are not of interest, that is, to minimize their effects
on the observations. In the example of chlorophyll concentrations, variability could be reduced by
sampling each of several coastal waters in the deepest part, with multiple depth samples or a vertically
integrated pump sample taken in early spring before stratification appears. Many coastal waters are
sampled to examine and characterize the variability due to different coastal waters (the sampling unit).
Each coastal water is sampled in the same way, in the same place, and in the same timeframe to minimize
variability due to location, depth, and season, which are not of interest in this particular study.
In the above example, chlorophyll concentrations vary with location within a coastal water, among
coastal waters, and time of sampling (day, season, year). If the spatial and temporal components of
variability within coastal waters are large (e.g., measurements of chlorophyll concentrations typically
vary more between spring and fall samples within a coastal water than they do among coastal waters),
then it may be best to use an index period. For this reason, coastal water chlorophyll concentrations
often are estimated as a growing season average, estimated from several determinations (e.g., monthly)
during the growing season.
In statistical terminology, there is a distinction between sampling error and measurement error that has
little to do with actual errors in measurement. Sampling error is the error attributable to selecting a
certain sample unit (e.g., a coastal water or a location within a coastal water) that may not be
representative of the population of sample units. Statistical measurement error is the ability of the
investigator to accurately characterize the sampling unit. Thus, measurement error includes components
of natural spatial and temporal variability within the sample unit as well as actual errors of omission or
commission by the investigator. Measurement error is minimized with methodological standardization:
selection of cost-effective, low-variability sampling methods; proper training of personnel; and quality
assurance procedures to minimize methodological errors. In analytical laboratory procedures,
measurement error is estimated by replicate determinations on some subset of samples (but not
necessarily all). Similarly, in field investigations, some subset of sample units should be measured more
than once to estimate measurement error.
Analysis of variance (ANOVA) can be used to estimate measurement error. All multiple observations of
a variable are used (from all coastal waters with multiple observations), and coastal waters are the
primary effect variable. The root means square error (RMSE) of the ANOVA is the estimated variance
of repeated observations within coastal waters. Note that a hypothesis test (F-test) is not of interest in
this application, only the RMSE of the analysis.
Natural variability that is not of interest for the questions being asked, but which may affect the ability to
address them, should be estimated with the RMSE method above. If the variance estimated from RMSE
is unacceptably large (i.e., as large or larger than variance expected among sample units), then it is often
necessary to alter the sampling protocol, usually by increasing sampling effort in some way to further
reduce the measurement error. Measurement error can be reduced by multiple observations at each
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sample unit, for example, multiple ponar casts at each sampling event; multiple observations in time
during a growing season or index period; depth-integrated samples; or spatially integrated samples.
A less costly alternative to multiple measures in space is spatially composite determinations. In nutrient
or chlorophyll determinations, a water column pumped sample, where the pump hose is lowered through
the water column, is an example of a spatially composite determination. Spatial integration of an
observation and compositing the material into a single sample is almost always more cost-effective than
retaining separate, multiple observations. This is especially so for relatively costly laboratory analyses
such as organic contaminants and benthic macroinvertebrates, but the price of this economy is loss of
information about the water column or about distribution over an area.
Statistical power is the ability of a given hypothesis test to detect an effect that actually exists and must
be considered when designing a sampling program (e.g., Peterman 1990, Fairweather 1991). The power
of a test (1-b) is defined as the probability of correctly rejecting the null hypothesis when the null
hypothesis is false (i.e., the probability of correctly finding a difference [impairment] when one exists).
For a fixed confidence level (e.g., 90%), power can be increased by increasing the sample size or the
number of replicates, except in cases where the variance is proportional to the mean. To evaluate power
and determine sampling effort, an ecologically meaningful amount of change in a variable must be set.
Optimizing sampling design requires consideration of tradeoffs among the measures used, the effect size
that is considered meaningful, desired power, desired confidence, and resources available for the
sampling program. Every study requires some level of repeated measurement of sampling units to
estimate precision and measurement error. Repeated measurement at 10% of sites is common among
many monitoring programs.
Alternative Sampling Designs
Sampling design is the selection of a part of a population to observe the attributes of interest to estimate
the values of those attributes for the whole population. Classical sampling design makes assumptions
about the variables of interest; in particular, it assumes that the values are fixed (but unknown) for each
member of the population until that member is observed (Thompson 1992). This assumption is perfectly
reasonable for some variables, say, length, weight, and sex of members of an animal population, but it
seems less reasonable for more dynamic variables such as nutrient concentrations, loadings, or
chlorophyll concentrations of estuaries. Designs that assume that the observed variables are themselves
random variables are model-based designs, where prior knowledge or assumptions are used to select
sample units.
Probability-Based Designs (Random Sampling)
The most basic probability-based design is simple random sampling, where all possible sample units in
the population have the same probability of being selected; that is, all possible combinations of n sample
units have equal probability of selection from among the N units in the population. If the population N is
finite and not excessively large, a list can be made of the N units, and a sample of n units is randomly
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selected from the list. This is termed list frame sampling. If the population is very large or infinite (such
as locations in an estuary), one can select a set of n random (x,y) coordinates for the sample.
All sample combinations are equally likely in simple random sampling, thus there is no assurance that the
sample actually selected will be representative of the population. Other unbiased sampling designs that
attempt to acquire a more representative sample include stratified, systematic, multistage, and adaptive
designs. In stratified sampling, the population is subdivided or partitioned into strata, and each stratum is
sampled separately. Partitioning is typically done so as to make each stratum more homogeneous than
the overall population; for example, estuaries could be stratified on ecoregion or coastal waters by
dominant current structure. Systematic sampling is the systematic selection of every k* unit of the
population from one or more randomly selected starting units, and it ensures that samples are not
clumped in one region of the sample space. Multistage sampling requires selection of a sample of
primary units, such as fields or hydrologic units, and then selection of secondary sample units, such as
plots or estuaries within each primary unit in the first-stage sample.
Estimation of statistical parameters requires weighting of the data with inclusion probabilities (the
probability that a given unit of the population will be in the sample) specified in the sampling design. In
simple random sampling, inclusion probabilities are by definition equal, and no corrections are
necessary. Stratified sampling requires weighting by the inclusion probabilities of each stratum.
Unbiased estimators have been developed for specific sampling designs and can be found in sampling
textbooks, such as Thompson (1992).
Model or Goal-Based Designs
Use of probability-based sampling designs may miss relationships among variables (models), especially
if there is a regression-type relationship between an explanatory and a response variable. As an example,
elucidation of estuary response to N loading with the Vollenweider-type model; that is, chlorophyll a
concentration regressed against a depth-normalized N concentration (Vollenweider 1968) requires a
range of trophic states from ultra-oligotrophic to hypereutrophic. A simple random sample of estuaries is
not likely to capture the entire range (i.e., there would be a large cluster of "mesotrophic" estuaries with
few at high or low ends of the trophic scale), and the random sample therefore may be biased with
respect to the model.
In model-based designs, sites are selected based on prior knowledge of auxiliary variables, such as
estimated phosphorus loading, estuary depth, and elevation. Often, these designs preclude an unbiased
estimate of the population response variable (e.g., trophic state), unless the model can be demonstrated to
be robust and predictive, in which case the population value is predicted from the model and from prior
knowledge of the auxiliary (predictive) variables. Selection of unimpacted reference sites is an example
of samples for a model (index development; response of index variables to measures of anthropogenic
influence) that cannot later be used for unbiased estimation of the biological status of estuaries. Ideally, it
may be possible to specify a design that allows unbiased estimation of both population and model.
Statisticians should be consulted in developing the sample design for a nutrient criteria program.
Nutrient Criteria—Estuarine and Coastal Waters 5-17
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Sampling and Analytical Designs for More Complex Ecological Questions
Complex ecological questions may not be required to develop numerical nutrient criteria. However, the
manager may require a biological and an ecological assessment of resources at risk to establish
management goals—that is, focus on biological resources of high social and economic value (e.g., the
Chesapeake Bay Program includes biological variables as part of the goal setting process). Questions on
how to sample different levels of biological organization (e.g., populations, communities, and
ecosystems), indicators of stress, diversity and similarity measures, and biotic indices may become
important. Criteria that go beyond the core variables will likely address one or more of these ecosystem
or community elements. Publications are available that provide conceptual and statistical guidance for
monitoring biological/ecological systems, including multivariate analytical approaches (e.g., Spellerberg
1991, Luepke 1979, Digby 1987, Clark and Warwick 1994a,b, Ott 1995, Ludwig and Reynolds 1988,
Eckblad 1991).
Monitoring Programs
The purpose of monitoring is to obtain data that can be used not only to determine reference condition,
but to help classify estuaries and coastal waters, or portions thereof, into groups (see Chapter 3).
Classification should aid in the determination of reference sites or stations that are representative and
have the lowest possible variability.
In some cases, a problem may exist where monitoring data indicate that the system has been greatly
impaired from nutrient enrichment over the period of record. This is analogous to the so-called "corn-
belt" problem in some lakes in the upper mid-West of the United States (see Lakes and Reservoirs
Nutrient Guidance Document). This problem suggests that a meaningful reference condition may no
longer exist. Or, the system has been greatly disturbed and it is not clear to what extent the impairments
are due to nutrient enrichment. In this case, both historical information and diagnostic sampling may be
required to clarify the reference condition and subsequent nutrient criteria.
Where data are insufficient, several approaches can be tried, for example, running a mathematical
nutrient model "backwards"; use of biostratigraphic approaches, including changes in algal dominance
and composition, loss of submerged aquatic vegetation (SAV), and pyritization of iron in sediments to
detect earlier anoxia (Brush 1984, 1986, 1992, Cooper 1995) or reference to old written accounts (e.g.,
newspapers, diaries). For example, there are accounts of water clarity in the mouth of the Patuxent River
estuary, Chesapeake Bay, in the late 1930's where engineers sitting in a "Beebe-like Bathysphere" on the
estuary bottom could see horizontally approximately 20 to 30 feet. The methods discussed here are often
qualitative to semiquantitative, but such information can be useful, especially if marked nutrient
increases have evidently occurred over historical conditions but ambient data are insufficient. Older
aerial photographs and other forms of watershed land use information and human population density
trends can help make extrapolations regarding the system's response to nutrient loading.
Parameters to Survey
Each of the core variables discussed in Chapter 4 must be included in the survey (e.g., concentrations of
TN, TP [total phosphours], chlorophyll a, and a measure of water clarity, such as Secchi disc,
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submersible PAR meter, or spectral radiometer). It is also appropriate to measure salinity, water
temperature, flow and direction, tide phase, pH, and nutrient load to help better interpret the core
variables. This is a much different problem than usually experienced in rivers and lakes. It may be
desirable in some circumstances to include secondary variables, for example, vertical dissolved oxygen
profiles, distribution and abundance of SAV/seagrasses, distribution of tidal emergent marshes,
distribution and density of benthic filter feeders (e.g., oysters), water color, dissolved organic carbon
(especially if humic-like materials are abundant), and particulate organic carbon. This more complex
array of variables would require a diagnostic justification.
Sampling Frequency
A single grab sample from an estuary or coastal water will be grossly inadequate. Estuaries are near the
bottom of watersheds, which makes them prone to episodic rainfall events. Coastal waters are also
subject to seasonal storms that churn the waters and physically disturb shallow sediments, and these
events may be seasonally highly variable. If information is available to set expectations when possible
seasonal pulses of freshwater occur, then it should be used to help schedule the sampling of wet and dry
periods. In north temperate estuaries, where winter to early spring is the dominant freshet period, this
interval should be included in the sampling scheme. A lag of hours, days, or several weeks to one or
more months is usually required to detect the system's response to the nutrient load, depending on
magnitude of the freshet relative to volume of the estuary or mixing zones of coastal waters. This also
may capture any spring blooms of diatoms if such occur. A midsummer and early fall survey should give
a first-order picture of the nutrient concentration and response variable pattern suitable for classification.
In the event of variable summer freshwater flows, then more frequent sampling may be required.
Because different patterns of rainfall exist around the coasts, regional considerations should weigh
heavily in the design of sampling schedules.
Long-term datasets have well-documented ecological value (Likens 1992, Wolfe et al. 1987, Livingston
200 la); however, all too frequently resources constrain longer term sampling which can average out
short-term variability. Recent data connected to long-term trends provide the strongest case for
classification, reference condition determination, and other criteria development. By measuring the
nutrient load, especially during freshet and low flow periods and concurrently with ambient water quality
and hydrographic sampling, one can get an estimate of the load and salinity/nutrient and response
variable relationships while keeping in mind the precautions noted above. For comparative purposes, it
is important to compare core monitoring variables under similar salinity conditions.
If tidal elevation is large (e.g., greater than 2.0 m), then this component of estuarine flushing probably
dominates over nontidal gravitational flows (Monbet 1992), and eutrophication symptoms are likely to be
of a small magnitude. In some estuaries (e.g., York River estuary of the Chesapeake Bay), spring tides
may break down density stratification, and the system responds differently to nutrient supplies than
during periods of relatively strong stratification (Hass 1977). Thus, for estuaries with tidal elevations
less than 2.0 m, it is important to note that they are likely to be quite vulnerable to nutrient enrichment.
Nutrient Criteria—Estuarine and Coastal Waters 5-19
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A general rule of thumb regarding freshwater run-off events to estuaries is that a large freshet may
displace the nutrient supply and responses will be detected seaward of the focal area. A modest freshet
may not deliver enough nutrients or physically affect the density stratification to make an estuary
vulnerable to nutrient enrichment. But an intermediate freshet may cause the focal area to receive a
significant nutrient load and establish a strong vertical density gradient so maximum responses will be
detected (e.g., high average chlorophyll a concentrations and minimum Secchi disk readings). This rule
is less easily applied to coastal waters.
Sampling Locations
Sampling locations depend on the size (and especially the length of an estuary), bathymetry, nutrient
source inputs, and hydrography (especially the longitudinal and vertical salinity profiles). In estuaries,
consideration should be given to tidal freshwater, the turbidity maximum (if one is present), mesohaline
and polyhaline regimes, as well as water below zones of density stratification.
In large tidal freshwater riverine systems, it is important to employ several stations because this portion
of the estuary may "store" a large supply of nutrients that later advect into the saline reach of the estuary
(e.g., the Hudson River system) (Lampman et al. 1999). Enough samples should be taken to detect
nutrient concentration gradients along the salinity gradient from tidal river to the estuary receiving
waters. Typically, this will require from five to seven stations at a minimum. If the estuary is relatively
wide (e.g., lagoonal systems such as Pamlico Sound, NC, or Pensacola Bay, FL) or has large tributary
creeks, then these features may need independent sampling. Where salinity gradients are distinct both
horizontally and vertically, composite sampling may have severe limitations. Depth variability also
should be considered, for example, main channel, shelf samples, and samples in shallow water near
SAV/seagrass meadows or in emergent marsh channels should be included. Emergent marsh creeks
should be sampled in the summer during high and low tides, because high system respiration may cause
hypoxia/anoxia in these tidal creeks that may be largely natural. Where SAV meadows are poorly
developed, resuspension of bottom sediments may be more common and not represented by open channel
samples.
Serious consideration should be given to some replicate sampling within salinity zones to estimate
variability; however, resources may require a broad picture where gradients become equally or more
important than the physical salinity "zones." In most cases, analytical levels of detection should be a
trivial aspect of data acquisition for reference characterization. This does not free one from application
of good laboratory quality asurance/quality control (QA/QC) practices, which must be maintained with
appropriate blanks, reference samples, and other considerations to standard analytical measurements.
Citizen Monitoring Programs
Citizen monitoring programs have greatly increased, especially since the early 1980's. Where there is
adequate technical oversight either from within the group expertise or from the outside, such monitoring
efforts can play an important role in assessing trends, identifying "hotspots," and locating likely sources
of nutrients, especially in smaller estuaries where larger research vessels are not required. Many Federal,
State, Tribal, and local agencies assist citizen monitoring efforts, and these agencies contribute to
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training and direction, development, and implementation of QA/QC procedures, act as a data repository;
and perform analyses on environmental samples collected by citizen groups. Citizen monitoring groups
often can provide more frequent observations, such as visiting a gauging station, than can State
personnel. Citizens also can identify those property holders or resource users not following best
management practices or operating within permit limits. See also Volunteer Monitoring Programs,
above.
5.4 QUALITY ASSURANCE/QUALITY CONTROL
The validity and usefulness of data depend on the care with which they were collected, analyzed, and
documented. EPA provides guidance on data QA/QC (U.S. EPA 1998b) to assure the quality of data.
Factors that should be addressed in a QA/QC plan are briefly described below, but the reader is referred
to published EPA guidance for specifics. The QA/QC plan should state specific goals for each factor and
should describe the methods and protocols used to achieve the goals. The five factors discussed below
are representativeness, completeness, comparability, accuracy, and precision.
Representativeness
Sampling program design (when, where, and how you sample) should produce samples that are
representative or typical of the environment being described. Sampling designs for developing nutrient
criteria are discussed earlier in this chapter.
Completeness
Datasets are often incomplete because of spilled samples, faulty equipment, and/or lost field notebooks.
A QA/QC plan should describe how complete the dataset must be to answer the questions posed (with a
statistical test of given power and confidence) and the precautions being taken to ensure that
completeness. Data collection procedures should document the extent to which these conditions have
been met. Incomplete datasets may not invalidate the collected data, but they may reduce the rigor of
statistical analyses. Therefore, precautions should be taken to ensure data completeness. These
precautions may include collecting extra samples, having backup equipment in the field, installing alarms
on freezers, copying field notebooks after each trip, and/or maintaining duplicate sets of data in two
locations.
Comparability
To compare data collected under different sampling programs or by different agencies, sampling
protocols and analytical methods must demonstrate comparable data. The most efficient way to produce
comparable data is to use sampling designs and analytical methods that are widely used and accepted
such as Standard Methods for the Examination of Water and Wastewater (APHA, AWWA, WEF, 1998)
and EPA methods manuals.
Nutrient Criteria—Estuarine and Coastal Waters 5-21
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Accuracy
To assess the accuracy of field instruments and analytical equipment, a standard (a sample with a known
value) must be analyzed and the measurement error or bias determined. Internal standards should
periodically be checked with external standards provided by acknowledged sources. At Federal, State,
Tribal, and local government levels, the National Institute of Standards and Technology (NIST) provides
advisory and research services to all agencies by developing, producing, and distributing standard
reference materials. For calibration services and standards see:
http://ts.nist.gov/ts/htdocs/230/233/home/calibration.html.
Standards and methods of calibration are typically included with CTD sondes, turbidity meters, pH
meters, DO meters, and DO testing kits. USGS, EPA, and some private companies provide reference
standards or QC samples for nutrients. Reference standards for chlorophyll are also available from EPA
and some private companies, although chlorophyll standards are time and temperature sensitive because
they degrade overtime.
Variability
Natural variability, rather than imprecision in the method used, is usually the greatest source of error in
the constituent measured. The variability in field measurements and analytical methods should be
demonstrated and documented to identify the source of variability when possible. EPA QA/QC guidance
provides an explanation and protocols for measuring sampling variability (U.S. EPA 1998). Methods for
creating a chlorophyll standard to determine if the spectrophotometer is measuring chlorophyll
consistently from one year to the next or from the beginning to the end of an analytical run are described
in Wetzel and Likens (1991). In addition, replicates for each sample time and site (usually three) must be
collected because the largest source of variation is likely to be natural (i.e., in the samples).
5.5 STATISTICAL ANALYSES
Statistical analyses are used to identify variability in data and to elucidate relationships among sampling
parameters. Several statistical approaches for analyzing data are mentioned here. We advocate simple
descriptive statistics for initial data analyses, that is, calculating the mean, median, mode, ranges, and
standard deviation for each parameter in the system of interest. The National Nutrients Database
discussed above calculates simple descriptive statistics for queried data. Specific recommendations for
setting criteria using frequency distributions are discussed in Chapter 7.
Data Reduction
Data reduction requires a clear idea of the analysis that will be performed and a clear definition of the
sample unit for the analysis. For example, a sample unit might be defined as "an estuary during
July-August." For each variable measured, a median value then would be estimated for each estuary in
each July-August index period on record. Analyses then are done with the observations (estimated
medians) for each sample unit, not with the raw data. Steps in reducing the data include:
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• Selecting the long-term time period for analysis
• Selecting an index period
• Selecting relevant chemical species
• Identifying the quality of analytical methods
• Identifying the quality of the data recorded
• Estimating values for analysis (mean, median, minimum, maximum) based on the reduction
selected.
Frequency Distributions
Frequency distributions can be used to aid in the setting of criteria. Frequency distributions do not
require prior knowledge of individual waterbody conditions before setting criteria. Criteria are based on
and, in a sense, developed relative to the population of systems in the Region, State, or Tribal
jurisdiction.
Data plotted on a scale of mean nutrient concentration versus frequency of occurrence for a specific
estuary, portion of an estuary, or coastal reach produces a frequency distribution of mean or median
nutrient concentration. Plots of frequency distributions of median TP, median TN, median chlorophyll a,
and Secchi depth for the index period (discussed in Chapter 4) should be examined to determine the
normality of the data in the distribution and to determine the potential for further subdivision of the
waterbody under investigation. Data that are not normally distributed often are transformed into a
distribution more approximating the normal distribution by taking the logarithm of each value. Analysis
of outliers may assist in explaining variability in small datasets; additional analysis can be conducted to
identify the statistical significance of population differences.
Correlation and Regression Analyses
The relationship between two variables may be of use in analyzing data for criteria derivation.
Correlation and regression analyses allow the relationship to be defined in statistical terms. A correlation
coefficient, usually identified as r, can be calculated to quantitatively express the relationship between
two variables. The appropriate correlation coefficient is dependent on the scale of measurement in which
each variable is expressed (whether the distribution of data is continuous or discrete) and whether there is
a linear or nonlinear relationship. Results of correlation analyses may be represented by indicating the
correlation coefficient and represented graphically as a scatter diagram that plots all of the collected data,
not just a measure of central tendency. The statistical significance of a calculated correlation coefficient
can be determined with the t test. The t test is used to determine if there is a true relationship between
two variables. Therefore, the null hypothesis states that there is no correlation between the data variables
measured within the population. A critical a value is chosen as a criterion for determining whether to
reject the null hypothesis. If the null hypothesis is rejected, the alternate hypothesis states that the
correlation at the calculated r value between the two variables is significant.
Regression analysis provides a means of defining a mathematical relationship between two variables that
permits prediction of one variable if the value of the other variable is known. In contrast to correlation
analysis, there should be a true independent variable (a variable under the control of the experimenter) in
Nutrient Criteria—Estuarine and Coastal Waters 5-23
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regression analysis. Regression analysis establishes a relationship between two variables that allows
prediction of the dependent variable (predicted variable) for a given value of an independent variable
(predictor variable). However, scientists (other than statisticians) apply regression analyses to field data
when a relationship is known to exist, even when there is no true independent variable (e.g., cell counts
of algae and chlorophyll concentration; nutrient concentrations and chlorophyll concentration) (Ott 1988,
Campbell 1989, Atlas and Bartha 1993, Ott 1995).
Tests of Significance
Various statistical tests are used to assess the hypotheses being tested. Statistical tests of significance
differ in their applicability to the dataset of interest and the power of the test (the ability of the test to
detect a false null hypothesis). A parametric test of significance assumes a normal distribution of the
population. Nonparametric analyses are valid for any type of distribution (normal, log-normal, etc.) and
can be used if the data distribution is not normal or unknown. A parametric test has more power than a
nonparametric test when its assumptions are satisfied. Two types of errors can be made when testing
hypotheses: Type I—where a correct null hypothesis is mistakenly rejected, and Type II—when there is
a failure to reject a false null hypothesis. The parametric test is less likely than a nonparametric test to
make a Type II error, when the assumptions are met. Therefore, if given a choice, the parametric test
should be used rather than the nonparametric test when the assumptions of the parametric test are
fulfilled. Less powerful, nonparametric tests of significance must be used in cases where the data do not
fit the assumption of a normal distribution (Ott 1988, Campbell 1989, Atlas and Bartha 1993).
Parametric tests include the student ^test, analysis of variance, multivariate analysis of variance, and
multiple range tests. Nonparametric tests include chi square, Mann Whitney U test, and the
Kruskal-Wallis test (Ott 1988; Campbell 1989; Atlas and Bartha 1993). Detailed descriptions of these
and other relevant statistical tests can be found in standard statistical texts.
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CHAPTER 6
Determining the
Introduction and Definition
Significance of Reference Conditions
Paucity of Similar Estuarine and Coastal Marine Ecosystems
. . Approaches for Establishing Reference Conditions
Reference Condition
6.1 INTRODUCTION AND DEFINITION
A reference condition is the comprehensive representation of data from several similar, minimally
impacted, "natural" sites on a waterbody or from within a similar class of waterbodies, i.e., median
values of TN, TP, chlorophyll a, or Secchi depth. However, in cases where severe degradation has
occurred, surrogate values for reference site data may be required as described in this chapter. There are
two basic approaches for their determination: (1) analysis of in situ estuarine and coastal data, and (2)
analysis of watershed nutrient loading to estuaries and, through advective transport, nutrient loading to
the coastal environment. These approaches reinforce each other, but one may be preferred or even
required depending on comparative or site-specific data. Reference conditions are a primary element of
nutrient criteria development, but should be used in conjunction with the other elements described in
Chapter 1 and Chapter 7. Classification of estuaries and coastal waters should facilitate development of
reference conditions but, as pointed out in Chapter 3, further research is essential to bring classification
of these systems to the level of practical utility comparable to that of most freshwater systems. Models
of estuarine susceptibility to nutrient overenrichment are at an early stage of development, and even less
may be known about coastal ecosystems (Chapter 3).
6.2 SIGNIFICANCE OF REFERENCE CONDITIONS
The reference condition is made explicit through several environmental measures. This manual focuses
on TN and TP as principal causative agents, but their relative roles depend on individual
watershed/estuary and conditions. There are two response variables: chlorophyll a, a measure of algal
biomass; water clarity, linked to algal biomass through chlorophyll a; and often a third, dissolved
oxygen deficiency, particularly in estuaries. These explicit measures are indicators of nutrient
enrichment but are linked conceptually to a continuum of biological resources and recreational
opportunities (Figure 6-1). These linkages show considerable variability and apparent elasticity because
ecological processes are not "hardwired" as are some physically engineered systems (e.g., cogs or pulleys
that drive a machine). States and authorized Tribes are encouraged to employ additional response
variables.
EPA assumes that ecosystems will support natural assemblages of aquatic life and high-quality
recreational activities if a nutrient supply is achieved and maintained at a level to support the natural
biological system. This can happen, of course, only if other environmental conditions are compatible.
Nutrient Criteria—Estuarine and Coastal Waters 6-1
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Extensive
Degradation,
Greatly
Overenriched
Designated
Uses Partially
Supported
Low
Water
Quality
Potential EPA Ecoregional
Nutrient Criteria
Reference
Condition
Minimally
Minimum
Water Quality i'mpac[ed
to Support
"Fishable/
Swimmable"
Uses
Pristine
Potential State/Tribal
Nutrient Criteria
High
Water
> Quality
Figure 6-1. Environmental quality scale representing reference conditions and potential nutrient criteria relative to
designated uses.
Conditions that support minimal or unimpaired aquatic resources are typically associated with very low
human population densities and limited land use activities in watersheds that would otherwise be a
source of increased human-mediated nutrient supplies to estuaries and coastal waters. This is a plausible
association because once a watershed is moderately to heavily developed, it is practically impossible to
control all nutrient inputs to the aquatic system. State and national preserves may have relatively high-
quality environmental conditions, but even in these systems atmospheric nitrogen deposition likely has
caused some nutrient enrichment. Ideal conditions may still exist in some estuaries and coastal waters,
and they should be identified. However, where these ideal nutrient conditions no longer exist, especially
in the 50% to 60% of moderately to heavily enriched estuaries (Bricker et al. 1999), the reference
condition should be sought from comparative analyses of similar systems and/or the historical record,
which provides an implied reference condition.
In this manual, a distinction is made for estuaries and coastal waters between "pristine" and minimally
impacted waters. The precolonial period may have had water quality and habitat conditions, including
nutrient loading, that were pristine. This would approximate the ideal of "restoration and protection of
physical chemical and biological integrity" that is the ultimate goal of the Clean Water Act. But this
ideal is largely hypothetical, because methods to estimate it invariably contain a relatively high degree of
uncertainty. However, "fishable and swimmable" conditions are commonly used as an interim goal of
the act and represent a goal of the Nutrient Program, exemplified by reference conditions associated with
minimal human-mediated nutrient enrichment.
6-2
Nutrient Criteria—Estuarine and Coastal Waters
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The term "fishable and swimmable" is not easily quantified in any waters because of their inherent
natural variability. A lake example helps provide a perspective. When a lake shifts from an oligotrophic
to a mesotrophic fishery because of nutrient overenrichment, at some point the change becomes
demonstrable. Likewise, in an estuary when nutrient enrichment shifts phytoplankton production and
algal species composition toward microbial dominance and away from oyster production, the change
becomes demonstrable. What is needed are early warning indicators of the impending change. This is
the role of nutrient criteria. Although it is clear that additional research is called for, there is sufficient
symptomatic knowledge such as changes of species with enrichment to merit setting such indicator
criteria.
Identifying a reference condition in degraded waters should start with an analysis of the best existing
estuarine or marine waters within a watershed or coastal area, or as commonly stated, "the best of what's
left." Because of the difficulties in identifying reference sites in some overenriched estuaries and
nearshore coastal waters, it may be necessary to derive an "implied" reference condition by comparing
the "best of what's left" with "what used to be" as established by the historical record. In any case, it is
still important to identify the best remaining sites in the waterbody of concern. Where the the "best"
sites are known to be severely degraded, more emphasis must be placed on the historical record, but some
knowledge of the continuum from past to present is necessary to establish protective criteria.
Nutrient enrichment-based impacts are a function of the concentrations and supply of nutrients as well as
the ecological conditions and nutrient processes characteristic of the system. Nutrient enrichment effects
may be exacerbated in estuaries where the dominant grazing populations (e.g., menhaden and oysters)
have been lost through human causes, natural causes, or a combination of the two over the past century or
the past several decades. For this reason, it is important to assess the factors that may have modified the
"assimilative capacity" of coastal waterbodies. Reference conditions are not threshold values (a
concentration less than some specified value). For example, in the lower Potomac River estuary,
Chesapeake Bay, the average N load caused a larger phytoplankton bloom than either drought or flood
conditions (Chapter 2). This response involved system hydrodynamics. The point is that reference
conditions should be interpreted in an ecosystem context, especially when the system has experienced
significant nutrient overenrichment and/or is subject to periodic natural disruption such as hurricanes,
winter storms, or droughts.
Some argue that one of the earliest symptoms of impairment involves a nutrient stimulation of harmful
algae relative to beneficial algae. Many types of algal blooms that become a nuisance or harmful clearly
are a form of pollution. There is mounting evidence that many harmful algal blooms are associated with
nutrient enrichment (NRC 2000). For example, substantial loss of seagrass habitat due to algal shading
or bottom habitat due to hypoxia are associated with negative effects on living resources. Seagrasses also
stabilize shorelines and provide cover from predators.
This manual emphasizes the importance of reference conditions to address nutrient problems in a timely
manner. They serve as the best initial measure for identifying nutrient loads that could cause use
impairments. Statistical and computer-based modeling can improve site-specific estimates of the load
Nutrient Criteria—Estuarine and Coastal Waters 6-3
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and response relationships. Classification may assist in extrapolating nutrient effect relationships
between systems in the same class, although classification of coastal waters probably has less
predictability and utility than it does for lakes and streams. As pointed out in Chapter 2, for coastal
waters, the relationships between nutrient loading (e.g., TN and TP) and the response indicators of
chlorophyll a and water clarity normalized to chlorophyll a can be less than straightforward. However,
an understanding of the reference condition will help prevent resource managers from being blindsided
by complications associated with cause-effect relationships.
6.3 PAUCITY OF SIMILAR ESTUARINE AND COASTAL MARINE ECOSYSTEMS
Estuaries and coastal marine ecosystems tend to be relatively individualistic in their sensitivity and
response to nutrient overenrichment. Susceptibility to nutrient enrichment ranks as a premier research
need (Hobbie 2000). The lack of physically similar waterbodies may severely limit grouping
(classifying) waterbodies as recommended for lakes, reservoirs, rivers, and streams where frequency
distributions are used to derive reference conditions (e.g., upper 75th percentile of a priori nutrient-
unimpaired waterbodies or lower 25th percentile of all waterbodies; EPA 2000a,b). As mentioned in
Chapter 1, an exception may be coastal embayments that form behind barrier islands. If several
relatively similar embayments can be identified within a given geographic area, then one or more may
serve as benchmarks against which the others may be compared.
6.4 APPROACHES FOR ESTABLISHING REFERENCE CONDITIONS
Three primary estuarine approaches may provide considerable flexibility to meet the diversity of
conditions encountered by resource managers. A fourth approach focuses on nutrient loading from the
watershed. A fifth approach is described for coastal marine waters. There are situations where light and
estuarine flushing limit the expression of nutrient enrichment effects. In such cases, downstream effects
may nonetheless be a problem requiring attention. Consequently, we have proposed a variety of criteria
development approaches. Reference conditions in any case should be defined with due consideration of
salinity gradients and seasonal and interannual variability. Table 6-1 summarizes the five approaches to
establishing reference conditions in estuaries and coastal waters.
The alternatives for reference condition determination presented in the following text and table represent
two general approaches to developing baseline nutrient quality measurements. The reference conditions
approach per se acknowledges that the system response to overenrichment can be extremely variable, as
described earlier in this manual. For that reason, measuring the nutrient characteristics of minimally
impacted sites provides a reliable nutrient goal regardless of how those nutrients may or may not be
assimilated. This is the approach upon which the National Nutrient Criteria Program is predicated, and
reference sites should always be sought when designing nutrient criteria protocols.
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Table 6-1. Summary of estuarine and coastal nutrient reference condition determinations
Degree of apparent estuarine degradation
Method recommended
Criterion measure
A. In Situ Observations as the Basis for
Estuarine Reference Condition
1. Recognized unique excellent condition
Median ambient concentration. Fig. 6-2.
2. Some degradation, but reference sites exist Upper quartile. Fig. 6-2.
Concentration of TP, TN,
chlorophyll a, Secchi depth
(m)
Same as above
3. Significantly degraded, including all potential
reference sites
B. Watershed-Based Approaches for
Estuarine Reference Condition
4. Same as approach 3 above, but insufficient
historical data
C. Coastal Reference Condition
5. Applicable to all coastal reaches
- Estuarine plumes
- Coastal areas
Intercept value on a regression or Same as above
distribution curve as illustrated in Fig. 6-
3 and 6-4 or by use of a comparable
comparative regression model.
Ref. sites along each trib. and calculate
delivery. Summation is reference
condition. Fig 6-5. Model required to
back-calculate load where all tribs. are
degraded
Load of TP and TN; model
is required to convert load to
estuarine concentration
Index site approach; models may help Concentrations
distinguish anthropogenic contribution
See also Appendix H.
However, many estuaries are so degraded and/or exhibit such short retention times that investigators
cannot determine reference sites with any degree of confidence. In this instance, dose-response curves
and similar approaches as described below may be more appropriately applied in determining historical
reference conditions, i.e., conditions before degradation was first exhibited. Although these approaches
may not provide the real-time affirmation of existing reference sites, their strength is the documentation
of system decline coincident with the overenrichment. The investigator is expected to assess the
characteristics of the given estuary and select the most responsive option for reference condition
determination from those presented.
In all cases, historical information is important either as an alternative reference condition for heavily
enriched systems or as one of the five elements of criteria development essential to providing a status and
trend perspective important to data evaluation before criteria are established.
In Situ Observations as the Basis for Estuarine Reference Condition
These approaches require nutrient and relevant hydrographic data within an estuary.
Nutrient Criteria—Estuarine and Coastal Waters
6-5
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Recognized Unique Excellent Condition
Typically, this condition is based on an extensive spatial and temporal scientific database. If an existing
excellent condition is agreed to by the RTAG and stakeholders, then the State, authorized Tribe, or
appropriate government agency may establish reference conditions based on data that document the
condition and address the remaining elements of criteria development. With limited data, in a very small
number of cases it may be possible to document that the watershed is unimpacted (e.g., has very little
human development, is distant from the influence of local population centers, adjacent land uses are
relatively undisturbed, and is outside of major atmospheric deposition of nitrogen). It is necessary to
document, or augment with additional data, to ensure the original hypothesis is confirmed. Segmenting
the estuary by salinity zones, typically based on estuarine circulation, may be required to reflect nutrient
conditions associated with salinity gradients. However, the geographic scale of the nutrient
overenrichment problem suggests that this a priori approach likely will be limited to a relatively small
number of estuaries and coastal waters (Chapters 1 and 2). The data can be summarized as either
medians of the indicator endpoints or frequency distributions (Figure 6-2). Areas that meet or come very
close to minimally impaired conditions include Plum Island Sound and Blue Hill Bay, ME, and lower
Narragansett Bay, RI.
Some Degradation Exists But Reference Sites Can Be Identified
Two situations may be applicable for characterizing reference conditions where some degradation occurs.
A. Some Minimally Impaired Sites Are Available
Reference sites should be representative of the system (e.g., a branched estuarine system where one
branch is unimpaired by nutrients and otherwise relatively similar ecological conditions prevail).
Comparisons should be made among similar salinity zones. This example may apply only rarely because
in atmospheric nitrogen deposition and land use practices likely will have so altered the landscape that
truly undisturbed conditions are unavailable. In those cases where minimal biological resource uses are
impaired by nutrient overenrichment, then reference conditions for nutrients should be deemed to occur.
Clearly, point and nonpoint source discharges must be at a minimum. Land cover in the watershed
should be very close to natural for the ecoregion or, if modified in the past, then recovery must be well
along (e.g., forests should be near the anticipated climax condition for the region). The reader is referred
to Chapter 4 of "Estuarine and Coastal Marine Waters: Bioassessment and Biocriteria Technical
Guidance," December 2000 (EPA-822-B-00-024), for additional information (U.S. EPA 2000b).
Although the upper estuary likely does not qualify, reference-quality sites probably occur in lower
Delaware Bay, especially under sustained periods of low Delaware River input. Under these conditions,
because the measurement values likely will not exhibit a trend or show high variability, the data can be
summarized in terms of median values or a frequency distribution.
6-6 Nutrient Criteria—Estuarine and Coastal Waters
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0>
MEDIAN
LOWER QUARTILE
(For Mixed Data)
UPPER QUARTILE
(For Reference Quality Data)
CONCENTRATION (TP, TN, CHLOROPHYLL-a)
SECCffl DEPTH
Figure 6-2. Hypothetical frequency distribution of nutrient-related variables showing quantities
for reference or high-quality data and mixed data (all data incuded).
B. Reference Condition Derived From A Priori Selection of a Subset of Reference Estuaries or Coastal
Waters Within a Class of Systems
This approach may apply to a series of coastal embayments located within a physically similar reach of
coastline (e.g., some estuaries along the Maine coast and possibly New England salt ponds). Where
freshwater inputs contribute to a strong salinity gradient, salinity zones may be compared. The goal is to
establish frequency distributions for similar systems. If a population of 15 or more estuaries or
embayments occurs within the class, then a frequency distribution may be applicable (see Chapter 6,
Nutrient Criteria Technical Guidance Manual-Lakes and Reservoirs, April 2000, EPA-822-BOO-001). A
sample size of fewer than 15 systems will not likely have enough predictive power to justify application
of this approach. However, if only a small number exists, then one or two embayments may serve
qualitatively or possibly semiquantitatively as references for the remaining systems. Seasonality and
similar freshwater inputs are important to ensure that physical processes are not masking potential
nutrient effects. To date, it is unknown whether there are enough reference embayments to make a
frequency distribution approach feasible. Other examples may be difficult to identify but tributaries to
North Inlet, SC, may qualify.
Nutrient Criteria—Estuarine and Coastal Waters
6-7
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Significant Degradation Exists—Reference Sites Cannot Be Identified From Current
Monitoring Data
Historical Analysis ofEstuarine Data
Because of nutrient overenrichment, it is highly likely that no minimally impacted estuarine waters are
available for many coastal areas. Under these circumstances, the "best" of the existing waters are clearly
degraded. An alternative is to establish reference conditions from the historical record. Three
approaches are recommended: (1) analysis of historical ambient nutrient and hydrographic data; (2)
analysis of sediment cores to reveal the historical record, including the paleoecological record; and (3)
model hind-casting. The analysis of ambient data is likely available only for a small number of estuaries
(e.g., Chesapeake Bay and some tributaries, San Francisco Bay, Narragansett Bay, Tampa Bay), because
many estuaries were impaired before nutrient data were initially collected.
Empirical In Situ Data Analysis
The first approach depends on the availability of an adequate database. Such a database will need
considerable scientific judgment even for systems with a relatively rich database. For example, the
ambient nutrient-based historical record for the entire main-stem Chesapeake Bay becomes much less
abundant during the 1970s and earlier (Harding and Perry 1997). In addition, the spatial coverage was
much less widespread in earlier years. In many systems, early studies focused primarily on the deep
channels because of the interest in hypoxia. Analysis of ambient trends in chlorophyll a may be
complicated by the interaction with freshwater flow, as reported by Harding and Perry (1997) and Hagy
et al. (2000). This co-linear effect may affect assessment of trends in water clarity and cause a nonlinear
relationship between a limiting nutrient and hypoxia (Chapter 2). Hypoxia is the result of an extended
temporal effect of nutrient loading (lag effect), so published empirical relationships usually are based on
seasonal or annual nutrient loads (Chapter 2), not on short-term concentration data.
An important aspect of the historical approach is the selection of a period when nutrient loading caused
minimal use impairments. Often the nutrient monitoring data are more continuous for past decades than
is documentation of use impairments (e.g., visual pollution and reduction in fish productivity) due to
nutrient overenrichment. The economics and technology that affect fishery statistics in estuaries and
coastal waters are not easily translated into quantitative fish population abundance data. In some systems
(e.g., Tampa Bay, Janicki and Wade 1996, Greening et al. 1997), seagrasses were monitored to
demonstrate recovery but not necessarily prior to nutrient-based losses. The effects of increased nutrient
loads may be confounded by increased suspended sediments. Sediments can cause light limitation and
impair benthic habitats, with potential negative effects on living resources.
Historical ambient nutrient-related data will encompass seasonal and interannual variability. It is
important that the key variables (TN, TP, chlorophyll a,1 and water clarity—usually Secchi depth) be
distributed in a representative spatial and temporal manner. Although the Secchi disc is inexpensive and
has been widely used, it is recommended that future measurements of water clarity employ submersible
It is recommended that future algal pigment analyses consider some high-performance liquid chromotography (HPLC)
measurements that can also quantify algal groups.
6-8 Nutrient Criteria—Estuarine and Coastal Waters
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PAR meters as a minimum and that submersible spectral radiometers will be used more frequently.
Some precautions are relevant. For example, comparing nutrient-related variables in the turbidity
maximum, if light limitation dominates algal net production, would seem to provide little insight into the
nutrient problem. Another consideration is to ascertain where on the long-term hydrograph the reference
period falls (e.g., a wet or dry decade). These past measurements can then be used as a basis for
comparison to present condition (Figure 6-3). If a spring bloom is evident in the data, a judgment is
required as to whether remineralization of the spring bloom helps fuel a summer algal bloom, and
whether the spring bloom contributed directly to summer hypoxia or other use impairments. Therefore, it
is recommended that two averaging periods, a spring and summer period, be used.
The magnitude and duration of historical algal blooms are expected to be much lower than those of
current blooms in nutrient overenriched waterbodies under similar physical conditions. Some evidence
indicates that the magnitude of algal blooms (e.g., coefficient of variation) may increase as systems
become more enriched. At some point light limitation may limit variability. The historical data should
be aggregated within a physical classification, such as salinity zones, similar to present data analysis.
Several options exist for data summation. A point-in-time estimate may not capture the anticipated
seasonal and interannual variability in the data. In some cases, data for one or more key variables may
exist before an inflection, indicating worsening conditions. Such variables may include increase in
concentrations or loads of TN or TP, increase in chlorophyll a, decrease in Secchi disk values, and
increase in hypoxic volumes. If this more ideal case prevails, then seasonal medians (medians are less
sensitive to outliers or extreme values in a distribution than are means) or medians of seasonal index
periods (month of highest chlorophyll a) should be calculated. If available, the median of seasonal
medians for one or more years is desirable and may be essential to capture interannual variability.
Chlorophyll-a
Increased
Unattached
Algae
Fish Kill
Frequent
Fish
Kills
Loss of
SAV
TN.TP
Clarity
1950
1975
2000
Figure 6-3. Hypothetical example of load/concentration response of estuarine biota to increased enrichment.
Dashed line represents the selected reference condition level.
Nutrient Criteria—Estuarine and Coastal Waters
6-9
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Years may need to be excluded from the reference period if the reference period includes increasing
problem-related values (e.g., chlorophyll a and extinction coefficients). Nutrient and chlorophyll a
concentrations and Secchi disk values often will show a gradient within a given salinity zone. In data-
rich cases within similar salinity zones, one may express the values as a frequency distribution (e.g.,
Figure 6-4; see EPA 2000a, Chapter 6, EPA-822-BOO-001, for more details). However, attention needs
to be paid to potential confounding effects (co-linear) of freshwater input, even when comparing similar
salinity zones, because vertical density stratification may be a dominant controlling influence and be
dissimilar over time. When the above precautions are addressed and confounding factors are poorly
understood, it may be appropriate to set the reference condition at the median between the historical
median and the median for present data (Figure 6-4). This simple procedure reflects the magnitude of the
departure from minimally impacted waters and is, in part, a function of the length of the historical
database, addresses inherent variability, and is a realistic approximation of a reference condition over the
time span.
Sediment Core-Based Reconstruction
Sediment core analysis is becoming widely used in assessment of historic biogeochemical and
climatological conditions in lake and marine environments (Brush 1984, 1986; Cooper 1995). The
approach is applicable in sediment depositional areas where bioturbation and other forms of sediment
disturbance are minimal. Improved sediment dating techniques (e.g., lead-210, cesium-154, carbon-14)
have contributed to the understanding of historical conditions when sediments were deposited. Certain
Historical Data
75th %ile
Present Data
Full or Random
Sample Distribution
Full or Random
Sample Distribution
Nutrient Criteria Variable Concentration
"Historical
Median "Quartile
Option" Option"
"Median of Both Distributions" Option
Historical Data = Observations Between 1950-1968.
Present Data = Observations Between 1990-2000.
Figure 6-4. An illustration of the comparison of past and present nutrient data to establish a reference condition for
intensively degraded estuaries. The option of selecting the distributions from both time periods is compared to an
expected frequency distribution if the observations were available.
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Nutrient Criteria—Estuarine and Coastal Waters
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other chemicals (e.g., plasticizers) began to leave signatures when developed in the 1930s. Sediment
cores may provide information about metals and organic chemicals with long half-lives. Many
environmental indicators of past sediment conditions now are widely reported in the literature. These
include pyrite formation related to anoxia, history of submerged aquatic vegetation, diatom composition
related to nutrient enrichment, presence and composition of certain dinoflagellate species, distribution
and abundance of indicator species of Foraminifera, hard body parts of molluscs and crustaceans, land
use based on oak and ragweed pollen, total nitrogen and phosphorus profiles, and stable isotope analysis
of carbon, nitrogen, and sulfur in organic materials. Hind-casting with sediment cores provides a means
to infer reference conditions at a time when nutrient concentrations were much lower than present. In
shallow estuaries where depositional areas are not present or questionable, sediment cores have limited
applicability.
Model Hind-Casting
Hind-casting is a controversial approach because it is difficult to calibrate and verify a model when the
data difficult to quantify, as with past nutrient and hydrographic conditions. Running a computer-based
environmental process model backwards involves many uncertainties; however, where ambient data are
inadequate, sediment cores are not applicable, and a model is available and in the hands of experienced
scientists, then with expert guidance it seems reasonable to estimate reference conditions on a first-order
basis. Use of geographic-based land use models coupled to estuarine hydrodynamic-algal growth models
is one approach to hind-casting. Chapter 9 describes several models that may be useful.
Watershed-Based Approaches for Estuarine Reference Condition
An alternative to using in situ coastal marine data to establish reference conditions is to base estimates on
watershed nutrient loading characteristics. One can, in some situations, use watershed loading estimates
to define conditions in the watershed where nutrient loads would represent minimally impaired waters
downstream in estuaries. This assumes that in an undisturbed estuary and its watershed, the nutrient
load would historically represent the most natural condition. In some cases, the watershed above the
confluence of major freshwater streams may contain a relatively low level of human development, but
the nearshore estuarine area may contain considerable development (e.g., Elevenmile Creek, Perdido
Bay, AL/FL, and development near lower Perdido Bay distributary systems; see Livingston 200la).
Some estuarine watersheds may still contain a tributary, or segment thereof, whose nutrient load
represents a minimally impaired stream system because human development is minimal. The minimally
impaired stream nutrient loading to head of tide may be used to estimate the nutrient load for the entire
watershed if certain assumptions are met. Application of a watershed stream segment to estimate
nutrient loading is based on seven assumptions, as described in Table 6-2, including atmospheric
deposition of nitrogen as a constant.
Areal Load Approach to Identification of Reference Condition
The nutrient load is measured for the minimally disturbed subtributary or segment. State and national
preserves, when available, typically offer appropriate conditions (Clark et al. 2000, Fulmer and Cooke
1990). A decision is required as to whether the geology is approximately homogeneous across the
watershed to extrapolate from the reference tributary to the entire suite of tributaries. If not, then a
Nutrient Criteria—Estuarine and Coastal Waters 6-11
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Table 6-2. Requisite assumptions for establishing watershed-based reference conditions
Assumptions
Description
Estuarine systems are usually unique rather
than being readily divided into similar classes of
estuaries.
There are some instances of coastal bays that can be classified together
so that one or more of them may be designated as reference sites that
collectively may comprise a reference condition against which other
similar estuaries in a given area can be compared. However, most
often each estuary must be addressed individually and the reference
condition must be derived from data within that system.
The tidal factor and large volume shifts make
regional subdivision of each overenriched
estuary difficult.
In the instance of biological criteria development, it was possible to
subdivide the estuary by salinity habitat regimes and thus reduce the
portions of the estuary and its water column dynamics to a manageable
level, especially for assessments of nonmobile benthic invertebrates.
This is not viable with nutrient-related planktonic organisms and
dissolved or suspended water column materials.
We can estimate "pristine" or natural loads
from estimations of concentration-flow
relationships, and therefore loading estimations
in unaltered subestuarine watersheds. That
loading can be extended to an entire estuary for
reference condition development.
Loading estimations are an established practice in water resource
management. The universal soil loss equation (USLE) is the best
known example, and it has been extended to other unit loading
estimations. This robust and rather straightforward concept estimates
most of the nutrients that enter an estuary regardless of how effectively
the system processes or assimilates nutrients.
Most of the measurable load to an estuary is
tributary based, and atmospheric deposition is a
constant over the system.
Admittedly, there is an atmospheric source of nitrogen to and from
estuarine waters, but this is variable and, generally, is much less than
runoff from streams and rivers. The effect is somewhat mitigated
because the atmospheric deposition to the stream surfaces is
incorporated in the loading estimation. Phosphorus and suspended
material as well as algal responses to nutrients are most certainly more
tributary related than atmospheric. Shoreline sheet runoff can be
incorporated into the loading approach.
Loading from coastal marine waters is usually
negligible compared with anthropogenic
watershed loads to the estuary
While many estuarine ecologists are properly concerned with this
aspect of estuarine dynamics, from the practical criteria development
approach, a large portion of the marine load may be presumed to have
originated in the previous outgoing estuarine tidal water. The
remainder is to some extent a part of a natural process inherent to
estuarine systems. This assumption may not prevail when estuaries
enter deep, upwelling, oceanic waters.
The predevelopment nutrient loading rates
expressed as yield per watershed land area are
similar within a single geographic region (e.g.,
province, ecoregion, or subecoregion). Local
regional uniformity of geography is assumed.
This is a reasonable expectation in the absence of extensive land
development with attendant anthropogenic discharges and runoff. The
geographic subdivisions of a natural landscape can be expected to be
homogeneous by definition (i.e., similar soil type, topography, and
vegetative cover).
Because the National Nutrient Criteria Program
assumes that groundwater influence is a
separate loading factor to surface water
eutrophication, groundwater-dominated
estuaries should be treated separately for the
development of nutrient criteria.
The groundwater factor can be highly significant in some localities
such as coastal Florida. This generalized approach does not address
that factor, but Regional Technical Assistance Groups should be aware
of the groundwater contribution and account for it in their estimations.
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Nutrient Criteria—Estuarine and Coastal Waters
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second minimally disturbed subtributary should be sought to represent a different geology; this logic
should be applied if additional subwatersheds differ substantially. Nutrient loads are then extrapolated in
a simple proportional manner from the reference tributaries to the entire tributary system within the
watershed or subunit and the aggregate load is calculated (Figure 6-5). This load represents the
"reference load" reflecting in situ reference conditions for the four primary indicators (TN, TP,
chlorophyll a, and water clarity).
Extrapolation from a reference tributary can be augmented by application of geographic-based nutrient
erosion and transport models. This nutrient load would become the target load for the downstream
estuary or coastal waterbody. Consideration should be given to the representative nature of the
freshwater flow over the average hydrograph. It is desirable and may be necessary to obtain a measure of
the average multidecadal freshwater flows at the head of tide. Nutrient loads based on a drought period
would not accurately represent conditions in terms of nutrient-based ecological impairments. Extremely
high flows are important, but they are likely to fall outside of resource managers' capabilities to solve a
nutrient problem. Therefore, this approach has its limits. Large fluvial streams do not necessarily
transport the most upstream load to the lowest fluvial portion of the stream tributary; in-stream
ecosystem processes modulate the load. In summary, a sequence of steps is outlined in the following box
to complete the areal load watershed approach for reference condition determination.
Coastal Reference Conditions
Following is an example of a nutrient loading assessment from a very large watershed, the Mississippi
River system. In 1997, the EPA Gulf of Mexico Program, through a Mississippi River/Gulf of Mexico
Watershed Nutrient Task Force, asked the White House Office of Science and Technology Policy to
conduct a scientific assessment of the causes and consequences of Gulf hypoxia through its Committee
on Environment and Natural Resources (CENR). The National Oceanic and Atmospheric Administration
(NOAA) was asked to lead the assessment. The assessment included various computer-based modeling
approaches to characterize the nutrient delivery to coastal Louisiana and Texas.
Nutrient load reduction within the various watersheds to meet resource management objectives is an
analogue of the reference condition approach (see NOAA website for details:
www.nos.noaa.gov/products/pubsjiypox.html). Two conditions are considered: coastal estuarine
plumes and shelf waters.
Coastal Estuarine Plumes
The foregoing example for the Mississippi River watershed is a large-scale effort to assess nutrient
conditions associated with a large coastal nutrient plume. For large estuarine watersheds with the
potential to cause nutrient overenrichment on the continental shelf, it is important to extend the reference
concept beyond the local area. In many cases, the concerns will be large enough to warrant development
of a hydrodynamic model with coupled nutrient-phytoplankton growth kinetics (Chapter 9). Smaller
estuarine plumes along the coast may be addressed through a well-designed research and monitoring plan
with expert input on design features. An initial EPA-sponsored sampling effort in the Mid-Atlantic is
currently underway to provide range-finding data to assist in development of a more comprehensive
Nutrient Criteria—Estuarine and Coastal Waters 6-13
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Figure 6-5. Areal load estimate approach to nutrient reference
condition determination. The watershed is estimated to be
approximately 368 square miles; the reference tributary streams
representative of above head of tide systems in the watershed are
approximately 20 square miles combined. The median load estimate at
the mouths of the tributaries could therefore be multiplied by 18.4 to
approximate a reference condition load for the river.
monitoring plan. The first step is to assess the shelf s mixing and dispersive capabilities to attenuate the
negative effects of nutrient enrichment. Bloom development along physical discontinuities should be
assessed. If dispersive mechanisms are large enough to attenuate phytoplankton blooms, then concerns
are given a lower priority. In coastal estuarine plumes where the physical processes may not attenuate
the nutrient enrichment effects to acceptable levels, then an appropriate level of research, monitoring,
and modeling may be required to assess nutrient reduction from upstream nutrient sources as well as
from seaward upwelling of nutrients and atmospheric deposition (Paerl and Whitall 1999, Vitousek et al.
1997,Howarthetal. 1996).
Open Coastal Water
The NRC (2000) publication (Chapter 6) recommends an index site approach for estuaries and coastal
waters. The index site approach in particular has merit because the continental shelf is very extensive and
too large for the Nation to conduct comprehensive studies of all sites potentially affected by nutrient
enrichment. A priority-setting rationale should be based on a physical classification system that arrays
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Nutrient Criteria—Estuarine and Coastal Waters
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coastal water masses relative to their susceptibility to nutrient enrichment, especially waters that are
likely to respond similarly to nutrient enrichment. This would allow resource managers an opportunity to
apply unique "dose-response" curves to a particular coastal water class. Because such a system likely is
relatively crude, the best professional judgment of experts should augment physical classification
systems. Because this approach has a significant research element, it would be appropriate to begin it in
the competitive research forum with oversight provided by some mix of Federal, State, and university
participation. Chapter 7 also discusses the roles of monitoring and modeling, which offer useful insights
applicable to the coastal ocean. This manual recommends that the index site approach be given serious
consideration, especially for coastal waters including estuarine plumes. The coastal ocean is large and
oceanographically dynamic and complex. Thus, assessment of individual States' contribution to nutrient
overenrichment will in most cases require a Federal and States' partnership. Even in large estuaries
whose watersheds or tidal waters are shared by more than one State, a multi-State agreement is probably
required (e.g., Chesapeake Bay Program and the New York and Connecticut TMDL Agreement).
The Coastal Research and Monitoring Strategy, an element of the Clean Water Action Plan, contains
approaches that can help determine reference conditions. The strategy provides for a coordinated effort
among Federal, State, and private agencies. Clearly, an approach that coordinates use of aerial
surveillance tools (e.g., satellite-based water quality sensors), data buoys, and ship-based measurements
(especially ships of opportunity such as the North Carolina Albemarle Sound ferryboat monitoring
program; H. Paerl, personal communication) within an index site to underpin a cause-and-effect
framework is highly desirable.
On a provisional basis, any additional monitoring might include a stratified random approach (e.g.,
EMAP), because this provides an opportunity to address known ecological structure and functional
processes and unbiased trend monitoring. It is important to continue monitoring that involves identified
relationships. A challenge will be to design a program that can distinguish the effects of natural coastal
processes (e.g., nutrient upwelling) from anthropogenic influences (atmospheric nitrogen deposition,
fluxes of nutrient from estuaries, and possible expansion of mariculture activities).
One such investigation is the design presently being tested by the National Nutrient Criteria Program in
the Mid-Atlantic Bight. This near-coastal marine nutrient sampling protocol is intended to identify
inshore (within the 3-mile limit) reference sites based on land use and physical coastal characteristics
together with comparisons to offshore nutrient water quality. A stratified-random approach is used, and
the compiled data from reference sites establish the reference condition for that portion of the coastal
marine waters. Riverine and estuarine plumes or other discharges can then be evaluated relative to this
minimally impacted condition. Sampling is recommended for spring and summer conditions with
multidepth collections. The technique is being tested in a variety of State waters. A description of the
design and preliminary results is presented in Appendix H.
Nutrient Criteria—Estuarine and Coastal Waters 6-15
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1. Identify major tributaries to the estuary. Classify similar tributaries by physical size and freshwater delivery and
compare similar geological features of the sub-watersheds for classification.
Three nutrient loadings can be estimated:
A. Existing (actual current load). Existing load is estimated from the known tributary loads as measured at the mouth
above head of tide and extrapolated to similar land units in the watershed, plus any shoreside runoff and discharges
that directly enter the estuary, plus direct atmospheric deposition. This is the present status of the waterbody.
B. Best existing load. Find the best existing tributary(ies) or best subtributary(ies) in the region and calculate the
nutrient yield. Extrapolate this value to the rest of the watershed tributaries or subtributaries as appropriate.
C. "Pristine " or unimpacted load. Identify a regional reference streams in an undisturbed watershed having little or
no development, such as State or national preserves, that can be used to estimate areal yield for the region
approximating an entire unimpaired, undeveloped condition as though there were no significant cultural impairments.
2. Find the tributary(ies) with the least impaired status and minimal disturbed lands contributing to nutrient
loads.
These are B and C-type conditions from above. Each of these tributaries is monitored enough to establish a nutrient
load. Seasonal and interannual variability should be assessed. Cross-sectional sampling at the head of tide is required
because of nutrient flux variability. The USGS has established protocols for this type of stream monitoring. Where
available, dams on rivers, if located near the head of tide (e.g., Conwingo Dam on the lower Susquehanna River, MD),
make very desirable locations to measure nutrient loads.
3. Estimate the annual areal nutrient yield for TN and TP.
Extrapolate from the unimpaired or minimally impaired watersheds above head of tide to all other similar tributaries in
the watershed and apply the load estimate to direct runoff portions of the tributary. Do this for each monthly or
seasonal increment throughout the year. Do the same for each major tributary.
4. Extrapolate the nutrient yield to the entire watershed land area within the region.
For example, generate total best existing nutrient loading. This will, in many cases, represent the best attainable loads.
5. Repeat for other regions if the estuary watershed covers more than one major geological landform.
This is necessary to comply with the assumption that regional homogeneity within the watershed covers only part of
the entire estuarine watershed.
6. Sum the nutrient yield for all tributaries within the estuarine watershed.
Do not factor in atmospheric nitrogen loads, as they were incorporated into the tributary loads. The atmospheric
nitrogen loads may need to be reduced to achieve an acceptable nutrient condition in the estuary. The summation of
tributary loads becomes the estuarine reference condition. Computer modeling may be required, especially in larger
watersheds, because in very long tributaries some nutrients, especially phosphorus, may become embedded in the
stream bottom, and some nitrogen and phosphorus is potentially lost on the scale of years to decades in the floodplain.
The chlorophyll a concentrations in the estuary will need to be modeled from the reference nutrient loads and some
measure of central tendency of freshwater inflows to the estuary. When a larger estuarine system may dominate the
lower tributary estuarine hydrodynamics (e.g., the mainstem of the Chesapeake Bay dominates the hydrodynamics in
the lower Patuxent River estuary), then as a minimum, a box modeling approach may be required to account for the
two-estuarine interactions of the freshwater inflow and lower estuarine interactions (Hagy et al. 2000).
6-16 Nutrient Criteria—Estuarine and Coastal Waters
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CHAPTER 7
Nutrient and Algal
Criteria Development
7.1 INTRODUCTION
Role of RTAGs
Classification
Descriptive Background Information
Elements of Nutrient Criteria
Hypothetical Examples of Nutrient Criteria Development
Determinations
Evaluation of Proposed Criteria
Nutrient Criteria Interpretation Procedures
Criteria Modifications
EPA, State, or Tribe Responsibility Under the CWA
Implementation of Nutrient Criteria
This chapter addresses the details of developing scientifically defensible quantitative criteria for
nutrients, algae, and measures of water clarity (referred to hereafter as nutrient criteria). The chapter is
divided into eight sections: (1) role of the RTAG; (2) classification; (3) descriptive background
information; (4) approaches to criteria development; (5) evaluation of proposed criteria; (6) criteria
modification; (7) EPA, State, and/or Tribal responsibilities under the Clean Water Act (CWA); and (8)
implementation of nutrient criteria into water quality standards. The five elements of criteria
development described in the Executive Summary and Chapter 1 are integrated herein, and additional
information relevant to criteria development is provided.
As explained in Chapters 2 and 3, estuaries and coastal waters are especially complex ecosystems where
it is often difficult to distinguish the effects of anthropogenic nutrient enrichment from natural variability
primarily because of intervening physical processes and their interaction with biological (e.g., grazing;
Cloern 1982) and chemical (e.g., flocculation and sedimentation) processes (Malone et al. 1996, NRC
2000). The individual nature of many of these ecosystems presents a particular challenge for criteria
development. The ideal goal is to establish nutrient criteria that are protective during periods when
estuaries and coastal ecosystems are most vulnerable to nutrient enrichment and that protect designated
uses. It is important to understand that designated uses may be met but some nutrient-based impairment
may have occurred. In such cases, it is also desirable to have restoration goals in mind whose objective
is to restore the original ecological integrity, at least as represented by the reference conditions and
criteria. This information helps determine if new designated uses are appropriate.
If a shallow estuary is dominated by point sources of nutrients, then low freshwater flow periods might
be times of greatest vulnerability because of limited flushing. For a deep estuary under the same
situation, weak density stratification may set up conditions where the algae spend considerable time
below the euphotic zone and, hence, bloom development is minimized. Now, consider the situation
where these two estuaries are dominated by nonpoint sources of nutrients. The low flow period would
likely contribute lower nutrient loads to both types of estuaries and with weaker density stratification the
algal bloom potential might be substantially lowered (see Chapter 2). It is also extremely difficult to
establish nutrient criteria for episodic events, either hurricanes or drought periods. The reference
condition and consistent reference site records, however, make it possible to adjust the criteria
accordingly to address these intervals.
Nutrient Criteria—Estuarine and Coastal Waters 7-1
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In this manual, the elements of criteria development are sequenced and emphasized somewhat differently
from those described in the two published freshwater manuals. Because of the relatively high individual
nature of estuaries and coastal waters, the role of the regional technical assistance groups (RTAGs) is
enlarged and historical data become especially important in development of nutrient criteria. Figure 7-1
provides a visual perspective of the elements that should be integrated to arrive at a criterion.
7.2 ROLE OF REGIONAL TECHNICAL ASSISTANCE GROUPS
Expert evaluations are important throughout the criteria development process. The role of the RTAG in
criteria development for estuaries and coastal waters has an added dimension over that applicable to
lakes and reservoirs and rivers and streams. In the latter case, most of the data used in development of
criteria recommendations resided in national electronic data sets (e.g., STORET) collected by State and
Tribal agencies. The RTAGs helped review these data for duplication and outliers. They also
encouraged States and authorized Tribes to submit additional data to STORET. Under these
circumstances, EPA developed criteria recommendations based on frequency distributions (U.S. EPA
2000a,b). However, for most estuaries and coastal waters, the majority of relevant data have been
collected by universities and other organizations (e.g., NOAA, Minerals Management Service, USGS);
many of their data are not entered into STORET. It is expected that regional RTAGs will likely be
considerably more knowledgeable about the data veracity and applicability for criteria development in
their region. In addition, it is anticipated that the regional RTAGs will have the knowledge and access to
other local scientific expertise to assist in the sampling design and collection of additional data. These
considerations lead to the expectation that the RTAGs will have a larger role in the development of
protective nutrient criteria for estuarine and coastal waters.
Values of
Contributing
-------
Because of the tendency of estuaries and coastal waters to exhibit a high degree of individuality relative
to nutrient susceptibility, little predictive success has been demonstrated in using published values of
nutrient concentrations or nutrient impairments. However, experience with an individual estuary or
coastal water system may demonstrate a general range of algal biomass accumulation that leads to
hypoxia at a known level of enrichment. This type of assessment often requires several to many years of
observation and measurement and is an example of incorporating the RTAG experience in the criteria
development process.
EPA expects to continue to lead the effort to identify potential estuarine and coastal waters for the
development of nutrient criteria and fund the overall data collection and analysis.
7.3 CLASSIFICATION
Classification is a pivotal step in the process of developing criteria. Physical classification of
waterbodies for nutrient criteria development reduces variability in ambient measurements as the
reference conditions and nutrient criteria represent and reflect a relatively similar natural state.
Classification helps ensure that appropriate comparisons are made among comparable waterbodies so
that the variables measured are influenced as little as possible by dissimilar inherent characteristics. This
will facilitate appropriate application of criteria through their implementation.
In contrast to rivers and lakes, physical classification of estuarine and coastal waters is scale-sensitive
(Giller et al. 1994) and may not be as predictable of nutrient enrichment effects or be as useful for
generalizations about effects among estuarine systems. Classification nonetheless can provide improved
understanding of the processes that contribute to ecosystem susceptibility and variability in the
expression of nutrient effects. Classification may have valuable applicability at smaller physical scales
within larger estuarine and coastal ecosystems (e.g., embayments; subestuaries and estuaries discharge
plumes). Classification based on salinity gradients, circulation patterns, depth, and flushing within larger
estuarine and coastal systems should also prove useful, especially in correlating the different biological
communities at risk to nutrient overenrichment.
7.4 DESCRIPTIVE BACKGROUND INFORMATION
Estuarine Watershed Characterization
One of the keys to understanding nutrient enrichment problems in waterbodies is an environmental
characterization of the watershed from a historical perspective. Such investigations may provide insights
into the potential for confounding cause and effect relationships (e.g., nutrient, not herbicides as the
primary cause of the SAV decline in Chesapeake Bay.) Historical changes in land use, land cover, and
population demographics may correlate with increased anthropogenic-based nutrient enrichment and
major changes in downstream water quality and biological community structure.
Farm acreage, crop types, and fertilizer application rates also provide useful information to assess the
potential historical magnitude in nonpoint source nutrient loading to estuarine systems. Atmospheric
Nutrient Criteria—Estuarine and Coastal Waters 7-3
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plumes that result in nitrogen deposition should also be assessed as they may help explain an increased
nitrogen load to coastal waters beyond that attributed to local land use activities. This also is a situation
where reference sites, if available, will help the manager distinguish atmospheric from local
anthropogenic causes of overenrichment. Conversely, increases in forest coverage and soil banking of
agricultural lands may help explain potential decreases in nitrogen and phosphorus loading to estuaries
and coastal waters.
Within Estuarine System Characterization
Changes within estuarine systems that influence basic hydrography should not be overlooked. For
example, opening new passes, and later deepening them in several Gulf of Mexico estuaries during the
early part of the 20th century (e.g., Perdido Bay, Alabama/Florida, and Choctahatchie Bay, Florida)
apparently modified estuarine circulation resulting in strengthened density stratification leading to
enhanced potential for hypoxia (Livingston 2001a). Also, major changes in freshwater supplies should
be considered as a potential factor that can modify estuarine susceptibility to nutrient enrichment.
The use of marine sediment cores is another tool to assist in assessment of nutrient enrichment patterns in
coastal waters (Brush 1984). These analyses are relatively expensive to perform but appear more
frequently in the literature because of the numerous important insights they provide. They can provide
estimates of sedimentation rates and initiation of anoxia, changes in algal community structure, initiation
of the loss of SAV and other responses to nutrient enrichment. This temporal picture is important in
setting approximate timelines when nutrient enrichment may have been a major cause of biological
impairment. However, correlation does not necessarily equate to causality (Havens 1999).
It is also important to attempt to collect long-term (e.g., multiple decades) fisheries landings data as many
stakeholders need to be appraised of whether such landings data are associated with nutrient enrichment.
Such analyses often prove to be difficult because information on catch per unit effort that helps
normalize for variable fishing pressure, is difficult to obtain. A change in fishery yield may be
confounded by overfishing, as well as the role of increased or decreased primary productivity. Increased
bottom water hypoxia related to overenrichment may explain the loss of benthic habitat for bottom-
dwelling marine life (e.g., flounder and croaker and benthic infauna that serve as fish food).
Historical decreases in water column visibility from nutrient-driven algal blooms (phytoplankton and
macroalgae) may explain reductions in water-borne recreation (e.g., swimming); however, human
perceptional responses are often subjective and other factors such as user conflicts may also be involved.
Reduced visibility may also be related more to inorganic suspended sediments then nutrients. So, both
chlorophyll a records and total suspended sediment concentrations, preferably measured from the same
local water mass, may be required to establish nutrient enrichment (Gallegos 1994). Such analyses need
to be assessed with freshwater flow records because of potential co-linear effects that complicate
interpretation of cause and effect relationships.
The above measurements performed as part of a general characterization may help relate nutrient
enrichment effects and thresholds to existing and designated uses and identify any reference systems that
7-4 Nutrient Criteria—Estuarine and Coastal Waters
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are minimally impacted by nutrient pollution. Such information may enter directly into predictive water
quality models or serve as indirect collaborative information contributing to a "weight of evidence"
analysis.
7.5 ELEMENTS OF NUTRIENT CRITERIA
Reference Condition
Chapter 6, beginning with Table 6-1, describes several approaches that can be employed to determine the
ambient minimally impacted nutrient condition of the water resource. The significance of this reference
condition to nutrient criteria development cannot be overstated. It represents the determination of the
existing, presently attainable nutrient water quality of the estuarine or coastal waters of concern.
The selection of the method for reference condition determination in this manual involves more options
than previous guidance manuals because so many estuaries are reported to be unique and/or severely
degraded, thus requiring an array of alternative approaches to approximate reference conditions in the
absence of acceptable reference sites. In selecting from the different approaches, the resource manager
should strive for the most direct measurement of the resource and with the least number of intermediate
interpretative steps to a determination. Of equal importance in this process are in situ reference sites and
supporting data showing the system response to the nutrient increases. The best of both these worlds is a
set of reference sites documenting an optimal nutrient condition as well as response data confirming that
system degradation occurs at levels beyond this measure, which also corresponds to the EPA regional
reference condition for that area and class of waters. Failing this, the manager should seek the greatest
approximation possible and a sufficient understanding of the divergence to be confident of the reference
values determined.
Even though the reference condition is salient to nutrient criteria development, it should not be
interpreted as the only necessary element. It should be interpreted in light of the historical condition of
the resource and projections of its future potential.
Historical Information
Knowledge of antecedent conditions is particularly important in the case of estuarine waters, where
causal relationships are often confounding and existing reference sites compromised. In such cases the
historical data may not only qualify present information, they may, in fact, be the requisite reference
condition demonstrating to the manager and RTAG not only previous nutrient quality, but least impacted
conditions as well.
Models
In addition to computer modeling to help determine reference conditions, models may be used both to
estimate nutrient loads and load reductions to achieve a targeted nutrient regime in the receiving
estuarine and coastal waters systems. Various models are available (Chapter 9) that estimate nutrient
erosion from different land uses, riverine transport of nutrients, and estuarine effects including hypoxia
and chlorophyll a concentrations. These models are typically expensive to calibrate and verify, but in
Nutrient Criteria—Estuarine and Coastal Waters 7-5
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large estuaries and coastal ecosystems their application can be cost-effective when costly nutrient control
decisions are involved. A coupled nutrient transport and hydrodynamically coupled algal nutrient uptake
and growth model provides the ability to address "what if scenarios" of nutrient load versus level of
biological impairments (see Chesapeake Bay Case Study and web site).
Statistical models can be used to help separate effects of nutrient loading from estuarine physical
processes as determinants of increases in system response variables (e.g., chlorophyll a concentrations;
Harding and Perry 1997). Effects of nutrient enrichment are inferred by a process of elimination when
the suspected physical forcing functions are de-trended in the analyses and are inferred not to explain the
variability. Box models, using salinity as a tracer of water masses, are useful in assessing net non-tidal
physical circulation, hydraulic residence times, the effect of river flow on residence time versus seaward
higher salinity processes, and extent that nutrient sources are conserved within an estuarine region or
transported seaward (Hagy et al. 2000). Dettmann (in press) used a regression modeling technique to
compare the degree of nitrogen export of a variety of estuarine systems. Boynton et al. (1996) used a box
model to mass balance nitrogen and phosphorus in Chesapeake Bay and calculated net transport from the
Bay of nitrogen and phosphorus, sedimentation, amount tied up in plants and the amount of nitrogen lost
from the Bay through denitrification. Properly applied box and regression models are relatively
inexpensive to construct and can provide useful information to the scientist and water quality manager.
Antidegradation Policy and Attention to Downstream Effects
A critical requirement for the use of reference conditions associated with nutrient criteria is the EPA
antidegradation policy, which protects against incremental deterioration of waterbodies and reference
conditions. An observed downward trend in the conditions of reference sites cannot be used to justify
relaxing reference expectations, reference conditions, and the associated nutrient criteria. Once
established, nutrient criteria should only be refined in a positive direction in response to improved
conditions. Without antidegradation safeguards, even the establishment of reference conditions and
nutrient criteria could still allow for continual deterioration of water quality.
To combat this, the States should implement an effective antidegradation policy that promotes
continually improving conditions. As an example, Maine has an antidegradation policy that requires that
waterbodies remain stable or improve in trophic state (Courtemanch et al. 1989, NALMS 1992). The
RTAG should assume a comparable sense of antidegradation responsibility.
Estuaries that supply nutrients to relatively static coastal waters may require more stringent nutrient
criteria, not only to protect estuarine designated uses (e.g., "fishable and swimmable" conditions), but the
water quality of coastal shelf waters. At present, there are little data to assess whether U.S. estuaries are
supplying nutrients to coastal shelf waters at levels that are causing widespread harm, except in the case
of world-class rivers (e.g. Mississippi River Plume on LA/TX shelf). Locally, river-dominated estuaries
with open passes to the coastal shelf supply nutrients at levels that may increase secondary productivity
of valued fisheries (Sutcliffe et al. 1977), but the potential threshold for overenrichment effects in such
cases is generally still poorly understood. If the RTAG determines that estuarine nutrient criteria may be
expected to fall between the existing present nutrient concentrations or load and the reference condition
7-6 Nutrient Criteria—Estuarine and Coastal Waters
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determined from similar unimpaired systems or from a historical load and response relationship. It is
then up to the States and Tribes to adopt the criteria into 303(c) water quality standards.
The RTAG
Assimilation of all of the above information is the responsibility of the RTAG when developing
ecoregional nutrient criteria and when reviewing State or Tribal nutrient criteria as part of its role to
assist EPA.
The RTAG should work with the States to develop a monitoring program that would evaluate the status
of the reference systems, the possible future negative anthropogenic nutrient effects, and the condition of
the estuary and coastal receiving water. A well-designed monitoring program should provide data to
assess whether there is incremental deterioration of the subject waterbodies and reference conditions.
7.6 HYPOTHETICAL EXAMPLES OF NUTRIENT CRITERIA DEVELOPMENT
DELIBERATIONS
To help illustrate the role and responsibility of the RTAGs, an abbreviated hypothetical
illustration of nutrient criteria development follows for a river-dominated estuary that has a relatively
deep channel with moderate density stratification, and well-developed seagrass meadows located in the
shallow waters. The estuary is located in the northern Gulf of Mexico. The estuary often borders on
both nitrogen and phosphorus limitation with nitrogen limitation occurring more frequently during the
summer. The nitrogen sources have been and continue to be primarily nonpoint sources in headwater
streams and point sources near the pass to the Gulf. Tidal action is minimal. The focus for this
illustration is on nitrogen criteria.
Scenario
In the subject nitrogen-limited estuary no existing areas qualified as a reference condition and no
meaningful analogs of the estuary were available to apply the spatial-based frequency/percentile
approach to reference conditions applied to lakes (e.g., the 25/75 percentile approach; see Lakes and
Reservoirs Nutrient Guidance Manual, U.S. EPA 2000a). The historical nutrient concentrations were
plotted overtime by classifying the summer estuarine salinity zones based on the 30-year average into
tidal fresh and brackish (0-5 psu), mesohaline (5-18 psu) and polyhaline (18-30 psu). The concentrations
were consistent with a calculated nutrient load and estuarine physical hydrodynamic model. Additional
analysis demonstrated that the system was most vulnerable to average summer freshwater inflows.
Flushing in the estuary was determined to be on the order of one month on average during the summer
under average freshwater flow conditions so the physical potential was high for phytoplankton bloom
buildup.
A 50 (iM TN reference condition was determined by plotting TN concentrations for the mesohaline zone
from 1970 to 2000. Because nutrient data were few for the estuary prior to 1970 and most of the data
were collected for the mesohaline zone, a decision was made to select the 1970 summer average TN
value as the reference condition with application for the mesohaline zone. Only an occasional bottom
Nutrient Criteria—Estuarine and Coastal Waters 7-7
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channel hypoxia occurred in the summer of 1970 and seagrass meadows were growing to a depth of 1 to
2 meters. By 1980, the average summer TN concentrations had increased to 60 (iM when it became
apparent from field monitoring that major loss of seagrass acreage was documented, hypoxia volumes
had doubled, and hypoxia-laden water had consistently reached the deep-channel/shallow water shelf
break. By 1980, seagrasses grew on average only to 0.5 meter depth. Some sediment core evidence
indicated that seagrasses in 1950 grew at a depth of 2 to 3 meters but the nutrient concentrations and
loading information were much weaker than for the 1970-2000 period. The reference condition of 50
(iM TN suggests a gradual rise upward in ambient TN, from an estimated average 20 (iM TN 100 years
ago, with a projected further upward trend indicated by use of demographic, land use, and hydrological
models. It was determined that a significant loss of ecological integrity had likely occurred prior to 1950.
The RTAG, therefore, concluded that setting a criterion any higher than the present reference condition
would eventually lead to an unacceptable trend in water quality degradation due to expected development
increases for this part of the estuary. If the model projections are accurate, any increased load implied by
raising the criteria above reference levels will hasten nutrient overenrichment problems. Some on the
RTAG argued that the criterion should be set at 54 (iM TN but the final consensus was to be somewhat
conservative. The RTAG therefore concludes that it will be prudent to set the criterion at 50 (iM TN (see
Figure 7-2).
7.7 EVALUATION OF PROPOSED CRITERIA
The RTAG will provide expert assessment of proposed criteria and assure that criteria protect designated
uses. Criteria will need to be verified in many cases by comparing criteria that apply across State and
Tribal borders. In addition, attention will need to be paid to downstream effects and designated uses,
Use Designations
Fishing/Swimming
Shellfisheries
Except. National
Resource Waters
Navigation 1
Boating 1
|m
^ Estuarine Reference
^ Condition & Criterion
50
57
64
TN,
7-8
Figure 7-2. Hypotetical illustration of developing a TN criterion in an estuary.
Nutrient Criteria—Estuarine and Coastal Waters
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especially in large estuaries that are shared by two or more States or Tribes. Criteria recommended by
the RTAG can be adopted by the State or Tribe and approved by EPA if evidence is presented that
assures no adverse effects will result downstream (e.g., criteria developed for tidal freshwaters may not
be stringent enough to protect uses located at higher salinities.) In estuaries, a consideration, not typical
of streams and lakes, is one where the life cycle of anadromous and semi-anadromous fish species must
be considered as well as marine spawners that utilize estuaries as nursery areas. The RTAG may need to
consult with neighboring Regional RTAGs in regions where estuaries are shared.
At present, EPA's Office of Water is developing a policy to address effects from nutrient transport that
causes water quality problems at downstream estuarine sites including river systems that deliver nutrients
from far inland to coastal tidal systems. If downstream designated uses are not protected by a proposed
criterion, then the river or stream criterion must be modified accordingly.
Guidance for Interpreting and Applying Criteria
A critical step in the criteria development process is to assess how realistically criteria can be
implemented into standards that are accepted by the public. It should be realized that today's designated
uses are not those that would be applicable in many estuaries at the turn of the century or in some cases
even several decades ago. Many estuaries have lost important fisheries that may not be easily recovered
if at all. For example, sturgeon are rare in many estuaries today when they were abundant decades ago in
several east coast estuaries. It is doubtful that the nutrient relationship for sturgeon growth and survival
is adequately known except for obvious factors such as hypoxia. The RTAG should make some
judgements about designated uses as exemplified by the sturgeon example that significantly improves
nutrient-based degraded water quality in terms of "fishable and swimmable" but maintains an important
degree of realism.
Do the Criteria Protect Designated Uses?
Section 303(c) of the CWA as amended (Public Law 92-500 [1972], 33 U.S.C. 1251, et seq.) requires all
States and authorized Tribes to establish designated uses for their waters. EPA's interpretation of the
CWA requires that wherever attainable, standards should provide for protection and propagation offish,
shellfish, and wildlife and provide for recreation in and on the water (section 101(a)). Note: this is the
secondary goal of the Act; the primary goal being the protection and restoration of the physical,
chemical, and biological integrity of the Nation's waters, and zero discharge of pollution. Other uses
identified in the Act include industrial, agricultural, and public water supply. However, no waters may
be designated to be used as repositories for pollutants (see 40 CFR 131.10(a)). Each waterbody must
have criteria that protect and maintain the designated use of that water.
Also discussed below are general guidelines for developing criteria to protect selected designated uses.
The values included here are not intended to represent proposed EPA or State estuarine and coastal
waters nutrient criteria. Rather, they are simply guidance and illustrate ranges of parameters associated
with the impairment of some designated uses in some States and Tribes. Criteria to protect these uses
should be developed on a site-specific basis when the individual nature of the estuary or coastal waters
require such specificity.
Nutrient Criteria—Estuarine and Coastal Waters 7-9
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Outstanding National Resource Waters
Some estuarine and coastal waters of the State may require special criteria based on unique
characteristics of that waterbody. Such characteristics might include undisturbed or unique fjords or
subestuaries or stretches of coastline that are markedly different from other coastal waters in the State.
Some areas may include threatened or endangered species that need to be protected. Such waters are the
very best of the reference set and are most in need of protection by rigid State and Tribal antidegradation
policies and procedures.
Aquatic Life Uses
Aquatic life uses, including fisheries and shellfisheries, are heavily dependent on the initial high quality
condition of the resource. Species will change as a function of trophic state, and it may be difficult to
defend why one species is necessarily "better" than another. The use of reference areas and their
accompanying biota is one measure that can be used to predict the species that should be expected in a
region.
Fisheries
Developing criteria to protect a specific fishery may be somewhat difficult because in open estuarine and
coastal waters fish species shift with seasonal migrations and salinity changes. However, basic response
variables such as available DO and turbidity can be incorporated to protect all seasonal fish and
crustacean communities and resident molluscan populations. Consultation with fisheries managers, the
recreational public, and commercial fishermen should help resolve any issues of targeted species
management through nutrient abatement.
Although our knowledge of the dynamics of change in the biota as a function of eutrophication requires
further development, there is sufficient evidence to conclude that eutrophication will bring species
changes. If an area has an existing aquatic life use, then that use must be maintained. (See 40 CFR
§131.12(a) (1).) Eutrophication will cause some species to change in relative abundance and cause
others to disappear; therefore, nutrient enrichment may be incompatible with the maintenance of a
specific biota. The ultimate extension of this concept is in the use classification of outstanding natural
resource waters.
Recreation
Swimming/Primary Contact Recreation
Criteria to protect a contact recreation use may be associated with the occurrence (or appearance) of
certain phenomena that affect certain types of recreation. For example, in general, swimmers will not be
affected by the trophic state of the estuary, but resulting changes in transparency or change in species
may be important. Dense planktonic or macrophyte growth may inhibit swimming opportunities and help
promote seasonal densities of sea nettles or "jelly fish." Excess nutrients feed not only nuisance algae
growth, but potentially health-endangering bacteria, especially when human or animal waste may be
involved such as when sewerage discharges are in the general area and may impact swimming when wind
and tide coordinate. This risk is not unusual for coastal ocean beaches where development promotes
sewerage expansion and offshore discharges.
7-10 Nutrient Criteria—Estuarine and Coastal Waters
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Boating and Secondary Contact Recreation.
It might be expected that the transparency of the water or the presence of algal scums would not deter
boating, unless water skiing were involved. However, boating may be affected by the presence of dense
inshore beds of tall or floating macrophytes.
Restoration Goals
As described in the introduction to this chapter, designated uses may be protected but some nutrient-
based ecological degradation already may have occurred. The public deserves to know what nutrient
conditions existed before anthropogenic nutrient enrichment initiated a shift from the natural nutrient
regime toward conditions of nutrient impairment within the limits of scientific knowledge or reasonable
scientific inference.
Sampling for Comparison to Criteria
Once criteria have been selected for each indicator variable (e.g., as a minimum, TN, TP, chl a, or
macroalgal biomass as AFDW and a measure of water clarity associated with chl a and, where
appropriate, the addition of dissolved oxygen), States and Tribes will want to develop implementation
procedures to assess the estuarine and coastal water with the criteria. Sampling to evaluate attainment
with criteria and adopted standards should be compatible with the procedures to establish the criteria in
the first place. If the criterion was developed for a particular season, then sampling should be compatible
with that season. In some cases, it is plausible that nutrient concentrations will not correlate predictably
with response variables because of estuarine hydrodynamics. In such cases, the useful relationship
should be directed toward nutrient loading. Many published estuarine nutrient relationships are based on
nutrient load, often normalized to estuarine surface area, not nutrient concentrations. In such cases, the
sampling of load relative to response variable should be scientifically based on the appropriate season
with consideration for appropriate time lags (see Chapter 2). In many cases, relationships will need to be
sought through application of a computer model (see Chapter 9).
Questions will arise about the size of area, depth of sample, frequency, and duration of any exceedances.
These are difficult questions but the RTAG must be prepared to address them. It is expected that
scientific judgment will be required in numerous cases that pushes the state of the science and in some
cases it may be necessary to make risk management decisions extend beyond the current state of science.
Some illustrative examples are provided but should not necessarily be interpreted literally. If an area is
small and does not limit life cycle completion of important species (e.g., deepwater hypoxia that may
serve as a bottleneck to estuarine species migration), then some tolerance is accepted. However, if the
duration and magnitude of noncompliance of a criterion lasts long enough to affect the distribution and
abundance or recruitment of an important species or a key food web component at a designated level
(e.g., 15% reduction in the population of a harvestable size class is estimated based on the best available
judgment), then the criterion needs to be adjusted. In some cases, empirical or computer models will be
required to address many of the more complex relationships.
The question arises, how many replicate samples are needed to obtain an acceptable precision of data in
order to detect differences between sites and changes over time? This depends on the nature of the
Nutrient Criteria—Estuarine and Coastal Waters 7-11
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variability in the variable of interest. Several approaches are available. However, this question involves
both statistical and practical considerations (e.g., cost). General experience suggests that field water
quality sampling will often vary by 20%. With this "rule of thumb," it may not be cost-effective to try to
achieve a lower percent difference. Eckblad (1991) provides some guidance on statistical considerations
in sampling power. The Kendall test with Sen slope estimate (Hirsch et al. 1982) allows the
determination of the number of replicate samples needed to detect a certain percent change in annual
means of a variable or a certain percent trend over a period such as 10 years (see Rivers and Streams
Nutrient Guidance Manual (U.S. EPA 2000b, Appendix A).
7.8 NUTRIENT CRITERIA INTERPRETATION PROCEDURES
However done, a State's or Tribe's nutrient criteria should include a procedural protocol to implement
the newly adopted nutrient criteria. The criteria and procedures should be reviewed by the RTAGs for
concurrence and are subject to further EPA review and approval if submitted as part of State or Tribal
standards.
The initial criteria variables include two causal variables (TN and TP) and two response variables (algal
biomas, e.g., chlorophyll a for phytoplankton and AFDW for macroalgae) and water clarity (e.g., Secchi
depth), and where hypoxia occurs, dissolved oxygen may be added as a third response variable. Failure
to meet either of the causal criteria should be sufficient to indicate a criteria "excursion," and usually the
biological response, as measured by chlorophyll a and Secchi depth, will follow this nutrient trend.
However, if the causal criteria are met but some combination of response criteria are not met, then there
should be some means of determining if the waters in question meet the nutrient criteria. Two suggested
approaches are described below.
Decisionmaking Protocol
One option is to establish a decisionmaking procedure equation all of the criteria. Such a rule might
state: "Both TN and TP causal nutrient criteria must be met, and a least three out of five response criteria
(e.g., water clarity, algal biomass as chlorophyll a or macroalgal biomass as AFDW, DO, seagrass or
SAV biomass, and phytoplankton species composition) must be met for three out of four sampling events
during the June through August survey period over 2 consecutive calendar years of sampling. No
sampling events may be less than 3 weeks apart [to avoid clustering sampling activities near a particular
flow condition or runoff event], and flow and tidal conditions must be recorded as well so that watershed
base flow and runoff events are evident and can be factored into the data assessment process."
Multivariable Enrichment Index
The second option is to establish an index that accomplishes the same result by inserting the data into an
equation that relates the multiple variables in a nondimensional comprehensive score much the same way
an index of biotic integrity (Karr 1981) does. An example of an enrichment index approach is presented
in Table 7-1.
7-12 Nutrient Criteria—Estuarine and Coastal Waters
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Table 7-1. Example of an enrichment index using the middle portion of a hypothetical estuary
Variable
Criterion
Hypothetical estuary or salinity zone of estuary
Median measured value
Enrichment Index (El) score*
Causal variables
Total P (mg/L)
Total N (mg/L)
<2.0
<68
2.5
70
4
5
Primary response variables
Secchi depth (M)
Chlorophyll a (mg/L)
>1.0
<60
0.6
75
3
3
Secondary response variables
Dissolved oxygen (mg/L in
hypoliminion)
>6.0
3.5
4
Enrichment Index Value**: = 19
* Each of the eight variables receives an El score. The scoring procedure is: 0 = meets criterion; 2 = fails to meet criterion by
10%; 3 = fails to meet criterion by 25%; 5 = fails to meet criterion by 50% or more.
** Enrichment Index Value is the sum of the El scores. The maximum i.e. worst score achievable is 25.
If necessary, the scoring process can be weighted by seasons. Thus, different emphasis can be given to
the results of winter surveys as compared with summer surveys, and year-round work can be conducted if
necessary or desired. For example, greater weight perhaps by a factor of 2 could be given to the primary
response variables in winter for north temperate waters because these variables would normally be
expected to be improved at this time of year. Similarly, the criteria for TP and TN might both be
changed to lower concentration for winter because less runoff or fewer fertilizer applications are
expected in the watershed. In the example, the estuary or region thereof fails anyway because it failed
the criterion for either TP or TN (in fact it failed both). With a score of 19 out of a possible 25, it is also
a prime candidate for extensive remediation management.
Such enrichment index scores are not intended at this time to be surrogate nutrient criteria. They may,
however, serve as a "translator" to implement multiparameter criteria. However, like biological criteria
index scores such as the Index of Biotic Integrity, the enrichment index may be a useful assessment tool
relating several parameters. This helps the resource manager plan the distribution of effort and funds
over the entire estuarine or coastal resource base in one procedure.
Frequency and Duration
Frequency and duration are important concerns when evaluating any water with respect to meeting
criteria. This is a difficult process at this initial phase of the program because the data sources for
criteria development are presently so diverse. In general, however, the method of data gathering for
compliance should be as near as possible to that used to establish the criteria, especially keeping in mind
Nutrient Criteria—Estuarine and Coastal Waters
7-13
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tidal phase and salinity. Once consistency is established, excursions from the criteria based on frequency
and duration can be evaluated whether based on a decision rule or a multivariable index.
Frequency of "excursion" from a criterion is a decision that can be best established by the State or Tribe
on the basis of their knowledge of the local water resources. An excursion that occurs less than 10
percent of the times when sampling is conducted (at regularly spaced or random intervals) may be
considered acceptable. Duration of the excursion may be stipulated as a set period of time (e.g., 2 weeks,
or as to not exist over more than two consecutive sampling intervals, whichever is the lesser period). The
State or Tribe in consultation with EPA will need to specifically define these terms as appropriate to the
region and should also determine the combination of these factors that constitutes an "excursion."
7.9 CRITERIA MODIFICATIONS
Some situations may require site-specific criteria because of unique environmental conditions. In such
situations, the general criterion is a starting point and it must be modified to protect designated uses in a
unique situation. Such criteria can be adopted into State or Tribal water quality standards and reviewed
by EPA.
7.10 EPA, STATE, OR TRIBE RESPONSIBILITY UNDER THE CLEAN WATER ACT
The Clean Water Act as amended (Pub. L. 92-500 (1972), 33 U.S.C. 1251, et seq.) requires all States to
establish designated uses for their waters (Section 303(c)). Designated uses are set by the State. EPA's
interpretation of the CWA requires that wherever attainable, standards should provide for the protection
and propagation offish, shellfish, and wildlife and provide for recreation in and on the water (Section
101(a)). Other uses identified in the act include industrial, agricultural, and public water supply.
However, no waters may be designated for use as repositories for pollutants (40 CFR 131.10 [a]). Each
waterbody must have criteria or measures of appropriate water quality that protect and maintain the
designated use of that water. It is recommended that the EPA nutrient guidance be followed. However,
States and Tribes may follow other guidance to adopt water quality criteria as long as the criteria are
based on scientifically sound methods and protect designated uses.
7.11 IMPLEMENTATION OF NUTRIENT CRITERIA INTO WATER
QUALITY STANDARDS
Nutrient criteria adopted into water quality standards by States and Tribes are submitted to EPA for
review and approval (see Section 40 CFR 131). EPA reviews the criteria (40 CFR 131.5) for consistency
with the requirements of the CWA and 40 CFR 131.5, which requires that water quality criteria be
sufficient to protect the designated use (40 CFR 131.6 (c) and 40 CFR 131.11). The procedures for
State/Tribal review and revision of water quality standards, EPA review and approval of water quality
standards, and EPA promulgation of water quality standards (upon disapproval of State/Tribal water
quality standards) are found at 40 CFR 131.20-22. The Water Quality Standards Handbook (U.S. EPA
1994) provides guidance for implementation of these regulations.
7-14 Nutrient Criteria—Estuarine and Coastal Waters
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CHAPTER 8
Using Nutrient Criteria To
Protect Water Quality
Managing Point Source Pollution
Managing Nonpoint Source Pollution
Comprehensive Procedure for Nutrient
Management
Resources
This chapter provides an introduction to the applications of nutrient criteria. Chapter 1 described the
ways in which nutrient criteria are used to (1) identify problems, (2) develop management plan, (3) assess
regulations, (4) evaluate projects, and (5) determine the status and trend of the water resource. In this
applications chapter, some of these are discussed in further detail. Section 8.1 addresses the management
of point source pollution, in the context of standards development, National Pollutant Discharge
Elimination System (NPDES) permits, and total maximum daily loads (TMDLs). Section 8.2 focuses
exclusively on nonpoint source management programs. Although some material is not directly related to
estuarine and coastal marine resources, it is included here because the coastal waters are the ultimate
recipients of all drainage, both coastal and inland, and the information may be useful to a manager
addressing various sources within a watershed. Coastal waters may, of course, be waters of the United
States (see 40 CFR 122.2) and may thus be subject to the requirements of the Clean Water Act. Section
8.3 sets out a comprehensive planning, application, and evaluation procedure for estuarine and coastal
marine nutrient quality management, and Section 8.4 lists publications on coastal/estuarine and
watershed management and protection.
8.1 MANAGING POINT SOURCE POLLUTION
The term "point source" means any discernible, confined, and discrete conveyance, including but not
limited to any pipe, ditch, channel, tunnel, conduit, well, discrete fissure, container, rolling stock,
concentrated animal feeding operation, or vessel or other floating craft, from which pollutants are or may
be discharged (CWA § 502(14)). This term does not include agricultural storm water discharges and
return flows from irrigated agriculture. This section describes some of the programs relevant to point
source discharges into rivers and streams.
The Clean Water Act and Water Quality Standards
The goals of the CWA are to achieve, wherever attainable, water quality that provides for protection and
propagation of fish, shellfish, and wildlife and recreation in and on the water. The CWA further
specifies that States adopt, and EPA approve, water quality standards consisting of designated uses,
criteria to project those uses, and an antidegradation policy (CWA section 303(c)). The criteria must be
based on a sound scientific rationale and must contain sufficient parameters or constituents to protect the
designated uses (40 CFR § 131.1 l(a)). For waters with multiple use designations, criteria must support
the most sensitive use (Id.). Finally, in designating uses and establishing water quality criteria, States
must ensure attainment of standards in downstream waters (40 CFR § 131.10(b)). With regard to
nutrient criteria, Section 304(a) of the CWA directs EPA to develop and publish criteria that reflect the
latest scientific knowledge of the effects of pollutants on biological community diversity, productivity,
Nutrient Criteria—Estuarine and Coastal Waters
8-1
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and stability, including information on the factors affecting rates of eutrophication for varying types of
receiving waters. In establishing water quality criteria, States should establish numeric values to protect
designated uses based on EPA's Section 304(a) criteria guidance, modifications of the guidance
recommendations reflecting site-specific conditions, or criteria based on other scientifically defensible
methods (40 CFR § 131.11(b)(l)).
As illustrated in Figure 8-1, States adopt water quality standards for waters of the United States that
comprise designated uses, criteria to protect those uses, and an antidegradation policy to protect existing
water quality. Additionally, States develop implementation procedures to describe how the water quality
standards will be applied. Once water quality standards are adopted and approved, they become the basis
for legally enforceable NPDES permit limitations and a variety of assessment activities under the Clean
Water Act.
Protecting Designated Uses
It has been amply demonstrated that nutrients are a major contributor to use impairment in waters of the
United States. Because States are required to designate uses in consideration of the goals of the CWA
and adopt criteria that contain sufficient parameters to protect designated uses, and because it is EPA's
responsibility to make related recommendations, the Agency is developing and publishing Section 304(a)
criteria for nutrients that provide for protection and propagation offish, shellfish, and wildlife and
recreation in and on the water.
%
H
Criteria
(narrative or
nurmric
Implementation
Anti-
degradation
Policy
aters
Figure 8-1. Components of water quality standards.
8-2
Nutrient Criteria—Estuarine and Coastal Waters
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EPA's Section 304(a) criteria for nutrients are issued on the basis of ecoregion and waterbody type. This
approach to nutrient criteria development provides a sound, scientifically defensible approach that
accounts for the characteristics of different types and locations of waterbodies. EPA's ecoregional
nutrient criteria are intended to represent enrichment conditions of surface waters minimally affected by
human development. These criteria may be developed and further refined on the basis of the five
elements described in this technical guidance manual.
Water quality criteria incorporating minimally affected (i.e., reference) conditions should provide for
protection and propagation of aquatic life and recreation and reflect conditions that will not adversely
affect the biological community. The parameters addressed in the Ecoregional Nutrient Criteria
Documents are total phosphorus, total nitrogen, chlorophyll a, and turbidity (e.g., Secchi depth for lakes;
turbidity for rivers and streams). These are the parameters that EPA considers important in nutrient
assessment because the first two (nitrogen and phosphorus) are the main causal agents of enrichment,
whereas the two response variables (chlorophyll a and turbidity) are indicators of system overenrichment
for most surface waters.
Maintaining Existing Water Quality
Antidegradation
State and Tribal water quality standards include an antidegradation policy and methods through which
the State or Tribe implements the policy. An antidegradation policy is required in State water quality
standards to protect existing water quality. At a minimum, States must maintain and protect the quality
of waters to support existing uses Antidegradation implementation procedures address the measures used
by States and Tribes to ensure that permits and control programs meet water quality standards and
antidegradation requirements. The water quality standards regulation sets out a three-tiered
antidegradation approach for the protection of water quality (40 CFR § 131.12).
Tier 1
Maintains and protects existing uses and the water quality necessary to protect these uses (40 CFR
131.12[a] [1]). An existing use can be established by demonstrating that fishing, swimming, or other uses
have actually occurred since November 28, 1975, or that the water quality is suitable to allow such uses
to occur, unless there are physical problems, such as substrate or flow, that prevent the use from being
attained (U.S. EPA 1994).
Tier 2
Protects the water quality in waters whose quality is better than that necessary to protect
"fishable/swimmable" uses (40 CFR 131.12[a][2]). The water quality standards regulation requires that
certain procedures be followed and certain showings be made (an "antidegradation review") before a
point source is authorized to lower water quality in high-quality waters. In no case may water quality for
a tier 2 waterbody be lowered to a level at which existing uses are impaired.
Nutrient Criteria—Estuarine and Coastal Waters 8-3
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Tier 3
Outstanding national resource waters (ONRWs) are provided the highest level of protection under the
antidegradation policy. The policy provides for protection of water quality in high-quality waters that
constitute an ONRW by prohibiting the lowering of water quality. ONRWs are often regarded as highest
quality waters of the United States: That is clearly the thrust of 131.12(a)(3). However, ONRW
designation also offers special protection for waters of "exceptional ecological significance." These are
waterbodies that are important, unique, or sensitive ecologically, but whose water quality, as measured
by the traditional parameters such as dissolved oxygen or pH, may not be particularly high or whose
characteristics cannot be adequately described by these parameters (such as wetlands).
The regulation requires water quality to be maintained and protected in ONRWs. EPA interprets this
provision to mean no new or increased discharges to ONRWs and no new or increased discharge to
tributaries to ONRWs that would result in lower water quality in the ONRWs. The only exception to this
prohibition, as discussed in the preamble to the Water Quality Standards Regulation (48 FR 51402),
permits States to allow some limited activities that result in temporary and short-term changes in the
water quality of ONRW. Such activities must not permanently degrade water quality or result in water
quality lower than that necessary to protect the existing uses in the ONRW. It is difficult to give an exact
definition of "temporary" and "short-term" because of the variety of activities that might be considered.
However, in rather broad terms, EPA's view of temporary is weeks and months, not years. The intent of
EPA's provision clearly is to limit water quality degradation to the shortest possible time. If a
construction activity is involved, for example, temporary is defined as the length of time necessary to
construct the facility and make it operational. During any period of time when, after opportunity for
public participation in the decision, the State allows temporary degradation, all practical means of
minimizing such degradation shall be implemented. Examples of situations in which flexibility is
appropriate can be found in the Water Quality Standards Handbook (U.S. EPA 1994).
General Policies
The water quality standards regulation allows States and Tribes to include implementation in their
standards policies and provisions, such as mixing zones, variances, and low-flow exemptions (40 CFR §
131.13). Such policies are subject to EPA review and approval. These policies and provisions should be
specified in the State's or Tribe's water quality standards document. The rationale and supporting
documentation should be submitted to EPA for review during the water quality standards review and
approval process.
Mixing Zones
States and Tribes may, at their discretion, allow mixing zones for dischargers. The water quality
standards should describe the methodology for determining the location, size, shape, outfall design, and
in-zone quality of mixing zones. Careful consideration should be given to the appropriateness of a
mixing zone where a substance discharged is bioaccumulative, persistent, carcinogenic, mutagenic, or
teratogenic.
8-4 Nutrient Criteria—Estuarine and Coastal Waters
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Low-Flow Provisions
State and Tribal water quality standards should protect water quality for the designated and existing uses
in critical low-flow situations. States and Tribes may, however, designate a critical low-flow below
which numerical water quality criteria do not apply. When reviewing standards, States and Tribes should
review their low-flow provisions for conformance with EPA guidance.
Water Quality Standards Variances
Variance procedures involve the same substantive and procedural requirements as removing a designated
use (see 40 CFR 131.10 (g)), but unlike use removal, variances are both discharger and pollutant specific,
are time-limited, and do not forego the currently designated use.
A variance should be used instead of removal of a use where the State believes the standard can
ultimately be attained. By maintaining the standard rather than changing it, the State will assure that
further progress is made in improving water quality and attaining the standard. With a variance, NPDES
permits may be written such that reasonable progress is made toward attaining the standards without
violating section 402(a)(l) of the Act, which requires that NPDES permits must meet the applicable
water quality standards.
State variance procedures, as part of State water quality standards, must be consistent with the
substantive requirements of 40 CFR 131. EPA has approved State-adopted variances in the past and will
continue to do so if:
Each individual variance is included as part of the water quality standard
• The State demonstrates that meeting the standard is unattainable based on one or more of the grounds
outlined in 40 CFR 131.10(g) for removing a designated use
• The justification submitted by the State includes documentation that treatment more advanced
than that required by sections 303(c)(2)(A) and (B) has been carefully considered, and that
alternative effluent control strategies have been evaluated
The more stringent State criterion is maintained and is binding upon all other dischargers on the
stream or stream segment
• The discharger who is given a variance for one particular constituent is required to meet the
applicable criteria for other constituents
• The variance is granted for a specific period of time and must be rejustified upon expiration but at
least every 3 years (Note: the 3-year limit is derived from the triennial review requirements of section
303(c) of the Act.)
Nutrient Criteria—Estuarine and Coastal Waters 8-5
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• The discharger either must meet the standard upon the expiration of this time period or must make a
new demonstration of "unattainability"
• Reasonable progress is being made toward meeting the standards
The variance was subjected to public notice, opportunity for comment, and public hearing. (See
section 303(c)(l) and 40 CFR 131.20.) The public notice should contain a clear description of the
impact of the variance on achieving water quality standards in the affected stream segment.
Providing Flexibility in Implementation
Abundant flexibility is built into the criteria-setting process and water quality standards regulations to
allow States to (1) develop their own criteria to protect specific uses or reflect local conditions, (2) use
different techniques to develop criteria as long as they are protective and scientifically defensible, and (3)
conduct use attainability studies and refine their use designations.
States also have the flexibility to adopt numeric criteria to protect designated uses or adopt methods and
procedures that "translate" narrative criteria into numeric values. Narrative criteria statements, often
referred to as "general criteria" in States' standards regulations, usually take the form of a description of
desired water quality condition or a preclusion of certain types of pollution or undesirable conditions
(i.e., the "free from" provisions). Narrative criteria are considered critical backstops for designated use
protection and are a powerful means of achieving desired water quality if they are interpreted in a clear
and consistent manner. In water quality standards parlance, a "translator" is a process, methodology, or
guide that States or Tribes use to quantitatively interpret narrative criteria statements. Translators may
consist of biological assessment methods (e.g., field measures of the biological community), biological
monitoring methods (e.g., laboratory toxicity tests), models or formulas that use input of site-specific
information/data, or other scientifically defensible methods. Translators are particularly useful in
describing water quality conditions that require a greater degree of sophistication to assess than typically
can be expressed by numerical criteria that apply broadly to all waters with a given use designation. The
translator may be either directly incorporated into State or Tribal water quality standards or incorporated
by reference. In either case, specific limits or values for a measurable pollutant derived using a translator
that interpret a narrative criterion statement should be attached to the State or Tribal regulations to ensure
public review, as would be required of any site-specific numerical criterion.
States have the flexibility under current law to adopt appropriate nutrient criteria. If a State determines
that its water quality criteria cannot be met, that State may consider refining use designation, adopting
site-specific criteria, or issuing a variance to ensure that the appropriate uses and criteria to protect the
uses are established. For example, if a regulated source faces expensive treatment to comply with a new
or revised requirement, the State or Tribe can authorize a variance, based on a justification using one of
the six factors at 131.10(g), to allow time for the discharge to come into compliance with a permit limit
for the criteria and/or weigh options on treatment technologies, use reclassification, or site-specific
development criteria as appropriate.
8-6 Nutrient Criteria—Estuarine and Coastal Waters
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Water quality criteria guidance published by EPA under Section 304(a) of the CWA, such as for
nutrients, serve as primary sources of information to States and Tribes as they develop numeric criteria as
part of their water quality standards. Under the CWA and EPA's implementing regulations, States and
Tribes also may use other information, including local water quality conditions, as they develop
standards. Typically, EPA uses its own water quality criteria guidance as the principal basis for
proposing and promulgating a replacement water quality standard when a State or Tribe fails to adopt an
acceptable standard. In doing so, EPA commits to a process that includes public review and comment.
EPA will solicit information from the public to determine if any such proposed Federal nutrient criteria
for State waters are sufficiently protective of uses. This public process will help ensure that any
promulgated Federal water quality standards are appropriately protective.
NPDES Permits
The CWA requires wastewater dischargers to have a permit establishing effluent limits on pollutant
discharges. The regulations at 40 CFR 122.41 et seq. require these permits to specify monitoring and
reporting requirements. More than 200,000 sources are regulated by the NPDES permits nationwide.
These permits regulate household and industrial wastes that are collected in sewers and treated at
municipal wastewater treatment plants. Permits also regulate industrial point sources and concentrated
animal feeding operations that discharge into other wastewater collection systems or that have the
potential to discharge directly into receiving waters. Permits regulate discharges with the goals of
protecting public health and aquatic life and ensuring that every facility treats wastewater. Typical
pollutants regulated by NPDES are "conventional pollutants" such as fecal coliforms or oil and grease
from the sanitary wastes of households, businesses, and industries; and "toxic pollutants" including
pesticides, solvents, polychlorinated biphenyls (PCBs), dioxins, and heavy metals that are particularly
harmful to animal or plant life. "Nonconventional pollutants" are any additional substances that are not
conventional or toxic that may require regulation, including nutrients such as N and P.
[Source: http://www.epa.gov/owm/gen2.htm]
Discharge monitoring data for pollutants limited and/or monitored pursuant to NPDES permits issued by
States, Tribes, or EPA are required to be stored in the central EPA Permit Compliance System (PCS).
The assessment of point source loadings is not a simple process of assessing PCS data, even though PCS
is an important data source. The PCS database does not provide complete information for important N
sources. Most PCS N data are generated by water quality-based permit limitations on ammonia, often
applied in discharges to smaller streams. Few data exist in PCS on other forms of N, or TN; and data for
TP is not frequently found in PCS. This situation exists largely because most permits do not include
limits and/or monitoring requirements for N or P. The lack of nutrient limits and/or monitoring
requirements in permits is due to a general lack of State water quality standards for these parameters.
[Source: http://www.epa.gov/msbasin/protocol.html]
The NPDES Storm Water Permitting Program
Storm water runoff is one of the remaining causes of contaminated lakes, streams, rivers, and estuaries
throughout the country. Pollution in storm water runoff is responsible for closing beaches and shellfish
harvesting areas, contaminating fish, and reducing populations of water plants and other aquatic life.
Nutrient Criteria—Estuarine and Coastal Waters 8-7
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High flows of storm water runoff cause flooding, property damage, erosion, and heavy siltation. The
1987 Congressional Amendments to the Clean Water Act required EPA to control pollution from storm
water discharges. In 1990, EPA promulgated Phase I regulations to control storm water discharges from
municipal separate storm sewer systems (MS4s) serving populations of more than 100,000, construction
activities disturbing more than 5 acres, and industrial facilities through issuance of NPDES storm water
permits. EPA promulgated Phase II of the program in 1999 to control storm water discharges from MF4s
less than 100,000 and small construction sites between 1 and 5 acres in size. The Phase II regulations
also expanded the exemption for industrial facilities that do not have exposure of industrial activities and
materials to storm water.
Construction Permits
The 1987 Congressional Amendments to the CWA required EPA to control pollution from storm water
discharges. EPA issued a general NPDES permit for construction sites disturbing 5 or more acres in
1992. General permits provide EPA with an effective mechanism to regulate discharges from tens of
thousands of construction sites, thus protecting and improving surface water quality across the Nation.
Several general permits for construction activity have been issued/reissued since that first permit in 1992.
EPA Regions 1, 2, 3, 7, 8, 9, and 10 have a general permit that authorizes the discharge of storm water
associated with construction activity disturbing 5 or more acres and smaller sources as designated by the
Agency on a case-by-case basis. This multiregional permit is known as the "Construction General
Permit" (CGP).
Region 4 has issued a separate CGP for the State of Florida and Indian Country lands in Florida,
Mississippi, Alabama, and North Carolina. Region 6 has also issued its own CGP for the States of Texas
and New Mexico; Indian Country lands in Texas, New Mexico, Oklahoma, and Louisiana; and
construction activity at oil, gas, and pipeline facilities in Oklahoma.
As used in these construction permits, the term "storm water associated with construction activity" refers
to category (x) of the definition of "discharge of storm water associated with industrial activity," which
includes construction sites and common plans of development or sale that disturb 5 or more acres (see 40
CFR 122.26 [b][14]). The CGP applies only to areas for which EPA is the permitting authority (certain
States, Federal facilities, and Indian Lands). The majority of the country (i.e., 44 States and the Virgin
Islands) has been granted authority for permitting storm water discharges and as such, each of these
States is required to develop permits to control discharges from construction activities. In response to the
Phase II regulations, permit applications from construction activities between 1 and 5 acres is required by
March 2003.
Combined Sewer Overflows (CSOs)
Combined sewer overflows, or CSOs, can be a significant water pollution and public health threat in
urban areas. EPA's 1994 CSO Control Policy is a comprehensive national strategy to ensure that cities,
NPDES authorities, water quality standards authorities, and the public engage in a comprehensive and
coordinated planning effort to implement cost-effective CSO controls that meet the objectives and
requirements of the Clean Water Act.
8-8 Nutrient Criteria—Estuarine and Coastal Waters
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During dry weather, combined sewer systems transport wastewater directly to sewage treatment plants.
In periods of rainfall or snowmelt, however, the wastewater volume in a combined sewer system can
exceed the capacity of the collection system or treatment plant. When this happens, combined sewer
systems overflow and discharge untreated wastewater directly to streams, rivers, lakes, or estuaries.
It provides guidance to municipalities and State and Federal permitting authorities on meeting pollution
control goals of the CWA in a flexible, cost-effective manner. Information on EPA's CSO Control
Policy, accompanying guidance documents, and other elements of the national CSO control program can
be found on the following website: http://www.epa.gov/npdes.htm.
Storm Water Planning
The Watershed Management Institute, Inc., recently published a new manual entitled Operation,
Maintenance, and Management of Stormwater Management Systems (1998). This manual presents a
comprehensive review of the technical, educational, and institutional elements needed to ensure that
storm water management systems are designed, built, maintained, and operated properly during and after
their construction. The manual was developed in cooperation with the EPA Office of Water to assist
individuals responsible for designing, building, maintaining, or operating storm water management
systems. It will also be helpful to individuals responsible for implementing urban storm water
management programs.
The manual includes fact sheets on 13 common storm water treatment best management practices
(BMPs) that summarize operation, maintenance, and management needs and obligations, along with
construction recommendations. Other chapters review planning and design considerations, programmatic
and regulatory aspects, considerations for facility owners, construction inspection, inspection and
maintenance after construction, costs and financing, and disposal of storm water sediments. Forms for
inspecting BMPs during construction and determining maintenance needs afterwards are included in the
manual and in a separate supplement.
[Source: http://www.epa.gov/owowwtrl/NPS/wmi/index.html]
[Additional information: http://www.epa.gov/owowwtrl/NPS/ordinance/osm6.htm and
http://www.epa.gov/owowwtrl/info/NewsNotes/issue05/nps05sto.html]
Total Maximum Daily Load (TMDL)
States, territories, and authorized Tribes establish section 303(d) lists of impaired waters based on
information contained in their 305(b) reports as well as other relevant and available water quality data.
The section 303(d) list is a prioritized list of waters not meeting water quality standards. States are
required to submit lists biennially (40 CFR § 130.7(d)). States must develop TMDLs for waters and
pollutants on their section 303(d) lists (CWA § 303(d)(l)(C)). State section 303(d) lists and TMDLs
must be submitted to EPA for approval or disapproval. If EPA disapproves a list or TMDL submission,
EPA must identify waters for the list or establish the TMDL itself (CWA § 303(d)(2)).
A TMDL is a written, quantitative plan and analysis for attaining and maintaining water quality standards
in all seasons for a specific waterbody and pollutant. Specifically, a TMDL is the sum of the allowable
Nutrient Criteria—Estuarine and Coastal Waters 8-9
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loads of a pollutant from all contributing point, nonpoint, and background sources (40 CFR § 130.2(i)).
TMDLs may be established on a coordinated basis for a group of waterbodies in a watershed. TMDLs
identify the loading capacity of the water, wasteload allocations (for point sources), load allocations (for
nonpoint sources), and a margin of safety (U.S. EPA 1999) (40 CFR § 130.2), and are calculated at levels
necessary to achieve applicable water quality standards (CWA § 303(d)(l)(C)).
A waste load allocation (WLA) is the proportion of a receiving water's TMDL that is allocated to point
sources of pollution. Water quality models are often utilized by regulatory agencies in conducting an
assessment to determine a WLA. Models establish a quantitative relationship between a waste load and
its impact on water quality. WLAs are used by permit writers to establish Water Quality Based Effluent
Limits (WQBELs).
[Source: http://www.epa.gov:80/owmitnet/permits/pwcourse/chapt_06.pdfj
Both the 1996 and 1998 section 303(d) lists, as well as more recent 305(b) reports, reflect similar
patterns: sediments, nutrients, and pathogens are the top three causes of waterbody impairment.
[Source: http://www.epa.gov/owowwtrl/tmdl/faq.html]
Continuing Planning Process (CPP)
Each State is required to establish and maintain a continuing planning process (CPP) as described in
section 303(e) of the CWA. A State's CPP contains, among other items, a description of the process that
the State uses to identify waters needing water quality-based controls, the process for developing
TMDLs, and a description of the process used to receive public review of each TMDL (40 CFR § 130.5
& 130.7(a) & (c)). Descriptions may be as detailed as the Regional office and the State determine is
necessary to describe each step of the TMDL development process. This process may be included as part
of the EPA/State Agreement for TMDL development.
[Source: http://www.epa.gov/owowwtrl/tmdl/decisions/dec4.html]
Look to the Future ... Pollutant Trading
Point and nonpoint source pollutant trading involves financing reductions in nonpoint source pollution in
lieu of undertaking more expensive point source pollution reduction efforts. A trading program is
intended to produce cost savings for point source dischargers while improving water quality. In order for
a trading program to be viable, there should be a waterbody identifiable as a watershed or segment, as
well as a measurable combination of point sources and controllable nonpoint sources. In addition, point
source dischargers can be expected to trade for nonpoint source reductions if they perceive this as an
alternative to upgrade facility treatment capabilities. In addition, there should be significant load
reductions for which the cost per pound reduced for nonpoint source controls is lower than the cost for
upgrading point source controls.
Such a program allows the private sector to allocate its resources to reduce pollutants in the most
cost-effective manner, and it encourages the development of a watershed-wide or basin-wide
approach to water quality protection. A pollutant trading program also requires cooperation between
agencies and requires a system to arrive at trading ratios between point and nonpoint source controls.
8-10 Nutrient Criteria—Estuarine and Coastal Waters
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For example, in a North Carolina watershed, the Tar-Pamlico Basin Association (a coalition of point
source dischargers) and State and regional environmental groups have proposed a two-phased nutrient
management strategy that incorporates point and nonpoint source pollutant trading. The plan requires
association members to finance nonpoint source reduction activities in the basin if their nutrient
discharges exceed a base allowance.
[Source: http://www.epa.gov/OWOW/NPS/MMGI/funding.html#9]
8.2 MANAGING NONPOINT SOURCE POLLUTION
During the first 15 years of the national program to abate and control water pollution, EPA and the States
focused most of their water pollution control activities on traditional "point sources," such as discharges
through pipes from sewage treatment plants and industrial facilities. These point sources have been
regulated by EPA and the States through the NPDES permit program established by section 402 of the
CWA. Discharges of dredged and fill materials into wetlands have also been regulated by the U.S. Army
Corps of Engineers and EPA under section 404 of the Clean Water Act.
The Nation has greatly reduced pollutant loads from point source discharges and has made considerable
progress in restoring and maintaining water quality as a result of the above activities. However, the gains
in controlling point sources have not solved all of the Nation's water quality problems. Recent studies
and surveys by EPA and by State/Tribal water quality agencies indicate that the majority of the
remaining water quality impairments in our Nation's rivers, streams, lakes, estuaries, coastal waters, and
wetlands result from nonpoint source pollution and other nontraditional sources, such as urban storm
water discharges and combined sewer overflows.
Nonpoint source pollution generally results from land runoff, precipitation, atmospheric deposition,
drainage, seepage, or hydrologic modification. Technically, the term "nonpoint source" is defined to
mean any source of water pollution that does not meet the legal definition of "point source" in section
502(14) of the CWA, defined in the preceding section. Although diffuse runoff is generally treated as
nonpoint source pollution, runoff that enters and is discharged from conveyances such as those described
above is treated as a point source discharge and hence is subject to the permit requirements of the Clean
Water Act. In contrast, nonpoint sources are not subject to Federal permit requirements.
The pollution of waters by nonpoint sources is caused by rainfall or snowmelt moving over and through
the ground. As the runoff moves, it picks up and carries away natural pollutants and pollutants resulting
from human activity, finally depositing them into lakes, rivers, wetlands, coastal waters, and
groundwaters. Nonpoint source pollution can also be caused by atmospheric deposition of pollutants
onto waterbodies. Furthermore, hydrologic modification is a form of nonpoint source pollution that often
adversely affects the biological and physical integrity of surface waters. A more detailed discussion of
the range of nonpoint sources and their effects on water quality and riparian habitats is provided in
subsequent chapters of this guidance. A summary of State laws related to nonpoint source pollution can
be found in the Almanac of Enforceable State Laws to Control Nonpoint Source Water Pollution (ELI
1988). This report can be accessed on the Internet at http://www.eli.org/bookstore/research.htm.
Nutrient Criteria—Estuarine and Coastal Waters 8-11
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Nonpoint Sources of Nutrients
Guidance Specifying Management Measures for Sources of Nonpoint Pollution in Coastal Waters (U.S.
EPA 1993a) was developed by EPA for the planning and implementation of Coastal Nonpoint Pollution
Programs. The guidance focuses on controlling five major categories of nonpoint sources that impair or
threaten waters nationally. Management measures are specified for (1) agricultural runoff, (2) urban
runoff (including developing and developed areas), (3) silvicultural (forestry) runoff, (4) marinas and
recreational boating, and (5) hydromodification (e.g., channelization and channel modification, dams,
and streambank and shoreline erosion). EPA guidance also includes management measures for wetlands,
riparian areas, and vegetated treatment systems that apply generally to various categories of sources of
nonpoint pollution. Management measures are defined in the Coastal Zone Act Reauthorization
Amendments of 1990 (CZARA) as economically achievable measures to control the addition of
pollutants to waters, which reflect the greatest degree of pollutant reduction achievable through the
application of the best available nonpoint pollution control practices, technologies, processes, siting
criteria, operating methods, or other alternatives.
The following section outlines some of the management measures specified in the CZARA guidance for
the various types of nonpoint sources. These measures should be considered when implementing
programs targeting nutrient releases into waters of the United States.
Agricultural Runoff
• Erosion and sediment control
Control of facility wastewater and runoff from confined animal facilities
Nutrient management planning on cropland
• Grazing management systems
• Irrigation water management
Urban Runoff
• Control of runoff and erosion from existing and developing areas
Construction site runoff and erosion control
• Construction site chemical control (includes fertilizers)
Proper design, location, installation, operation, and maintenance of on-site disposal systems
Pollution prevention education (e.g., household chemicals, lawn and garden activities, golf courses,
pet waste, on-site disposal systems, etc.)
Planning, siting, and developing roads, highways, and bridges (including runoff management)
Silvicultural Runoff
• Streamside management
• Road construction and management
Forest chemical management (includes fertilizers)
• Revegetation
• Preharvest planning, harvesting management
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Marinas and Recreational Boating
• Siting and design
• Operation and maintenance
Storm water runoff management
• Sewage facility management
Fish waste management
Pollution prevention education (e.g., proper boat cleaning, fish waste disposal, and sewage pumpout
procedures)
Hydromodification (i.e., channelization, channel modification, dams)
• Minimize changes in sediment supply and pollutant delivery rates through careful planning and
design
Erosion and sediment control
• Chemical and pollutant control (includes nutrients)
Stabilization and protection of eroding streambanks or shorelines
Wetlands, Riparian Areas, Vegetated Treatment Systems
• Protect the NPS abatement and other functions of wetlands and riparian areas through vegetative
composition and cover, hydrology of surface and groundwater, geochemistry of the substrate, and
species composition
• Promote restoration of preexisting function of damaged and destroyed wetlands and riparian systems
Promote the use of engineered vegetated treatment systems if they can serve a NPS pollution
abatement function
Efforts To Control Nonpoint Source Pollution
Efforts to control nonpoint source pollution include nonpoint source management programs, the National
Estuary Program, atmospheric deposition, coastal nonpoint pollution control programs, and Farm Bill
conservation provisions. These efforts are described below.
Nonpoint Source Management Programs
In 1987, in view of the progress achieved in controlling point sources and the growing national
awareness of the increasingly dominant influence of nonpoint source pollution on water quality,
Congress amended the Clean Water Act to focus greater national efforts on nonpoint sources. In the
Water Quality Act of 1987, Congress amended section 101, "Declaration of Goals and Policy," to add the
following fundamental principle:
It is the national policy that programs for the control of nonpoint sources of
pollution be developed and implemented in an expeditious manner so as to enable
the goals of this Act to be met through the control of both point and nonpoint sources
of pollution.
More importantly, Congress enacted section 319 of the CWA, which established a national program to
control nonpoint sources of water pollution. Under section 319, States address nonpoint pollution by
Nutrient Criteria—Estuarine and Coastal Waters 8-13
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assessing nonpoint source pollution problems and causes within the State, adopting management
programs to control the nonpoint source pollution, and implementing the management programs.
Although not required, many States have incorporated the management measures specified in the 1993
CZARA guidance into their State Nonpoint Source Management Programs.
Section 319 also authorizes EPA to issue grants to States to assist them in implementing those
management programs or portions of management programs that have been approved by EPA. As of FY
2000, more than $1 billion in grants have been given to States, Territories, and Tribes for the
implementation of nonpoint source pollution control programs.
For additional information on the Nonpoint Source Management Program and distribution of Section 319
grants in your State, contact your State's designated nonpoint source agency. For many States, the
nonpoint source agency is the State Water Quality Agency. However, in several instances, other
agencies or departments are given nonpoint source responsibility (see Table 8-1).
National Estuary Program
EPA also administers the National Estuary Program under section 320 of the CWA. This program
focuses on point and nonpoint pollution in geographically targeted, high-priority estuarine waters. Under
this program, EPA assists State, regional, and local governments in developing comprehensive
conservation and management plans that recommend priority corrective actions to restore estuarine water
quality, fish populations, and other designated uses of the waters. For additional information, contact
your local estuary program. The following estuaries are currently enrolled in the program:
Table 8-1. States for which the nonpoint source agency is not the water quality agency
State
State Nonpoint Source Agency
Arkansas
Delaware
Oklahoma
Tennessee
Texas
Vermont
Virginia
State Department of Soil and Water Conservation
State Department of Soil and Water Conservation
State Department of Soil and Water Conservation
State Department of Agriculture
Department of Soil and Water Conservation (for agriculture)
Texas Water Quality Board (all other nonpoint sources)
State Department of Agriculture
State Department of Soil and Water Conservation
8-14
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Albemarle-Pamlico Sounds, NC
Barataria-Terrebonne Estuarine Complex, LA
Barnegat Bay, NJ
Buzzards Bay, MA
Casco Bay, ME
Charlotte Harbor, FL
(Lower) Columbia River Estuary, OR and WA
Corpus Christi Bay, TX
Delaware Estuary, DE, NJ, and PA
Delaware Inland Bays, DE
Galveston Bay, TX
Indian River Lagoon, FL
Long Island Sound, NY and CT
Maryland Coastal Bays, MD
Massachusetts Bays, MA
Mobile Bay, AL
Morro Bay, CA
Narragansett Bay, RI
New Hampshire Estuaries, NH
New York-New Jersey Harbor, NY and NJ
Peconic Bay, NY
Puget Sound, WA
San Francisco Estuary, CA
San Juan Bay, PR
Santa Monica Bay, CA
Sarasota Bay, FL
Tampa Bay, FL
Tillamook Bay, OR
Atmospheric Deposition
Even though runoff from agricultural and urban areas may be the largest source of nonpoint pollution,
growing evidence suggests that atmospheric deposition may have a significant influence on nutrient
enrichment, particularly from nitrogen (Jaworski et al. 1997). Gases released through fossil fuel
combustion and agricultural practices are two major sources of atmospheric N that may be deposited in
waterbodies (Carpenter et al. 1998). Nitrogen and nitrogen compounds formed in the atmosphere return
to the earth as acid rain or snow, gas, or dry particles (http://www.epa.gov/acidrain/effects/envben.html).
EPA has several programs that address the issue of atmospheric deposition, including the National
Ambient Air Quality Standards, the Atmospheric Deposition Initiative, and the Great Waters Program.
National Ambient Air Quality Standards
The Clean Air Act provides the principal framework for national, State, and local efforts to protect air
quality. Under the Clean Air Act, national ambient air quality standards (NAAQS) for pollutants that are
considered harmful to people and the environment are established.
The Clean Air Act established two types of national air quality standards. Primary standards set limits to
protect public health, including the health of "sensitive" populations such as asthmatics, children, and the
elderly. Secondary standards set limits to protect public welfare, including protection against decreased
visibility and damage to animals, crops, vegetation, and buildings (http://www.epa.gov/airs/criteria.html).
Atmospheric Deposition Initiative
In 1995, EPA's Office of Water established an "Air Deposition Initiative" to work with the EPA Office of
Air and Radiation to identify and characterize air deposition problems with greater certainty and examine
solutions to address them. The Air and Water Programs are cooperating to assess the atmospheric
deposition problem, conduct scientific research, provide innovative solutions to link Clean Air Act and
Clean Water Act tools to reduce the of these pollutants, and communicate the findings to the public. To
date, most efforts have focused on better understanding of the links between nitrogen and mercury
Nutrient Criteria—Estuarine and Coastal Waters
8-15
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emissions and harmful effects on water quality and the environment. Significant work has also been done
towards quantifying the benefits to water quality of reducing air emissions and developing sensible, cost-
effective approaches to reducing the emissions and their ecosystem and health effects
(http://www.epa.gov/owowwtrl/oceans/airdep/index.html).
Great Waters Program
On November 15, 1990, in response to mounting evidence that air pollution contributes to water
pollution, Congress amended the Clean Air Act and included provisions that established research and
reporting requirements related to the deposition of hazardous air pollutants to the "Great Waters." The
waterbodies designated by these provisions are the Great Lakes, Lake Champlain, and Chesapeake Bay.
As part of the Great Waters Program, Congress requires EPA, in cooperation with the National Oceanic
and Atmospheric Administration, to monitor hazardous pollutants by establishing sampling networks,
investigate the deposition of these pollutants, improve monitoring methods, monitor for hazardous
pollutants in fish and wildlife, determine the contribution of air pollution to total pollution in the Great
Waters, evaluate any adverse effects to public health and the environment, determine sources of
pollution, and provide a report to Congress every 2 years. These reports provide an information base that
can be used to establish whether air pollution is a significant contributor to water quality problems of the
Great Waters, determine whether there are significant adverse effects to humans or the environment,
evaluate the effectiveness of existing regulatory programs in addressing these problems, and assess
whether additional regulatory actions are needed to reduce atmospheric deposition to the Great Waters.
For more detail, the Great Waters biennial reports to Congress discuss current scientific understanding of
atmospheric deposition (http://www.epa.gov/airprogm/oar/oaqps/gr8water/xbrochure/program.html).
Coastal Nonpoint Pollution Control Programs
In November 1990, Congress enacted CZARA. These amendments were intended to address several
concerns, a major one being the impact of nonpoint source pollution on coastal waters.
To address more specifically the impacts of nonpoint source pollution on coastal water quality, Congress
enacted section 6217, "Protecting Coastal Waters," which was codified as 16 USC-1455b. This section
provides that each State with an approved coastal zone management program must develop and submit a
Coastal Nonpoint Pollution Control Program for EPA and National Oceanic and Atmospheric
Administration (NOAA) approval. The purpose of the program "shall be to develop and implement
management measures for nonpoint source pollution to restore and protect coastal waters, working in
close conjunction with other State and local authorities."
States with Coast Nonpoint Pollution Control Programs are required to include measures in their
programs that are "in conformity" with the 1993 CZARA guidance, as discussed previously. A listing of
States with Coastal Nonpoint Pollution Control Programs is presented in Table 8-2. For additional
information on the programs in these States, contact the State water quality agency.
8-16 Nutrient Criteria—Estuarine and Coastal Waters
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Table 8-2. States and Territories with coastal nonpoint pollution control programs
Alabama
Alaska
American Samoa
California
Connecticut
Delaware
Florida
Guam
Hawaii
Louisiana
Maine
Maryland
Massachusetts
Michigan
Mississippi
New Hampshire
New Jersey
New York
North Carolina
Northern Mariana Islands
Oregon
Pennsylvania
Puerto Rico
Rhode Island
South Carolina
Virgin Islands
Virginia
Washington
Wisconsin
Farm Bill Conservation Provisions
Technical and financial assistance for landowners seeking to preserve soil and other natural resources is
authorized by the Federal Government under provisions of the Food Security Act (Farm Bill). Provisions
of the 1996 Farm Bill relating directly to installation and maintenance of BMPs are summarized in the
following sections. Contact your Natural Resources Conservation Service (NRCS) State
Conservationist's office for State-specific information.
Environmental Conservation Acreage Reserve Program (ECARP)
ECARP is an umbrella program established by the 1996 Farm Bill and contains the Conservation Reserve
Program (CRP), Wetlands Reserve Program (WRP), and Environmental Quality Incentives Program
(EQIP). It authorizes the Secretary of Agriculture to designate watersheds, multi-State areas, or regions
of special environmental sensitivity as conservation priority areas eligible for enhanced Federal
assistance. Assistance in priority areas is to be used to help agricultural producers comply with NPS
pollution requirements of the CWA and other State or Federal environmental laws. The ECARP is
authorized through 2002.
Conservation Reserve Program (CRP)
First authorized by the Food Security Act of 1985 (Farm Bill), this voluntary program offers annual
rental payments, incentive payments, and cost-share assistance for establishing long-term,
resource-conserving cover crops on highly erodible land. CRP contracts are issued for a duration of 10 to
15 years for up to 36.4 million acres of cropland and marginal pasture. Land can be accepted into the
CRP through a competitive bidding process through which all offers are ranked using an environmental
benefits index, or through continuous signup for eligible lands where certain special conservation
practices will be implemented.
The Conservation Reserve Enhancement Program (CREP) is a new initiative of CRP authorized under
the 1996 Federal Agricultural Improvement and Reform Act. CREP is a joint, State-Federal program
designed to meet specific conservation objectives. CREP targets State and Federal funds to achieve
Nutrient Criteria—Estuarine and Coastal Waters 8-17
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shared environmental goals of national and State significance. The program uses financial incentives to
encourage farmers and ranchers to voluntarily protect soil, water, and wildlife resources.
Wetlands Reserve Program (WRP)
The WRP is a voluntary program to restore and protect wetlands and associated lands. Participants may
sell a permanent or 30-year conservation easement or enter into a 10-year cost-share agreement with
USDA to restore and protect wetlands. The landowner voluntarily limits future use of the land, yet
retains private ownership. The NRCS provides technical assistance in developing a plan for restoration
and maintenance of the land. The landowner retains the right to control access to the land and may lease
the land for hunting, fishing, and other undeveloped recreational activities.
Environmental Quality Incentives Program (EQIP)
The EQIP was established by the 1996 Farm Bill to provide a voluntary conservation program for
farmers and ranchers who face serious threats to soil, water, and related natural resources. EQIP offers
financial, technical, and educational help to install or implement structural, vegetative, and management
practices designed to conserve soil and other natural resources. Current priorities for these funds dictate
that one-half of the available monies be directed to livestock-related concerns. Cost-sharing may pay up
to 75% of the costs for certain conservation practices. Incentive payments may be made to encourage
producers to perform land management practices such as nutrient management, manure management,
integrated pest management, irrigation water management, and wildlife habitat management.
Wildlife Habitat Incentives Program (WHIP)
This program is designed for parties interested in developing and improving wildlife habitat on private
lands. Plans are developed in consultation with NRCS and the local Conservation District. USDA will
provide technical assistance and share up to 75% of the cost of implementing the wildlife conservation
practices. Participants generally must sign a 5- to 10-year contract with USDA that requires they
maintain the improvement practices.
Forestry Incentives Program (FIP)
Originally authorized in 1978, the FIP allows cost sharing of up to 65% (up to a maximum of $10,000 per
person per year) for tree planting, timber stand improvement, and related practices on nonindustrial
private forest land. The FIP is administered by NRCS and the U.S. Forest Service. Cost share funds are
restricted to individuals who own no more than 1,000 acres of eligible forest land.
Conservation of Private Grazing Land
This program was authorized by the 1996 Farm Bill for providing technical and educational assistance to
owners of private grazing lands. It offers opportunities for better land management, erosion reduction,
water conservation, wildlife habitat, and improving soil structure.
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Cooperative Extension
State land grant universities and Cooperative Extension play an important role in management
implementation. They have the expertise to research, transfer, and implement agriculture management
systems that will be needed to meet nutrient criteria. In addition, they have developed models and other
predictive management tools that will aid in selecting the most appropriate management activities.
Contact your local Cooperative Extension Agent or the Agriculture Department at a State land grant
university for more information on the services they can provide.
8.3 COMPREHENSIVE PROCEDURE FOR NUTRIENT MANAGEMENT
Numeric water quality criteria adopted by States into their water quality standards can also serve as
effective scientific tools for comprehensive water and land resource management. Effective programs
incorporate aspects of prevention and maintenance as well as restoration. It is important that existing
high-quality waters be managed wisely as a public resource, and waters whose uses are not yet threatened
or impaired, but nonetheless are at risk from ongoing pollution, should be identified and managed so that
designated uses are maintained in the future. The following 10-step management program that originated
with the Wisconsin Inland Lakes Program (Gibson et al. 1983) has since been refined to become a natural
resources management approach with utility for any water resource. States and Tribes can use this
approach in addition to established regulatory protocols. Though intended to protect or enhance coastal
marine or estuarine waters, the program does not, however, establish or replace any mandated procedures
as part of a regulatory requirement.
Management of tidal and marine waters may be approached as a rational progression of actions beginning
with a statement of major stressors and symptoms and progressing logically to a course of action and
final assessment to determine the relative success of the effort. The following steps illustrate this
management approach. States or communities are encouraged to adapt this technique to suit their
particular needs and expectations. Where considerable information is already available, some of these
steps may be skipped, but the methodology is presented here in detail for consideration.
Step 1: Status Identification
Data used during the preliminary nutrient criteria development process and the application of the criteria
will present the resource manager with the general status of the estuary or coastal area and the need for
responsive action. The information associated with these efforts, however, usually indicates a broad
status condition, for example, high nutrient concentrations, algal blooms, fish kills, loss of seagrasses, or
low-dissolved oxygen. Available data should be evaluated carefully to tease out potential connections to
land use practices or recent changes in practices (e.g., development, fishing pressure, stocking or lack
thereof). In particular, previous investigations should be reviewed to make a preliminary determination
of anthropogenic cause(s) as opposed to natural cycling of the system. Trend assessments on the
Chesapeake Bay, for example, are based on 10-year intervals in an effort to distinguish seasonality from
cultural impacts. Essentially, a preliminary evaluation of readily available data is necessary to ascertain
that there is indeed a problem or potential problem brought on by nutrient overenrichment and that the
sources of the threat probably can be addressed to the betterment of these waters and the public good.
Nutrient Criteria—Estuarine and Coastal Waters 8-19
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Step 2: Background Investigation
Given that the initial information reveals a viable management concern, it then becomes necessary and
justifiable to gather as much background information as possible about the waters in question. There are
three primary sources of such information.
Literature Searches
The initial effort here should be a search of the "gray literature" (often internal regional State and Federal
agency reports that provide specific information about the relevant waters). Sources of such information
include natural resource and fisheries agencies, forestry services, water quality administrations,
hydrological and geological survey offices, planning offices, multi-State or county commissions, and
community or environmental groups. A second source to comb through is peer-reviewed professional
literature journals and related publications such as proceedings of conferences and symposiums, which
may include specific studies of the estuary or coastal region of concern. The primary value of this
source, however, probably will be discussions of methods and techniques of investigation and
management. As the management investigation progresses, these sources of information become more
pertinent.
Questionnaires
In preparing a list of agencies from which reports may be solicited, the names of key personal contacts
should emerge. These contacts are the biologists, chemists, specialists, academics, resource managers,
and citizen activists most familiar with these waters. As the literature and baseline data are reviewed,
particular questions should develop, the answers to which will provide a fuller understanding of the
resource and lend direction to the investigation and eventual management plan. Particularly helpful will
be an understanding of the historical antecedents of the present status of the waters.
A standardized questionnaire can be prepared listing concerns such as the availability of any reports or
data or understanding of the history of development in the watershed, perhaps including industries,
agricultural practices, or development and structures associated with the resource. Particular episodes
may be noted for comment in the questionnaire, such as hurricanes, fish kills, algal blooms, or spill
events, as well as historical problems, such as marina development, port facilities, spoils disposal sites,
agricultural runoff, erosion problems, or development concentrations.
All discharge sites should be documented, such as industrial discharges, "Superfund sites," wastewater
treatment plants, drains, concentrations of housing with onsite wastewater treatment, marinas, major road
crossings, and tributaries potentially bearing large loadings of sediments or nutrients. Problem land use
areas along the shore also should be noted, for example, degraded wetlands, embayments where blooms
or fish kills regularly occur, or areas where seagrasses recently have contracted. In addition, it is helpful
to include a large, fairly detailed line drawing of the estuary or portion of the estuary or coastal reach and
its proximal watershed that the respondent may use to reference particular observations.
If at all possible, the questionnaire should be limited to no more than two pages of questions, including
space for answers plus the line drawing. Questions should be direct and concise. Determine exactly
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what you wish to learn and ask questions specifically related to this information. Opportunities for
additional comments should include an open-ended question at the end of the questionnaire.
To get the best response to a questionnaire, the potential respondents should be called first to confirm
their mailing addresses and availability. They should be advised of the nature of the study and their
cooperation then requested. Other potential respondents may be identified through these calls. If a large
survey is necessary, this preliminary step may not be possible. Most regional inquiries, however, are
made usually to no more than 50 specialists, and the additional information gained is well worth the
telephone calls.
Interviews
By this point in the background investigation process, the key people to contact for detailed information
should be evident. Their names will have come up in conversations and on reports, and they will be the
people providing the most helpful responses on the questionnaires.
Other valuable contacts are Basin Commissions; Interstate and State management agencies; EPA,
NOAA, and U.S. Fish and Wildlife Service specialists; the USDA Cooperative Extension Service agents
for the area; State, county, and municipal planners; and university faculty. Long-term residents and
commercial as well as sport fishermen and their organizations should be contacted also. Anecdotal
information can be invaluable and helps add perspective to other sources of data.
The interviews should assess the questionnaire data gathered already; they should clarify and elaborate
on the basic information generated. The interviews also are the means by which apparent contradictions
in perceptions or observations may be resolved at least partially. It should be noted that many people are
uncomfortable with recorded interviews; note-taking often is less intimidating. In either case,
immediately after each interview, a record of answers and observations should be prepared while the
impressions of the interviewer are still fresh.
The information compiled from each background investigation should clarify further the initial problem
stated. It should help resolve any ambiguities about the dynamics of the system and the human
community. In addition, the compilation should identify areas where more definitive, primary data
collection is required to clearly understand the nutrient problems of a particular waterbody and provide
direction for the subsequent management project.
Step 3: Data Gathering and Diagnostic Monitoring
Data obtained during the nutrient criteria development process are the mainstay of the database to be
prepared for any subsequent investigation. The intent of such a process is to develop a reasonable image
of the status of the estuary, bay, or coastal region. Diagnostic monitoring should expand on that structure
and extend the understanding from status of the resource to a diagnosis of causes of the overenrichment.
For example, where several tributaries of an estuary or salinity zone have been sampled and two are
identified as being of concern, they and other higher order feeder streams must now be sampled to further
target the locations of probable loadings. Though earlier sampling was done to portray the enrichment
Nutrient Criteria—Estuarine and Coastal Waters 8-21
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state of the waters, subsequent sampling should focus on near-shore areas of potential loadings,
tributaries, and portions of the tributaries where loadings may originate.
Diagnostic monitoring supports the identification of water quality problems and helps to develop an
appropriate management plan. General guidelines for conducting diagnostic monitoring are as follows.
Parameters To Sample
Diagnostic monitoring is conducted after nutrient criteria have been established. It might not be
necessary, therefore, to sample some parameters that are not related to the criteria.
• Diagnostic sampling for nutrients requires an estimation of nutrient loading and sources. Major
potential sources of nutrients (e.g., tributary streams, groundwater flow, runoff, illegal discharges,
atmospheric deposition) should be identified and sampled in such a way as to obtain an estimate of
annual loads from each source.
• The variables and techniques employed in the preliminary survey should be reviewed for adequacy
and either repeated or augmented. A manager should not eliminate the basis of the original
classification by dropping any variables or stations at this point. Documenting potential success or
failure of the subsequent management program will require "before" and "after" databases, and the
initial survey design should be modified only after careful consideration and due attention to
reestablishing the baseline survey.
Flow measurements also are an essential part of this survey. Perhaps more important than in any other
water resource investigation, attention to tidal state and amplitude and seasonal hydrologic characteristics
is imperative to determining the extent and source of nutrient loadings to estuaries and coastal marine
waters. If nutrient concentrations are to be compared meaningfully and loading estimates made,
cross-sectional areas and flow rates for all tributary streams and discharges also must be included in the
survey design. These measurements must be made or extrapolated whenever water quality samples are
collected. Without this information, assigning priorities to various loading sources identified in the
investigation will be difficult or impossible.
Sampling Frequency
Sampling frequency will increase for diagnostic monitoring because the sample population is now a
particular area of the waterbody. Sampling should occur repeatedly during the growing season to
precisely characterize individual areas as well as discharges and loadings. Statistical power analysis can
be used to determine the appropriate sample size based on the purpose of the sampling and the acceptable
error (see Chapter 5).
In addition to expanding the number of stations and parameters to accommodate diagnostic
determinations, the survey design should address temporal variables by sampling these stations during
each season of the year at times calibrated to the particular climate and locale. Accommodation may be
needed for periods of base flow, maximum runoff, turnovers, periods of maximum and minimum
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productivity, and, in some instances, migratory patterns offish or waterfowl. Seasonal changes in land
use such as peak summer or winter vacation periods, agricultural applications and harvests in the
watershed, and seasonal commercial or industrial activities also should be addressed.
To separate signals from seasonal "noise," it may be necessary to gather survey data for 2 or more
consecutive years to strengthen data assessment (as noted above, some estuary management programs
survey over as many as 10 years, requiring an extensive operating budget). Such assessments will require
a robust statistical evaluation of the data; this element should be incorporated into the study design at the
outset. As with the initial survey design, the preliminary statistical tools chosen may be carried into this
subsequent design as well. Care should be taken to replicate sample collections to ensure representative
sample design and confidence in the results obtained. Early inclusion of a skilled environmental
statistician on the management team is advisable.
Sampling Location
If turbidity, nutrients, and algae are known to be variable across the surface of an estuary or salinity zone,
then multiple sample sites within that zone are required. The exact number of sampling sites in a zone is
determined by the spatial variability of nutrients, turbidity, and chlorophyll and the desired precision. In
general, within a basin or zone, variation in time is larger than variation in space (Knowlton and Jones
1989). Thus, chlorophyll samples taken 2 weeks apart may differ severalfold, but samples taken on the
same day 500 meters apart are likely to differ much less. Depending on the questions being addressed in
the investigation, spatially composite samples may be more cost-effective than separate samples from
several sites.
The design and placement of these sample stations will rely heavily on the proximal and watershed land
use information garnered from the background investigation. The overall objective should be to bracket
suspected sources of nutrient loadings in the tributaries and near-bank areas so parcels can be either
selected or eliminated as potential candidates for management attention.
Step 4: Source Identification
The cumulative information gathered should now provide a clear image of the state of the estuarine or
coastal segment, the most likely sources of nutrient loadings or related degradation, and their relative
contributions to the problem. It is important to note that this process reveals only local sources of the
overenrichment. Atmospheric deposition of nitrogen compounds and other broad-scale impacts beyond
the watershed scale are not specifically addressed and must be assumed as essentially an environmental
constant. With all the risks this constant entails, it is probably not an undue assumption; remediation of
such depositions is probably beyond the scope of most nutrient management projects employing this
guidance.
The problems to be identified are likely to be as diverse as the geology, hydrology, and land use practices
of the waterbody and watershed. Typical developments include sediment resuspension and nutrient
re-release; biotic imbalances affecting nutrient utilization caused by overfishing or stock
mismanagement; discharge of excess nutrients directly to the waters by wastewater treatment plants,
Nutrient Criteria—Estuarine and Coastal Waters 8-23
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storm water runoff, or failing septic systems; and runoff from municipalities, subdivisions, farms,
commercial enterprises, and industrial activities. Other problems have included concentrations of
migratory and resident waterfowl contributing to an excess of nutrients, removal or filling of bank areas
and wetlands that once intercepted nutrient runoff, runoff of herbicide applications that killed
macrophytes and promoted nuisance algal blooms, and chronic, low-dissolved oxygen problems attendant
to overenrichment and vegetative imbalances.
Any combination of these in situ and land use problems can cause a cumulative overenrichment problem.
Management planning requires identifying first the loading sources and, second, of those sources, the
ones that are most significant. Proximity of a source to a lake or reservoir (or in some cases, the
ubiquitous nature of a source throughout a watershed such as subdivision or farm runoff), the relative
loading estimate of that source, and the likelihood of successful remediation are the key factors in
deciding which problem sources are priorities for inclusion in a management plan.
Loading estimation models are valuable for estimating the relative significance of various nutrient
sources in the watershed with respect to the likely response of the waters. Chapter 9 describes many of
these models and their relative utility. Modeling permits a manager to try out various scenarios and
combinations of techniques to estimate their likely effectiveness. Some of these options for a nutrient
management plan are discussed in the next step.
Step 5: Management Practices for Nutrient Control
Once the major sources of concern are identified and agreed on by the management planners, remedial
measures appropriate to these sources must be identified. Management practices are well defined and
documented for a variety of land uses in EPA guidance documents, USDA manuals, U.S. Forest Service
manuals, and urban land use planning guides. Resource managers should study these references for
likely approaches to consider and then consult regional experts in each of the subject land uses for
qualification and other suggested management practice recommendations. Bringing these specialists
together as a small workgroup is an effective, although sometimes contentious, way to develop the most
technically sound approaches to such problems.
Fitting the various components together in a comprehensive management plan is challenging. It calls for
both imagination and cooperation. Usually no one approach stands out as the obvious best choice.
Instead, two or three permutations of several generally agreed on BMPs will evolve from the planning
sessions.
Selection of the optimal approach—or more likely, the best candidates—should first involve careful
assessment by the planning workgroup and then consultation with all elements of the watershed
community, both organized interest groups and private landowners. The first phase should be conducted
using the threefold framework of evaluation developed by the Department of Resource Development at
Michigan State University (Figure 8-2). The premise behind this approach holds that the most effective
and achievable management plan should address three elements of practicality: scientific validity,
sociopolitical consideration, and economic consideration.
8-24 Nutrient Criteria—Estuarine and Coastal Waters
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• No resource or environmental management plan should be considered unless it is scientifically valid.
The technology proposed should be based in sound science and tested and validated. No attempt to
manipulate the environment and peoples' land use prerogatives should be made unless it can be
demonstrated in advance that the technique is reliable or at least that the risks are quantifiable and
understandable.
• The proposed approach should be cost-effective and affordable by the community. Among
technically sound plans to achieve desired goals, the most cost-effective (typically those with
elements that have the greatest benefit-cost ratios) are the easiest to implement and most likely to
satisfy the public interest.
• The management plan should have adequate social and political acceptability. A plan that seems
rational and cost-effective may conflict with the collective values of the local public. Any action
taken in addition to existing requirements should always be researched carefully for justification,
efficacy, lead time required, and likely effects on various segments of the community.
The resource manager most likely to achieve success will consider and responsibly address each of these
three elements. All candidate alternatives should be evaluated in this manner and revised as necessary.
Such evaluation not only generates the optimal plan (or plans where competing but different strengths are
evident), but documents the rationale, essential for public review before the final selection is made.
Socio-political
consideration
Economic
consideration
Optimal management plan if
all factors are considered.
Figure 8-2. "Threefold framework" of evaluation.
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Involving the public in the process throughout is highly beneficial, and invitations to meetings or
advisories to all potentially interested parties should be provided regularly, if not from the outset then
certainly before plan selection and approval are needed. A balance must be struck between making
public announcements too early, which could arouse people before sufficient information has been
generated, and making announcements too late, which may lead to suspicions that information is being
withheld from the public.
Step 6: Detailed Management Plan Development
A detailed management plan should include all 10 steps of the process described here. The first five
steps are necessary to achieve the design of the plan, but they also should be included so that anyone
reading it will understand what has gone into the effort.
Natural resource management efforts can include one or more of these three elements: education,
financing, and regulation; it is best to initiate them in the order presented here. Start with relatively
low-cost information and education efforts to acquaint people with the problem and how you propose to
address it, and to obtain their suggestions and perceptions. A good educational effort should be the
incentive for volunteer agreements and cooperative action. A grant-in-aid or other assistance often is the
key element to encourage individuals or jurisdictions to adopt appropriate water resource protection
practices. Regulatory actions are necessary and appropriate when mandated by law, when cooperation
and compliance are unlikely to occur otherwise, and when voluntary efforts have not succeeded.
Step 7: Implementation and Communication
Periodic progress reports during implementation of the management project are opportunities to
communicate with administrators, other involved agencies, politicians interested in the project, the
general public and landowners, and other interest groups. Progress reports should be brief and candid.
They will be part of the public record so that all parties are properly informed, the post-project cry of
inadequate notification is avoided or at least minimized, and techniques and methods used are
documented.
Regional public meetings and hearings are excellent ways to communicate. The more controversial an
issue, the more communication is necessary. To continue implementing a plan despite significant
opposition without evaluating the consequences is a mistake, especially if a change to a step in the
management plan with additional public consultation would still achieve the same objective.
Step 8: Evaluation Monitoring and Periodic Review
The management plan should always include "before, during, and after" water resource quality
monitoring to demonstrate the responses of the system to management efforts. This is the reason for
maintaining and expanding the initial survey stations. Monitoring data are important for evaluating
progress and are included in the requisite progress reports described above. The change or lack thereof
in the status of the estuarine or coastal waters is the ultimate determination of management success.
8-26 Nutrient Criteria—Estuarine and Coastal Waters
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These built-in monitoring schedules should include seasonality and periodic data assessment intervals for
management review to permit responses to changing circumstances, modifications of methods, schedules,
and changes of emphasis as needed.
Step 9: Completion and Evaluation
Management projects are frequently planned, initiated, and concluded and even new initiatives
undertaken to meet pressing schedules without sufficient evaluation of the initial project.
Reviews of progress reports, of the original objectives, and of the monitoring data will reveal whether the
lake or reservoir trophic state was successfully protected or improved. Evaluation also provides
documentation for determining whether the project's methods and techniques can be applied elsewhere,
perhaps with modification. Finally, it will reveal any mistakes that should be avoided in future projects
and perhaps will demonstrate that a sequel project is required to fully accomplish the original objectives.
Step 10: Continued Monitoring of the System
The monitoring initiated and expanded in the course of the project can now be reduced to periodic
measuring of key variables at critical times and locations. The purpose now is to keep sufficiently
informed of the status of the waters to ensure that the protection or remediation achieved is maintained.
If periodic evaluation monitoring indicates a return of trophic decline, intervention should be possible at
an early point so that costs of preserving that which was achieved are reduced. These evaluation and
periodic monitoring steps essentially complete the process. If new issues arise, the manager returns to
Step 1 with a new problem statement. General guidelines associated with evaluation monitoring are
provided below.
Parameters To Sample
Each water quality parameter discussed in the Indicators chapter—TP, TN, chlorophyll a, Secchi depth,
and dissolved oxygen as well as perhaps selenium and vegetation and indicator organisms—should be
sampled during maintenance monitoring. Because the purpose of maintenance monitoring is to
determine if conditions have changed or if criteria are exceeded, other physical or chemical variables
need not be measured.
Sampling Frequency
Sampling efforts for maintenance monitoring can be adaptive and sequential, so that a certain minimum
of information is collected at regular intervals, and if data indicate change or uncertainty, the sampling
effort (in both time and space) can be increased to attempt to reduce the uncertainty. For example, a
station in an undisturbed area could be sampled once every 5 years, from a single visit during an index
period (say, midsummer). If results suggest a change in conditions beyond what is normally expected for
these waters, then additional and more frequent sampling can be continued to determine if the departure
from "normal" conditions is real and if it is ecologically significant. If TP, TN, chlorophyll a, and Secchi
depth relationships have been established, it may be cost-effective to use Secchi depth as a preliminary
indicator; if a trigger value is detected, more parameters can be measured.
Nutrient Criteria—Estuarine and Coastal Waters 8-27
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Such variation also suggests different levels of maintenance monitoring, depending on existing
knowledge of the waters and expectations. Maintenance monitoring may be done for several purposes:
• Routine monitoring of waters of known quality (i.e., sampled before) that are not expected to change
greatly
• Initial sampling of a station of unknown quality
• Monitoring of a station or stations of known quality expected to change, say, as the result of
watershed development or restoration efforts
Routine monitoring of stations of known quality is the least intensive and typically requires sampling
once every several years, as in the example above. However, initial sampling of an estuary or coastal
area of unknown quality requires the same sampling effort, and parameters, as the classification survey.
Monitoring a known estuary or portion of an estuary or coastal area that is expected to change or
suspected to have changed requires more intensive effort, typically an increase in sampling frequency to
several times during the growing season, to obtain seasonal averages of indicator values.
The actual frequency of sampling should be determined by the number of samples required to detect an
ecologically relevant change in the indicators, resources available for the monitoring program, and
amount of time for a change to be detected. These considerations require power analysis using existing
or preliminary data, and tradeoffs of desired significance level, desired power, desired effect size that is
detectable, ecological significance, and most important, resources (labor and money) available for the
monitoring program.
Sampling Location
For routine monitoring, it is recommended that the sampling locations be at least the same as for the
classification survey to provide the database with a certain minimum continuity.
8.4 RESOURCES
Listed below are selected publications concerning coastal or estuarine and watershed management and
protection.
• National Research Council. 2000. Clean Coastal Waters - Understanding and Reducing the Effects
of Nutrient Pollution. National Academy Press. Washington, DC.
• Gibson, G. R., M.L. Bowman, J. Gerritsen, and B.D. Snyder. 2000. Estuarine and Coastal Marine
Water Bioassessment and Biocriteria Technical Guidance. EPA 822-B-00-024. U.S. Environmental
Protection Agency; Office of Water; Washington, DC.
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Sharpley, AN, ed. 2000. Agriculture and phosphorus management: the Chesapeake Bay. Boca
Raton, FL: Lewis Publishers.
This text is a compilation of conference proceedings describing nutrient dynamics in the watershed
of Chesapeake Bay, with emphasis on agricultural loadings and practices. Although directed at an
estuarine environment, much of the agriculturally based nutrient information has broad application.
U.S. Environmental Protection Agency. 1993. Guidance Specifying Management Measures for
Sources of Nonpoint Pollution in Coastal Waters. EPA-840-B-92-002.
The EPA Office of Water produced the 1993 guidance document to support the Coastal Zone Act
Reauthorization Amendments of 1990. This document describes several management measures to
control nonpoint sources of pollution, including nutrients.
U.S. Environmental Protection Agency. 1995. Watershed Protection: A Project Focus. EPA
841-R-95-003.
This document focuses on developing watershed-specific programs or projects. It provides a
blueprint for designing and implementing watershed projects, including references and case studies
for specific elements of the process. The document illustrates how the broader principles of
watershed management, including all relevant Federal, State, Tribal, local, and private activities, can
be brought to bear on water quality and ecological concerns.
U.S. Environmental Protection Agency. 1995. Watershed Protection: A Statewide Approach. EPA
841-R-95-004.
This document is primarily designed for State water quality managers. A common framework for a
statewide watershed approach focuses on organizing and managing a State's major watersheds (called
basins in this document). In this statewide approach, activities such as water quality monitoring,
planning, and permitting are coordinated for multiple agencies on a set schedule within large
watersheds or basins.
U.S. Environmental Protection Agency. 1997. Monitoring Consortiums: A Cost-Effective Means to
Enhancing Watershed Data Collection and Analysis. EPA 841-R-97-006.
This document addresses coordination in watershed monitoring. As demonstrated in the document's
four case studies, consortiums can stretch the monitoring dollar, improve cooperation among
partners, and increase sharing of expertise as well as expenses of data collection and management.
U.S. Environmental Protection Agency. 1997. Land Cover Digital Data Directory for the United
States. EPA 841-B-97-005.
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Land cover, which is the pattern of ecological resources and human activities dominating different
areas of the Earth's surface, is one of the most important data sources used in watershed analysis and
the management of water resources throughout the country. The 75 land cover data summaries in
this directory include contact information to assist readers who may want to acquire copies of the
digital data for their own use.
U.S. Environmental Protection Agency. 1997. Designing an Information Management System for
Watersheds. EPA 841-R-97-005.
This document is an introduction to the information management responsibilities and challenges
facing any watershed group. The document reviews the fundamentals of identifying information
management needs, integrating different databases, evaluating hardware and software options, and
developing implementation plans.
U.S. Environmental Protection Agency. 1997. Information Management for the Watershed
Approach in the Pacific Northwest. EPA 841-R-97-004.
This document centers on a series of interviews with leaders and key participants in the statewide
watershed approach activities in the State of Washington. The document reviews Washington's
statewide watershed activities in case study fashion.
U.S. Environmental Protection Agency. 1998. Inventory of Watershed Training Courses. EPA
841-D-98-001.
This inventory provides one-page summaries of 180 watershed-related training courses offered by
Federal and State agencies; it also lists resource professionals in the private sector.
U.S. Environmental Protection Agency. 1997. Statewide Watershed Management Facilitation. EPA
841-R-97-011.
This document addresses statewide watershed management and the process of facilitating the
development or reorientation of statewide watershed programs. It includes State case histories.
U.S. Environmental Protection Agency. 1996. Watershed Approach Framework. EPA
840-S-96-001.
This publication revisits and updates EPA's vision for a watershed approach, first explained in a 1991
document entitled "Watershed Protection Approach Framework." It describes watershed approaches
as coordinating frameworks for environmental management that focus public and private efforts to
address the highest priority problems in defined geographic areas, involving both ground and surface
water flow.
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U.S. Environmental Protection Agency. 1997. Top 10 Watershed Lessons Learned. EPA
840-F-97-001.
Watershed work has been going on for many years now, and this 60-page document summarizes the
top lessons that have been learned by watershed practitioners across the United States regarding what
works and what does not.
U.S. Environmental Protection Agency. 1999. Catalog of Federal Funding Sources for Watershed
Protection (second ed.). EPA 841-B-99-003.
Many sources of Federal funding are available to support different aspects of watershed protection
and specific types of local-level watershed projects. This document presents information on 52
Federal funding sources (grants and loans) that may be used to fund a variety of watershed projects.
U.S. Environmental Protection Agency. 1997. Watershed Training Opportunities. EPA
841-B-97-008.
This is a 22-page booklet developed to highlight watershed training opportunities offered by EPA's
Office of Water and the Watershed Academy. It covers training courses and educational materials on
watersheds produced throughout the EPA Office of Water.
U.S. Environmental Protection Agency. 1997. Stream Corridor Restoration: Principles, Processes
and Practices. EPA 841-R-97-011.
This document is a practical reference manual and logical framework to help environmental
managers recognize stream restoration needs and design and implement restoration projects.
U.S. Environmental Protection Agency. 1997. Protocol for Developing Nutrient TMDLs. EPA
841-B-99-007.
This protocol is an organizational framework for the TMDL development process for nutrients. It
leads to an understandable and justifiable TMDL.
Nutrient Criteria—Estuarine and Coastal Waters 8-31
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Model Identification and Selection
Model Classification
Use of Models for Nutrient Investigation
Management Applications
CHAPTER 9
Use of Models in Nutrient Criteria
Development
All models are wrong. Some models are useful.
(George E. P. Box)
9.1 INTRODUCTION
This chapter addresses the role of models in nutrient criteria development. It is closely linked to Chapter
5, which addresses database development, sampling designs, and monitoring. One system is said to
model another when the observable variables in the first system vary in the same fashion as the
observable variables in the second (NRC 2000). Chapter 7 of the NRC report goes on to state that
models come in many forms. They may be empirically derived statistical relationships plotted on a
graph, physical analogues (e.g., mesocosms) of the system of interest, analogues of different systems that
have useful parallel relationships of observable factors (e.g., from physics, the flow of water through
pipes to model the flow of electrons through an electrical circuit), or numerical models run on computers
that are based on first principles or empirical relationships.
Environmental water quality models have several uses (e.g., reduce ecosystem complexity to a
manageable level, improve the scientific basis for development of theory, provide a framework to make
and test predictions, increase understanding of cause-and-effect relationships, and improve assessment of
factor interaction). Reliable predictions stand out as a salient requirement because of the social and
economic consequences if predictions are unreliable. Many times decisionmakers rely on models to
guide their environmental management choices, especially when costly decisions are involved and the
problem and solution involve complex relationships. This is exemplified by the decision of the Long
Island Sound Hypoxia Management Conference (see Case Study for Long Island Sound). Generally,
empirical and mathematical models are the most widely used models that statistically or mathematically
relate nutrient loads or concentrations to important ecological response variables (e.g., dissolved oxygen
deficiency, algal blooms and related decrease in water clarity, and loss of seagrasses). They both depend
on the scientific robustness and accuracy of underlying conceptual models.
This chapter addresses both empirical and mathematical models. Considerably more space is devoted to
mathematical models, because they are capable of addressing many more details of underlying processes
when properly calibrated and validated. They also tend to be more useful forecasting (extrapolation)
tools than simpler models, because they tend to include a greater representation of the physics, chemistry,
and biology of the physical system being modeled (NRC 2000). A great danger in complex mathematical
models is that error propagation is difficult to explicitly measure, and there is a tendency to use a more
complex model than required, which drives costs up substantially and unnecessarily. Another
consideration that is gaining acceptance is that mathematical models need to be appropriately scaled to
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9-1
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spatial and temporal processes, or they may suffer problems similar to empirical models when one
extrapolates the results of scaled experiments to full-sized natural systems. Also, empirical coefficients
introduced into equations often hide the degree of uncertainty concerning the fundamental nature of the
processes being represented.
Use of Empirical Models in Nutrient Criteria Development
Statistical models are empirical and are derived from observations. To be useful as predictive tools,
relationships must have a basis in our understanding, typically represented by conceptual models.
However, extrapolation from empirical data is known to be uncertain. Thus, these models are most
reliable when used within the range of observations used to construct the model. When shown to meet
program objectives and requirements, empirical models are a desirable place to begin model development
and, if later determined to be required, they often provide insights into the structure needed for
development of mathematical models. Empirical models typically are useful if only a subsystem of the
larger ecosystem is of primary interest.
Frequently, the impression is given that the only credible water quality modeling approach is that of
mathematical process-based dynamic computer modeling. This is not the case. For example, a Tampa
Bay water quality modeling workshop in 1992 (Martin et al. 1996) produced the consensus
recommendation that a multipronged (mechanistic and empirical) modeling approach be implemented to
provide technical support for the water quality management process. The Tampa Bay National Estuary
Program produced an empirical regression-based water quality model. The estimated N loads were
related to observed chlorophyll concentrations using the regression model (Janicki and Wade 1996):
where C t s = average chlorophyll a concentration at month t and segment s,
L t s = total N load at month t and segment s,
a u and P s are regression parameters.
A related model equated Secchi depth to average chlorophyll a concentrations (Greening et al. 1997).
This analysis was followed by an empirical model that related N loadings to in-bay chlorophyll a
concentrations.
There are many other examples of empirical models used to relate environmental forcing functions to
ecological responses, especially nutrient load/concentration and response relationships. Much of the
professional aquatic ecological literature reports on use of empirical models (e.g., Chapters 2 and 3).
Empirical models have their limitations, but when judiciously applied, they offer a highly useful tool to
water quality managers.
9-2 Nutrient Criteria—Estuarine and Coastal Waters
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Use of Mathematical Models in Nutrient Criteria Development
Mathematical models can play an important role in assessing acceptable nutrient loads and
concentrations in estuaries and near-coastal areas. For example, models are used to:
• Develop a relationship between external nutrient loads and resulting nutrient concentrations, which
can then be used to define allowable loads
• Define the relationship between nutrient concentrations and other endpoints of concern, such as
biomass or dissolved oxygen
• Provide an increased understanding of the factors affecting nutrient concentrations, such as the
relative importance of point and nonpoint source loads
• Simulate relationships between light attenuation and expected depth of sea grass growth
The intent of this section is to describe the models available for assessing the relationship between
nutrient loading and nutrient-related water quality criteria for estuaries and near-coastal waters. This
chapter provides general guidance and some specific procedures for selecting and applying an
appropriate model. It is divided into the following sections: (1) Model Identification and Selection, (2)
Model Classification, (3) Use of Models for Nutrient Investigation, and (4) Management Applications.
Extensive EPA guidance (i.e., U.S. EPA 1985, 1990a-c; 1997; EPA document # 841-B-97-006) currently
exists on these topics. This section serves primarily to condense the existing guidance with some
modifications, to reflect changes in the science that have occurred subsequent to their publication. In
addition, emphasis is placed on the simpler, more empirical techniques that are applied most easily.
Readers are referred to the original guidance materials for more detailed discussions of the concepts
described in this section.
9.2 MODEL IDENTIFICATION AND SELECTION
The first steps in the modeling process are model identification and selection. The goals are to identify
the simplest model(s) that addresses all of the important phenomena affecting the water quality problems,
and to select from those the most useful analytical formula or computer model. Selection of too simple a
model can result in predictions of future water quality that are too uncertain to achieve the decisions or
objectives of the study. On the other hand, selection of an overly complex model may also result in
misdirected study resources, delays in the study, and increased cost. Predictive uncertainty may increase
to unacceptable levels because of model parameters that cannot be adequately estimated with available
data. Study costs will increase because of the additional data requirements and the expanded computer
and staff time needed for model runs, analysis, and sensitivity studies.
Nutrient Criteria—Estuarine and Coastal Waters 9-3
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Model Identification
Model identification entails four basic steps:
• Establish study objectives and constraints
Determine water quality pollutant interactions
Determine spatial extent and resolution
• Determine temporal extent and resolution
Each is discussed below.
Study Objectives and Constraints
The first step in identifying an appropriate model for a particular site is to clearly delineate the objectives
of the modeling analysis. These objectives address questions such as:
• What are the nutrients of concern?
• What are the environmental endpoints of concern?
What spatial and temporal scales are adequate for management concern?
• What management issues must the model address?
• What is the acceptable level of uncertainty in model predictions?
The nutrients of concern addressed in this document are nitrogen or phosphorus (depending on which is
the limiting nutrient or will become limiting after controls are implemented). Environmental endpoints
of concern are total nutrient concentration and other indicators of excessive nutrients such as
chlorophyll/biomass and minimum dissolved oxygen. Local, State, and Federal regulations contribute to
the definition of objectives by specifying time and space scales that the model must address: for
example, the averaging period, or the season at which the criteria are applicable.
All expected uses of the model are to be stated clearly in advance. If the model will be used to predict
future allowable nutrient loads, the specific conditions to be evaluated must be known. Then a final
study objective is established that pertains to the required degree of reliability of model predictions,
which may vary depending on whether the model application is designed for screening level estimation
or for more detailed predictions.
The reliability objective is directly related to project constraints, as there is often a mismatch between
desired model reliability and available resources. Resource constraints can cover four areas: data, time,
level of effort, and expertise. Appropriate model selection must be balanced between competing
demands. Management objectives typically favor a high degree of model reliability, but resource
constraints generally prohibit the degree of reliability desired. Decisions often are required regarding
whether to proceed with a higher-than-desired level of uncertainty, or to postpone modeling until
additional resources can be obtained.
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Water Quality/Pollutant Interactions
After the pollutants and water quality indicators are identified, the significant water quality processes
must be determined. These processes directly or indirectly link the pollutants to be controlled with the
primary water quality indicators. All other interacting water quality constituents thought to be significant
should be included at this point. This consolidation can best be seen in a diagram or flow chart
representing the mass transport and transformations of water quality constituents in a defined segment of
water. Figure 9-1 illustrates variables and processes important to the eutrophication process. Not all of
these need to be included in the actual model selected for use. Those excluded, however, should be
considered externally and reflected in the coefficients.
At the end of this step all the available knowledge of a system should be assimilated in a way that permits
major water quality processes and ecological relationships to be evaluated for inclusion in the numerical
model description. This conceptual model is the starting point from which systematic reductions in
complexity can be identified to provide an adequate representation of the system while meeting the
objectives of the study.
The simplest level of model complexity considers only total nutrient concentrations and assumes that all
of the processes shown in Figure 9-1 either have no effect on total nutrient concentrations (as is
sometimes assumed for total nitrogen), or can be lumped into a single overall loss coefficient.
Models that simulate phytoplankton concentrations or dissolved oxygen typically include all of the
processes shown in Figure 9-1, and sometimes many more, to describe such processes as sediment
diagenesis and competition among multiple phytoplankton classes. Denitrification in the model is
expressed in terms of the water column carbonaceous biochemical oxygen demand (CBOD).
Spatial Extent and Scale
Two spatial considerations must be addressed in the model identification process: spatial extent and
scale. Spatial extent pertains to the specific boundaries of the area to be assessed. Spatial scale pertains
to the number of dimensions to be considered and the degree of resolution to be provided in each
dimension.
Several guidelines can help locate proper model boundaries. In general, the boundaries should be located
beyond the influence of the discharge(s) being evaluated. Otherwise, proper specification of boundary
concentrations for model projections is very difficult. Boundaries should be located where flow or stage
and water quality are well monitored. Upstream boundaries should be located at a fall line, or at a gaging
station in free-flowing, riverine reaches. Downstream boundaries are best located at the mouth of an
estuary, or even nearby in the ocean. For large estuaries with relatively unaffected seaward reaches, the
downstream boundary can be located within the estuary near a tidal gage and water quality monitoring
station.
Nutrient Criteria—Estuarine and Coastal Waters 9-5
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SOD
Settling
Sediment
Figure 9-1. Eutrophication model framework : an example of hypoxia-based conceptual framework
for water quality model. Source: Bierman et al. 1994.
Appropriate model spatial scale requires consideration of two factors: (1) the extent to which spatial
gradients in water quality occur and (2) the extent to which these variations need to be considered from a
management perspective. Real estuaries and near-shore waters all exhibit three-dimensional properties.
There are gradients in hydrodynamic and water quality constituents over length, width, and depth. The
effective dimensionality of an estuary includes only those gradients that affect the water quality analysis
significantly.
One-dimensional models consider the change in pollutant concentration over a single dimension in space,
typically oriented longitudinally down the length of an estuary. Two-dimensional models can consider
concentration gradients in the lateral and longitudinal directions (termed x-y orientation), or
concentration gradients that occur longitudinally and vertically (termed x-z orientation). Three-
dimensional models describe changes in concentration that occur over all three spatial dimensions.
These models provide the most detailed assessment of pollutant distribution with respect to a discharge;
they also have the most extensive model input requirements and are the most difficult to apply.
Justifiable reductions in dimensionality result in savings in model development, simulation, and analysis
costs. Usually the vertical and/or lateral dimension is neglected. Eliminating a dimension from the water
quality analysis implies acceptable uniformity of water quality constituents in that spatial dimension. For
example, use of one-dimensional models implies acceptably small deviations in concentration from the
9-6
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cross-sectional mean, both vertically and laterally. This judgment requires understanding both the
transport behavior of estuaries and the specific objectives of the study.
Spatial variations are best determined by plotting observed water quality concentrations versus distance
along the dimensions of concern. If such data are not available, other types of methods are available to
estimate the importance of spatial variations. These are described in U.S. EPA (1990a) and discussed
briefly here. The methods can be divided into three categories:
• Relatively simple desktop methods that compare the stratification potential caused by freshwater
inputs to the mixing potential caused by tides and other currents
• Dye studies that observe the degree of mixing
Geomorphological classification, which categorizes estuaries and the degree of mixing based on
standard morphological categories (e.g., drowned river valleys)
Two situations exist where the observed spatial variations can be ignored. The first is when the primary
location of water quality concern occurs in an area where these gradients are not important. A good
example would be a nutrient modeling study to consider the impacts of a discharge on phytoplankton. If
it takes 2 miles for a bankside discharge to undergo complete lateral mixing, but the location of
maximum algal density is 5 miles downstream, lateral variability in water quality need not be described
by the model. The second situation where a known gradient need not be modeled is where management
objectives are not concerned with the gradient. Examples of this include water quality standards that are
expressed on a spatially averaged basis.
The choice of spatial scale and layout of the model network requires considerable judgment. Knowledge
of the regulatory problem must be combined with knowledge of the loading, transport, and
transformation processes and an understanding of the model chosen to perform the simulations.
Competing factors often must be balanced, such as precision and cost, or the better fit of one section of
the network versus another.
Temporal Extent and Scale
The temporal resolution of water quality models falls into one of two broad categories, steady state or
dynamic (i.e., time-variable). Steady-state models predict pollutant concentrations that are expected to
result from a single set of loading and environmental conditions. Dynamic models predict changes in
water quality over time in response to time-variable loads and environmental conditions.
Steady-state models are much easier to apply and require considerably fewer resources than dynamic
models. This ease of application makes them the preferred modeling framework when loading to the
system can be assumed to be constant and information on changes in concentration over time is not
required. Potential uses of steady-state models include calculation of seasonal average total nutrient
concentrations in response to seasonal average loads. Steady-state models also have been used to predict
Nutrient Criteria—Estuarine and Coastal Waters 9-7
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"critical condition" low-flow phytoplankton densities. Steady-state models are poorly suited for
evaluating individual intermittent events (i.e., runoff) but can be used to evaluate the cumulative effect of
multiple, intermittent events on a seasonal basis.
The timeframe to be represented for a particular steady-state simulation (e.g., monthly average, seasonal
average) must be longer than the water residence time (flushing time) of the estuary. The water residence
time is the time required to remove a parcel of water from an upstream location in an estuary. Factors
that control flushing include tidal action, freshwater inflow, and wind stress. Typical flushing times
range from days in small estuaries, or those dominated by tributary flow, to months in large estuaries
during low tributary flow conditions. Several formulae have been used to estimate flushing times. The
Fraction of Fresh Water Method, the Tidal Prism Method, and the Modified Tidal Prism Method are
fully discussed in Mills et al. (1985) and briefly described in the following section.
Dynamic models should be used when information on changes in concentration overtime is required.
Dynamic models can be divided into two categories, quasi-dynamic and fully dynamic. Quasi-dynamic
simulations predict variations on the order of days to months. The effects of tidal transport are time-
averaged, and net or residual flows are used. Fully dynamic simulations predict hour-to-hour variations
caused by tidal transport.
The duration of dynamic simulations can range from days to years, depending on the size and transport
characteristics of the study area, the reaction kinetics and forcing functions of the water quality
constituents, and the strategy for relating simulation results to the regulatory requirements. One basic
guideline applies in all cases: the simulations should be long enough to eliminate the effect of initial
conditions on important water quality constituents at critical locations. Flushing times provide the
minimum duration for simulations of dissolved, nonreactive pollutants. The annual sunlight and
temperature cycles almost always require that eutrophication simulations range from seasons to years.
Predicting the year-to-year eutrophication response of large estuaries is best accomplished by quasi-
dynamic simulations. In general, if the regulatory need or kinetic response is on the order of hours, then
fully dynamic simulations are required; if regulatory needs are long-term averages and the kinetic
response is on the order of seasons to years, then quasi-dynamic or steady-state simulations are preferred.
Model Selection
The goal of model selection is to obtain a simulation model that effectively meets all study objectives. In
the final analysis, how a model is used is more important to its success than exactly which model is used.
Nevertheless, although selection of an appropriate model will not guarantee success, it will help.
Selection of an inappropriate model will not guarantee failure, but will render a successful outcome more
difficult.
Models may be classified in different and somewhat arbitrary ways. Some models may not quite fit in
any category, or may fit well in several. In addition, models tend to evolve with use. The exact
capabilities of the individual models described here may change. In particular, pollutant fate processes
9-8 Nutrient Criteria—Estuarine and Coastal Waters
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may be modified. Usually the computational framework and the basic transport scheme remain stable
over time. For this reason, transport characteristics will provide the basis for the model classification
scheme used here. Models selected for discussion here are general purpose, in the public domain, and
available from or supported by public agencies.
9.3 MODEL CLASSIFICATION
Estuarine and near-coastal models consist of two components: hydrodynamics and water quality.
Although the hydrodynamic component is independent of the water quality component, water quality
depends on the transport processes controlled by hydrodynamics. As a result, estuarine models can be
classified as Level I to Level IV, according to the temporal and spatial complexity of the hydrodynamic
component of the model.
Level I includes desktop screening methodologies that calculate seasonal or annual mean total nutrient
concentrations based on steady-state conditions and simplified flushing time estimates. Steady-state
models use an unvarying flow condition that neglects the temporal variability of tidal heights and
currents. These models are designed for relatively simple screening level analyses. They also can be
used to highlight major water quality issues and important data gaps in the early model-identification
stage of a more complex study.
Level II includes computerized steady-state or tidally averaged quasi-dynamic simulation models, which
generally use a box or compartment-type network. Tidally averaged models simulate the net flow over a
tidal cycle. These models cannot predict the variability and range of nutrient concentrations throughout
each tidal cycle, but they are capable of simulating variations in tidally averaged concentrations over
time. Level II models can predict slowly changing seasonal water quality with an effective time
resolution of 2 weeks to 1 month.
Level III includes computerized one-dimensional (1-d) and quasi two-dimensional (2-d) dynamic
simulation models. These real-time models simulate variations in tidal heights and velocities throughout
each tidal cycle. One-dimensional models treat the estuary as well-mixed vertically and laterally. Quasi
2-d models employ a link-node approach that describes water quality in two dimensions (longitudinal and
lateral) through a network of 1-d nodes and channels. Tidal movement is simulated with a separate
hydrodynamic package in these models. The required data and modeling resources typically are
unavailable to support models of Level III or above on a widespread basis.
Level IV consists of computerized 2-d and 3-d dynamic simulation models. Dispersive mixing and
seaward boundary exchanges are treated more realistically than in the Level III 1-d models. These
models are almost never used for routine nutrient assessment, because of excessive resource
requirements.
Nutrient Criteria—Estuarine and Coastal Waters 9-9
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Level I Models
Level I desktop methodologies may be employed using a hand-held calculator or computer spreadsheet
and are based on steady-state conditions, first-order decay coefficients, simplified estimates of flushing
time, and seasonal pollutant concentrations. The EPA screening methods provide a series of Level I
analyses as described below.
EPA Screening Methods
WQAM is a set of steady-state desktop models that includes both one -dimensional and two-dimensional
box model calculations (Mills et al. 1985). Specific techniques contained in WQAM are the Fraction of
Freshwater Method, the Modified Tidal Prism Method, Advection-Dispersion Equations, and Pritchard's
Two-Dimensional Box Model.
Fraction of Freshwater Method
The Fraction of Freshwater Method estimates pollutant concentrations in one-dimensional estuaries from
information on freshwater and tidal flow by comparing salinity in the estuary with salinity in the local
seawater. The fraction of freshwater at any location in the estuary is calculated by comparing the volume
of freshwater at that location with the total volume of water:
Os ~ ijx
fa = — (9-D
where ^ = fraction of freshwater at location x, Ss = seawater salinity at the mouth of the estuary, and Sx =
salinity at location x.
This ratio can be viewed as the degree of dilution of the freshwater inflow (as well as pollutants) by
seawater. With this in mind, the total dilution of a pollutant input can be calculated by multiplying the
seawater dilution by the freshwater dilution. This then provides a simple way to calculate concentrations
of conservative pollutants. For any location x, at or downstream of the discharge, pollutant loads are
diluted by tidal mixing and upstream flows. The amount of dilution can be calculated by:
w
Cx = fx— (9-2)
where Cx = constituent concentration at location x at or downstream of discharge,^ = fraction of
freshwater at location x, W= waste loading rate (mass/time), and Q = freshwater inflow (volume/time).
The right side of Equation 9-2 can be divided into two distinct terms. The term W/Q represents the
classical equation for determining dilution in rivers caused by upstream flow. The second term, fx,
accounts for the further dilution of the river concentration by tidal influx of seawater.
9-10 Nutrient Criteria—Estuarine and Coastal Waters
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Concentrations upstream of the discharge are estimated from the concentration at the point of mix and
the relative salinity of the upstream location. The upstream concentrations are assumed to be diluted by
freshwater to the same degree that salinity is diluted. The equation is:
„ ,w s*
Cx = fd (9-3)
QS*
where/, = fraction of freshwater at discharge location, Sx = salinity at upstream location x, and Sd =
salinity at discharge location.
The fraction of freshwater at the discharge location,/,, is determined by applying Equation 9-1 at the
discharge location. Equation 9-3 can be modified to assess the impact of nutrients entering from the
seaward boundary by replacing the leading fdW/Q term with the boundary nutrient concentration.
Cumulative pollutant impacts from multiple sources are obtained through a two-step process. First,
pollutant concentration caused by each source independent of all other discharges must be determined.
This determination is accomplished by applying Equation 9-2 or 9-3, one discharge at a time, for any
estuary location of interest. The second step is to determine the total concentration at that location. This
determination is accomplished by adding all of the incremental concentrations caused by each discharge,
as calculated in the first step. This process can be repeated for any location of interest.
The Fraction of Freshwater Method can be used to predict cumulative impacts in one-dimensional (i.e.,
narrow) systems with significant freshwater inflow. Upstream freshwater flow must be large with
respect to total pollutant inflow for this method to be applicable. The method assumes conservative
pollutant behavior. It is consequently best used to investigate total nitrogen concentrations, because
overall loss rates of total nitrogen from the water column generally are small.
Modified Tidal Prism Method
The Modified Tidal Prism Method estimates dilution from the total amount of water entering an estuary.
It is more powerful than the Fraction of Freshwater Method because it can consider not only tidal
dilution but also nonconservative reaction losses. It is best applied to investigate total nutrient
concentrations, but provides additional flexibility to describe pollutant losses that may occur through
settling or denitrification.
The method divides an estuary into segments whose lengths and volumes are calculated using low-tide
volumes and tidal inflow. The tidal prism (i.e., total volume of tidal inflow) is compared for each
segment with the total segment volume to estimate flushing potential in that segment over a tidal cycle.
The Modified Tidal Prism Method assumes complete mixing of the incoming tidal flow with the water
resident in each segment.
Nutrient Criteria—Estuarine and Coastal Waters 9-11
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The Modified Tidal Prism Method requires seven inputs:
Freshwater inflow to the estuary
• Salinity of seawater at the downstream boundary
Pollutant loading rate
Salinity of each segment
• Low-tide volume for each segment
Intertidal volume (tidal prism) for each segment
• First-order constituent loss rate for each segment
The first step of the method is to segment the estuary. This requires an initial time-consuming step of
dividing the estuary into segments with lengths equal to the distance traveled by a particle over a tidal
cycle. Cumulative subtidal and intertidal water volumes must be plotted for the estuary, and a graphical
procedure is used to define model segmentation. Once the estuary has been segmented, a series of
calculations can be performed to estimate constituent concentrations in each segment. Specific methods
for dividing the estuary and performing the calculations are provided in Mills et al. (1985).
Advection-Dispersion Equations
Analytical equations have been developed to predict the concentration of nonconservative constituents in
one-dimensional estuaries. These types of equations consider the processes of net seaward flows
(advection) and tidally averaged mixing (dispersion), as well as simple decay. They can be used to
predict total nutrient concentrations at various locations in an estuary in response to alternative nutrient
loading rates. One-dimensional advection-dispersion equations can be expressed in several different
forms (O'Connor 1965), with the most common form contained in the water quality assessment
methodology. These equations require numerous simplifying assumptions, such as constant geometry
and tidal mixing along the length of the estuary, but have proven to be a useful screening tool.
The advection-dispersion equations require five inputs: upstream freshwater flow rate (R), constituent
loading rate (W), estuarine cross-sectional area (^4), tidally averaged dispersion coefficient (£), and first-
order decay rate coefficient (k). The first three inputs can be measured directly. The latter two inputs
must be determined indirectly through the model calibration process described below. Two equations are
provided, one which predicts concentrations at any distance (x) upstream of the discharge of concern and
another for concentrations at any distance seaward of the discharge. Co is the concentration at the point
of discharge. The equations are:
C = C0eJ2X X>0 (down estuary) (9-4)
C = C0eJlX X<0 (up estuary)
where:
9-12 Nutrient Criteria—Estuarine and Coastal Waters
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R
J\ =
w
Co =
R
These equations can be used to evaluate multiple loading sources by independently applying Equation
9-4 for each loading source and summing the predicted concentrations across the estuary.
Pritchard's Two-Dimensional Box Model
Vertically stratified estuaries add a significant degree of complexity to the modeling analysis. Pritchard
(1969) developed a relatively simple approach, which can predict nutrient concentration distributions
along the length of an estuary in both an upper and lower layer. This approach is based on numerous
simplifying assumptions, including:
Steady-state conditions
• Conservative pollutant behavior
• Uniform constituent concentration within each layer or each segment
The following information is required: (1) freshwater flow rate into the head of the estuary, (2) pollutant
mass loading rates, and (3) longitudinal salinity profiles along the length of the estuary in the upper and
lower layers. The method solves a series of linear equations describing the salinity balance around each
segment to determine net flows and dispersion between each segment. Specific methods for performing
the calculations are provided in Mills et al. (1985).
Results from Pritchard's model can be used to directly calculate conservative constituent concentrations
throughout the estuary or to serve as the hydrodynamic input to one of the Level II models described
below.
Level II Models
Level II models include computerized steady-state and tidally averaged simulation models that generally
use a box or compartment-type network. Steady-state models are difficult to calibrate in situations where
hydrodynamics and pollutant releases vary rapidly. Consequently, these models are less appropriate
when waste load, river inflow, or tidal range vary appreciably with a period close to the flushing time of
the waterbody. Level II models are the simplest models available that are capable of describing the
Nutrient Criteria—Estuarine and Coastal Waters 9-13
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relationship between nutrient loads and some of the endpoints of concern of the eutrophication process
(i.e., chlorophyll a, minimum dissolved oxygen).
The Level II models by EPA are QUAL2E and the Water Quality Analysis Simulation Program
(WASPS), with its associated eutrophication program EUTRO5.
QUAL2E
QUAL2E is a steady-state, one-dimensional model designed for simulating conventional pollutants in
streams and well-mixed lakes (U.S. EPA 1995) and is not recommended for estuaries. Rather, QUAL
TX, which allows tidal boundary conditions, may be more appropriate, but documentation on this model
is very sparse.
WASP6.0
The Water Quality Analysis Simulation Program (WASP6.0) is a general, multidimensional model that
utilizes compartment modeling techniques (Ambrose et al. 1993). The equations solved by WASP6.0 are
based on the principle of conservation of mass. Operated in the quasi-dynamic mode, WASP6.0 requires
the user to supply initial segment volumes, network flow fields, and inflow time functions. The user also
must calibrate dispersion coefficients between compartments. WASP6.0 has the capability of simulating
nutrient-related water quality issues at a wide range of complexity.
EUTRO5
EUTRO5 is the submodel in the WASP6.0 system that is designed to simulate conventional pollutants.
EUTRO5 combines a kinetic structure adapted from the Potomac Eutrophication Model with the WASP
transport structure. EUTRO5 predicts DO, carbonaceous BOD, phytoplankton carbon and chlorophyll a,
ammonia, nitrate, organic nitrogen, organic phosphorus, and orthophosphate in the water column and, if
specified, the underlying bed. In addition to segment volumes, flows, and dispersive exchanges, the user
must supply deposition and resuspension velocities for organic solids, inorganic solids, and
phytoplankton. Rate constants and half-saturation coefficients for the various biochemical
transformation reactions must be specified by the user. Finally, the time- and/or space-variable
environmental forcing functions, such as light intensity, light extinction, wind speed, cloud cover,
temperature, and benthic fluxes, must be input.
Level III Models
Level III includes computerized 1-d and 2-d models that simulate variations in tidal height and velocity
throughout each tidal cycle. Level III models enable characterization of phenomena varying rapidly
within each tidal cycle, such as pollutant spills, stormwater runoff, and batch discharges. Level III
models also are deemed appropriate for systems where the tidal boundary impact, as a function of the
hydrodynamics and water quality, is important to the modeled system within a tidal period.
Tidally varying (intratidal) models have found most use in the analysis of short-term events, in which the
model simulates a period of time anywhere from one tidal cycle to a month. Some seasonal simulations
also have been conducted.
9-14 Nutrient Criteria—Estuarine and Coastal Waters
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In using Level III models, one must decide whether a 1-d longitudinal system is sufficient, or whether a
2-d model is required to capture the longitudinal and lateral variations in the estuary. For estuaries
whose channels are longer than their width and reasonably well mixed across their width, a 1-d model
may be chosen. If large differences exist in water quality from one side of an estuary to the other, or if
vertical stratification is important, then a 2-d model is appropriate.
All Level III models considered here can simulate nutrient-eutrophication interactions. These models
also include settling rates and benthic flux rates for several different constituents, such as phosphorus,
nitrogen, and sediment oxygen demand. The Level III model distributed by EPA is the WASP6.
Level IV Models
Level IV includes a variety of computerized 2-d and 3-d dynamic simulation models. Dispersive mixing
and seaward boundary exchanges are treated more realistically than in the Level III 1-d models.
Although not routinely used in nutrient analyses, they are now finding use by experts in special studies.
Level IV models are required when variations in concentrations in all three dimensions are of concern.
The time-variable nature of a Level IV model ensures the need for a time-variable watershed model in
order to provide for the nonpoint source inputs. Fully 3-d models that can predict longitudinal, lateral,
and vertical transport are the most complex and expensive to set up and run.
At present, no Level IV model is supported by EPA. Three current Level IV models, CE-QUAL-W2,
Integrated Compartment Model (ICM), and EFDC, are described below.
CE-QUAL-W2
CE-QUAL-W2 is a dynamic 2-d (x-z) model developed for stratified waterbodies (Environmental and
Hydraulics Laboratories 1986). This is a U.S. Army Corps of Engineers modification of the Laterally
Averaged Reservoir Model (Edinger and Buchak 1983; Buchak and Edinger 1984a,b). CE-QUAL-W2
consists of directly coupled hydrodynamic and water quality transport models. Hydrodynamic
computations are influenced by variable water density caused by temperature, salinity, and dissolved and
suspended solids. Developed for reservoirs and narrow, stratified estuaries, CE-QUAL-W2 can handle a
branched and/or looped system with flow and/or head boundary conditions. With two dimensions
depicted, point and nonpoint loadings can be distributed spatially.
CE-QUAL-W2 simulates as many as 20 other quality variables. Primary physical processes included are
surface heat transfer, shortwave and longwave radiation and penetration, convective mixing, wind- and
flow-induced mixing, entrainment of ambient water by pumped-storage inflows, inflow density current
placement, selective withdrawal, and density stratification as influenced by temperature and dissolved
and suspended solids. Major chemical and biological processes in CE-QUAL-W2 include the effects on
DO of atmospheric exchange, photosynthesis, respiration, organic matter decomposition, nitrification,
and chemical oxidation of reduced substances; uptake, excretion, and regeneration of phosphorus and
nitrogen and nitrification-denitrification under aerobic and anaerobic conditions; carbon cycling and
alkalinity-pH-CO2 interactions; trophic relationships for total phytoplankton; accumulation and
decomposition of detritus and organic sediment; and coliform bacteria mortality.
Nutrient Criteria—Estuarine and Coastal Waters 9-15
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CH3D-ICM
CH3D is a 3-d, finite-difference hydrodynamic model, developed by the U.S. Army Corps of Engineers
Waterways Experiment Station (WES) in Vicksburg, MS. Results from CH3D have been linked to the
ICM to model water quality in the Chesapeake Bay. The ICM was developed as the integrated-
compartment eutrophication model component of the Chesapeake Bay model package. The model
contains detailed eutrophication kinetics, modeling the carbon, nitrogen, phosphorus, silica, and
dissolved oxygen cycles.
CH3D-ICM is a linkage of CH3D, a hydrodynamic model, and ICM, a water quality model. CH3D is a
hydrodynamic model developed for the Chesapeake Bay Program (Johnson et al. 1991). The model can
be used to predict system response to water levels, flow velocities, salinities, temperatures, and the three-
dimensional velocity field. CH3D makes hydrodynamic computations on a curvilinear or boundary-fitted
platform grid. Deep navigation channels and irregular shorelines can be modeled because of the
boundary-fitted coordinates feature. Vertical turbulence is predicted by the model and is crucial to a
successful simulation of stratification, destratification, and anoxia. A second-order model based on the
assumption of local equilibrium of turbulence is employed.
ICM is a finite-difference water quality model that may be applied to most waterbodies in one, two, or
three dimensions (Cerco and Cole 1995). The model predicts time-varying concentrations of water
quality constituents and includes advective and dispersive transport. The model also considers sediment
diagenesis benthic exchange. ICM incorporates detailed algorithms for water quality kinetics.
Interactions among state variables are described in 80 partial-differential equations that employ more
than 140 parameters. An improved finite-difference method is used to solve the mass conservation
equation for each cell in the computational grid and for each state variable.
EFDC
EFDC is a linked three-dimensional, finite-difference hydrodynamic and water quality model developed
at the Virginia Institute of Marine Sciences (Hamrick 1996). EFDC contains extensive water quality
capabilities, including a eutrophication framework based on the ICM model. EFDC is a general-purpose
hydrodynamic and transport model that simulates tidal, density, and wind-driven flow; salinity;
temperature; and sediment transport. Two built-in, full-coupled water quality/eutrophication submodels
are included in the code.
EFDC solves the vertically hydrostatic, free-surface, variable-density, turbulent-averaged equations of
motion and transport; transport equations for turbulence intensity and length scale, salinity, and
temperature in a stretched, vertical coordinate system; and horizontal coordinate systems that may be
Cartesian or curvilinear-orthogonal. Equations describing the transport of suspended sediment, toxic
contaminants, and water quality state variables also are solved.
The model uses a finite-difference scheme with three time levels and an internal-external mode splitting
procedure to achieve separation of the internal shear, or baroclinic, mode from the external free-surface
gravity wave, or barotropic, mode. An implicit external-mode solution is used with simultaneous
9-16 Nutrient Criteria—Estuarine and Coastal Waters
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computation of a two-dimensional surface elevation field by a multicolor successive overtaxation
procedure. The external solution is completed by calculation of the depth-integrated barotropic velocities
using the new surface elevation field. Various options can be used for advective transport, including the
"centered in time and space" scheme and the "forward in time and upwind in space" scheme.
Summary of Model Capabilities
The important features of the models selected for discussion in this manual are summarized in Tables 9-1
and 9-2. The information provided in these tables is primarily qualitative and sufficient to determine
whether a model may be suitable for a particular application. For complete information, consult the
appropriate user's manuals, the supporting agency, and other experienced users.
9.4 USE OF MODELS FOR NUTRIENT INVESTIGATION
This section describes procedures for using models to perform nutrient assessment in estuaries and near-
coastal waters. It describes the model calibration and validation process, where model parameters that
best describe the waterbody of interest are selected. In addition, guidance is provided on using models
for nutrient management and assessment.
The first subsection describes a general procedure for calibrating nutrient models, and briefly describes
the validation procedure used to estimate the uncertainty of such models. The subsection also describes
some statistical methods for testing the calibrated models. These methods are useful to aid in the various
calibration phases and also in the validation phase to measure how well model predictions and
measurements of water quality agree.
The second subsection provides guidance on the management application of a calibrated model.
Methods to project effects of changes in waste loads and to determine causes of existing conditions are
discussed. Finally, a case study application is provided.
Model Calibration and Validation
Model calibration is the process of determining model parameters most appropriate for a given site-
specific application. Calibration of a model involves a comparison of the measured and simulated
receiving water quality conditions. The nature of the model calibration process depends upon the
complexity of the model selected. Simpler models contain relatively few parameters that need to be
calibrated, whereas more complex models contain many.
Calibration alone is not adequate to determine the predictive capability of a model for a particular
estuary. To map out the range of conditions over which the model can be used, one or more additional
independent sets of data are required to determine whether the model is predictively valid. This model
validation exercise defines the limits of usefulness of the calibrated model. Without validation testing,
the model merely describes the conditions defined by the calibration data set. The degree of uncertainty
of any projection or extrapolation of the model remains unknown.
Nutrient Criteria—Estuarine and Coastal Waters 9-17
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Table 9-1. Basic model features
Methods/Model
Fraction of
Freshwater
Modified Tidal
Prism
Advection-
Dispersion
Equations
Pritchard's 2-D
Box Model
QUAL2E
WASPS
CE-QUAL-W2
CH3D-ICM
EFDC
Time
Scales
SS
ss
SS
ss
ss
Q/D
D
D
D
Spatial
Dimensions
ID
ID
ID
2D(xz)
ID
ID, 2D (xy), or
3D
2D(xz)
3D
3D
Hydro-
dynamics
0
0
0
0
I
I, S
s
s
s
Data Expertise
Requirements
Minimal
Minimal
Minimal
Minimal
Moderate
Moderate to
substantial
Substantial
Extreme
Extreme
Distributing
Agency
EPA
EPA
EPA
EPA
EPA
EPA
Army Corps
EPA
EPA
Scale of
Effort
Days
Days
Days
Days
Few months
Few months
Several months
Months to years
Months to years
D - dynamic; Q - quasi-dynamic (tidal-averaged); SS - steady state; x - 1-dimensional, xy - 2-dimensional, longitudinal-lateral;
xz - 2-dimensional, longitudinal-vertical; xyz - 3-dimensional; B - compartment or box 3d; xx - link node branching 2d; 0 - No
hydraulics specified, inferred from salinity data; I - hydrodynamics input; S - hydrodynamics simulated.
In general, models are calibrated in phases, beginning with the selection of the model parameters and
coefficients that are independent of parameters to be calibrated later. For purposes of this discussion, the
process is divided into the categories of hydrodynamic calibration and water quality calibration. The
discussion covers the parameters that need to be calibrated for each level of model as well as the specific
model outputs to be used for the calibration comparison. Calibration of the more complex models
requires detailed guidance; the reader is referred to other documents (e.g., U.S. EPA 1990b; Thomann
and Mueller 1987) for a discussion that is more comprehensive than is feasible here.
Hydrodynamic Calibration
The first phase of calibration concentrates on the hydrodynamic and mass transport models. Two Level I
models, the Fraction of Freshwater Method and the Tidal Prism Method, have no hydrodynamic
parameters that require calibration. In these simplest cases, all hydrodynamic and mass transport
processes are implicitly considered via specification of observed salinity values. Although there are no
parameters to calibrate for these models, there is merit in testing the model's predictive validity by
comparing predicted concentrations with field observations of a conservative (i.e., nondecaying)
substance, if such data are available.
For the remaining Level I and Level II models, only one hydrodynamic parameter requires calibration:
the tidal dispersion coefficient. It is possible to calibrate the hydrodynamic and mass transport portions
of these models by determining values for this coefficient that best describe observed salinity or
conservative tracer measurements.
9-18
Nutrient Criteria—Estuarine and Coastal Waters
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Table 9-2. Key features of selected models
Model
Key Features
Advantages
Disadvantages/ Limitations
WQAM
Simplified equations to
simulate dilution, advection,
dispersion, first-order decay,
empirical relationships
between nutrient loading,
and total nutrient
concentration
Few data requirements;
can be employed easily
with a hand calculator or
computer spreadsheet
Limited to screening and
midlevel applications
QUAL2E
Steady-state model provides
adequate simulation of water
quality processes, including
DO-BOD and algal growth
cycles
User-friendly Windows
interface; widely used and
accepted; able to simulate
all of the conventional
pollutants of concern
Limited to simulation of time
periods during which stream
flow and input loads are
essentially constant
WASPS
Based on flexible
compartment modeling
approach; can be applied in
1, 2, or 3 dimensions
Has been widely applied
to estuarine situations;
considers comprehensive
DO and algal processes;
can be used in 3-d
simulations by linking
with hydrodynamic
models
Coupling with multi-
dimensional hydrodynamic
models requires extensive
site-specific linkage efforts
CE-QUAL-W2
Uses an implicit approach to
solve equations of continuity
and momentum; simulates
variations in water quality in
the longitudinal and lateral
directions
Simulates the onset and
breakdown of vertical
stratification; most
appropriate where vertical
variations are an
important water quality
consideration
Application requires
extensive modeling
experience
CH3D-ICM
Finite-difference model can
be applied to most water
bodies in 1 to 3 dimensions;
predicts time-varying
concentrations of
constituents; includes
advective and dispersive
transport
State-of-the-science
eutrophication kinetics
Computationally intensive;
requires extensive data for
calibration and verification;
requires a high level of
technical expertise to apply
effectively
EFDC
Linked 3-d, finite-difference
hydrodynamic and water
quality model contains
extensive water quality
capabilities; water quality
concentrations can be
predicted in a variety of
formats suitable for analysis
and plotting
3-d description of water
quality parameters of
concern; entire range of
hydrodynamic, sediment,
eutrophication, and toxic
chemical constituents can
be considered
Computationally intensive;
requires extensive data for
calibration and verification;
requires a high level of
technical expertise to apply
effectively
Nutrient Criteria—Estuarine and Coastal Waters
9-19
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Level III models typically contain two calibration parameters, the channel roughness coefficient and the
dispersion coefficient. Occasionally these models are calibrated with current velocity and water surface
elevation data, but more often are indirectly calibrated from salinity or conservative tracer measurements
that also must be used to calibrate the mass transport model. Indirect calibration can result in an
imprecise description of both the circulation and mass transport algorithms, but this is not a severe
drawback unless the critical water quality components of the waste load allocation model are sensitive to
small changes in circulation and mass transport.
Level IV hydrodynamic models contain several calibration parameters, including bottom and surface
friction coefficients; vertical, lateral, and horizontal eddy viscosity coefficients; and wind speed
coefficients. Calibration efforts for these types of models are beyond the scope of this document.
Kinetic Process Calibration
The second phase of calibration involves selection of the set of kinetic coefficients describing nutrient
cycles and other aspects of the eutrophication process. Again, the effort required is directly related to the
complexity of the model selected.
Two Level I models—the Fraction of Freshwater Method and Pritchard's model—have no kinetic
parameters that require calibration. The models assume that constituent concentrations undergo no
kinetic processes that affect their concentration, and typically are appropriate only for estimating total
nitrogen concentrations. The remaining Level I models can describe nonconservative constituents, and
lump all kinetic processes into a single overall decay coefficient. Model calibration in these cases
consists of a comparison of predicted versus observed total nutrient concentrations.
The calibration of higher level nutrient and phytoplankton models requires significant expertise because
of the complexity of the interactions between a number of the components of the cycles involved.
Coefficients that require calibration in these models pertain to: transformation rates among various forms
of a given nutrient; maximum phytoplankton growth rates; phytoplankton respiration rates;
phytoplankton growth sensitivity to light and nutrients; and phytoplankton and detrital settling velocities.
Model Validation
Validation testing is designed to confirm that the calibrated model is useful at least over the limited range
of conditions defined by the calibration and validation data sets. The procedure is not designed to
validate a model as generally being useful in every estuary, or even as useful over an extensive range of
conditions found in a single estuary. Validation, as employed here, is limited strictly to indicating that
the calibrated model is capable of producing valid results over a limited range of conditions. Those
conditions are defined by the sets of data used to calibrate and validate the model. As a result, it is
important that the calibration and validation data cover the range of conditions over which predictions
are desired.
9-20 Nutrient Criteria—Estuarine and Coastal Waters
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Validation testing is performed with an independent data set collected during a second field study. The
field study may occur before or after the collection of calibration data. For the best results, however, the
validation data should be collected after the model has been calibrated. This schedule of calibration and
validation ensures that the calibration parameters are fully independent of the validation data. Often it is
difficult to assemble the necessary resources to conduct the desired number of surveys. Therefore, it is
important that surveys be scheduled in an innovative manner and the choice of calibration and validation
data sets remain flexible to make the test of the calibrated model as severe as possible.
Too often, limited studies attempt calibration but not validation. This approach, in effect, limits the
study to describing the conditions during the calibration data collection period and increases the
uncertainty associated with the waste load allocation. In fact, model prediction uncertainty cannot be
reliably assessed in these cases.
Model Testing
During and after the calibration and validation of a model, at least two types of testing are important.
First, throughout the calibration procedure, a sensitivity test helps determine which parameters and
coefficients are the most important. Second, a number of statistical tests help define the extent of
agreement between model simulations and measured conditions.
The sensitivity analysis is simply an investigation of how much influence changes in model coefficients
have on simulated results. Typically, important coefficients, parameters, boundary conditions, and initial
conditions are varied by a positive or negative constant percentage to see what effect the change has on
critical predictions. The coefficients and parameters are changed one at a time and the effects typically
are ranked to show which parameters have the most influence and which have the least.
The second type of testing involves assessment of the "goodness of fit" for model simulations, compared
with measurement of important water quality parameters. In addition to making a visual assessment, a
number of statistical tests have proven useful. These include root mean square error, relative error, and
regression analysis. Other more detailed statistical analyses are described in U.S. EPA (1990b).
The root mean square (rms) error is a criterion that is widely used to evaluate the agreement between
model predictions. The rms error can be defined as:
i0.5
(9-5)
where Cm = measured concentration, Cs = simulated concentration, and N = number of measurements.
The rms error can be used to compute simultaneous discrepancies at a number of points, or it can be used
to compute discrepancies between measurements and predictions at a single point over time. Global rms
errors can be computed for a series of measurements at multiple points over time.
Nutrient Criteria—Estuarine and Coastal Waters 9-21
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When discrepancies between model simulations and measurements are not uniform over parts of the
estuary or over time, the relative error may be a more appropriate statistic for testing calibration or
validation. The relative error is defined as:
e= V- (9-6)
where the overbars denote the average measured or simulated value. Averages can be performed over
multiple sites or over time. The relative error behaves poorly for small values of measurements if
discrepancies are not proportional to the magnitude of the measurement (i.e., small values of Cm magnify
discrepancies) and if Cm > Cs (as the maximum relative error is usually taken to be 100%). Therefore, the
relative error is best for computing composite statistics when discrepancies are not constant, as may
occur when calibration over an extensive range is attempted.
Regression analysis is very useful in identifying various types of bias in predictions of dynamic-state
variables. The regression equation is written as:
Cm = a + bCs + 8 (9-7)
where a = intercept value, b = slope of the regression line, and e = the error in measurement mean, Cm.
The standard linear regression statistics computed from the above equation provide insight into the
goodness of fit for a calibration. The square of the correlation coefficient, r2, measures the percent of the
variance accounted for between measured and predicted values. The slope estimate, b, and intercept, a,
can indicate any consistent biases in the model calibration. A model calibration that perfectly described
all available data would have a correlation coefficient of 1.0, a slope of 1:1, and a zero intercept.
9.5 MANAGEMENT APPLICATIONS
Once the model is calibrated and validated, it can be used to simulate future conditions to determine
effects of changes in waste loads or to investigate causes of existing problems. This section describes
three types of management application: (1) load-response analysis, (2) determination of acceptable
nutrient loads, and (3) investigation of causes of nutrient problems.
Load-Response Analysis
A load-response analysis consists of performing multiple model simulations using different loading rates
and examining the water quality predicted for each simulation. The most common use of a model to
investigate nutrients in estuaries is to determine the water quality throughout an estuary in response to
changes in nutrient loads. Models are designed to predict water quality based on loadings and
environmental conditions (Figure 9-2).
This type of analysis also requires specification of the environmental conditions (e.g., freshwater inflows,
tidal conditions) to be considered. The results of the load-response analysis are directly related to the
9-22 Nutrient Criteria—Estuarine and Coastal Waters
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environmental conditions specified for the model simulation. For example, use of summer-average
environmental conditions in the model will show the response in summer-average water quality to
changes in loads. For the simplest Level I models, environmental conditions are specified implicitly
through the use of salinity observations. Predictions from these models will correspond to the
environmental conditions that were in effect when the salinity was measured. The more complex models
require explicit definition of environmental condition and can be used to provide predictions for a wide
range of environmental conditions.
Acceptable Nutrient Loads
The most common use for water quality models is to define allowable loads necessary to achieve water
quality objectives. As seen in Figure 9-2, models predict water quality for a specified set of loads and
environmental conditions. Determination of acceptable loads typically requires an iterative procedure, as
shown in Figure 9-3. The first iteration consists of performing a model simulation using existing loads
and comparing predicted water quality with objectives. Assuming that the existing loads result in
unacceptable water quality, additional model simulations are performed using incremental reductions in
nutrient loads until water quality objectives are achieved.
The approach shown in Figure 9-3 can be used to define necessary reductions in total loads as well as
reductions in individual contributors to the total load.
The results of the above approach are highly dependent on the environmental conditions selected, as
allowable loading rates can vary substantially across different environmental conditions. Two
approaches are available for selecting critical conditions for use in defining allowable loads. These
approaches are termed the critical conditions approach and the continuous simulation approach. In the
critical conditions approach, a single set of environmental conditions is selected for analysis. These
conditions typically represent critical or worst-case conditions, that is, those environmental conditions
that will result in the poorest water quality for a given set of loads. The rationale for the critical
conditions approach assumes that if loads are defined to meet water quality objectives during "critical"
conditions, the same loads will result in attaining water quality objectives during most other conditions as
well.
Pollutant Loads
Environmental
Conditions
Water Quality
Model
Predicted
Water Quality
Figure 9-2. Use of models in load-response analysis.
Nutrient Criteria—Estuarine and Coastal Waters
9-23
-------
Pollutant Loads
Environmental
Conditions
Water Quality
Model
Predicted
Water Quality
Reduce Loads
no
Acceptable?
yes
Done
Figure 9-3. Use of models in determining allowable loads.
The continuous simulation approach performs simulations for as long a duration as is feasible, using
historically observed variations in environmental conditions. The predicted water quality resulting from
the continuous simulation is analyzed to determine the frequency with which water quality problems are
observed to occur.
The overall intent of the modeling analysis is to define loads that will restrict the occurrence of water
quality problems to an acceptable frequency. Each of the above two methods has particular strengths and
weaknesses for performing this task. The continuous simulation approach provides a direct means to
consider frequency of occurrence (i.e., number of years per problem) but has extreme resource
requirements. The ability to perform continuous simulation of sufficient duration typically is constrained
by the availability of data describing historical environmental conditions, or by computational
requirements for the higher level models. The critical conditions approach has much more manageable
resource requirements; however, there are no clear methods to establish appropriate critical conditions.
In estuaries, freshwater, tides, wind, and other factors all can affect water quality. Selection of
appropriate values for each environmental parameter requires considerable judgment. Furthermore, the
specific level of protection associated with any single set of environmental conditions cannot be
evaluated without performing a continuous simulation.
Case Study Example
Water quality models also can be used to gain an increased understanding of the relative importance of
various loading sources to an estuary or near-coastal water. It is possible to investigate the contribution
of individual loading sources to the water quality problem by performing a series of simulations
examining each loading source separately. Because most water quality models assume a linear
relationship between pollutant load and resulting water quality impact, it is possible to determine overall
impacts to the estuary by summing the impacts from each source.
9-24
Nutrient Criteria—Estuarine and Coastal Waters
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Investigation of Causes of Nutrient Problems
Shipps Creek (Figure 9-4) is a long, narrow, tidal tributary receiving nutrient inputs from upstream runoff
and a single wastewater treatment plant (WWTP). This case study example demonstrates the use of
models to perform three tasks:
Estimate the contribution of various loading sources to the overall summer-average total nitrogen
concentration
Estimate the effect of a 50% reduction in loads from the WWTP on total nitrogen concentrations
throughout the estuary
Estimate the reduction in loading necessary to achieve an average total nitrogen concentration of
0.100 mg/L in the lower half of the estuary
Short-term answers were required, and screening-level accuracy was judged acceptable because of the
short timeframe and limited data available.
The Fraction of Freshwater Method was selected because the estuary was considered one-dimensional,
long-term average results were acceptable, and the water quality target was specified in terms of total
nutrient levels.
City
WWTP
Figure 9-4. Shipps Creek site map and salinity monitoring location.
Nutrient Criteria—Estuarine and Coastal Waters
9-25
-------
Information available to support the study included summer-average salinity measurements at 13
locations along the length of the estuary, summer-average freshwater inflows and total nitrogen loads to
the estuary, nitrogen loads from the WWTP, and nitrogen and salinity concentrations outside of the
estuary. This information was compiled as shown in Table 9-3 to allow implementation of the Fraction
of Freshwater Method. The top two rows of Table 9-3 show the measured loading rates and seaward
boundary conditions. The two leftmost columns define segments centered around each of the salinity
measurements. The third column applies Equation 9-1 to calculate the fraction of freshwater in each
segment.
The fourth through sixth columns in Table 9-3 apply Equations 9-2 or 9-3 as appropriate to determine the
incremental contribution to the total nitrogen (TN) concentrations from each of three possible
contributing sources: upstream (TNup), wastewater treatment plant (TNwwxp), and the downstream
seaward boundary (TNSEA). Equation 9-2 is applied only to those segments downstream of the loading
source, whereas Equation 9-3 is applied to those segments upstream of the loading source. Equation 9-2
is applied to all segments for examining impacts from upstream sources, whereas Equation 9-3 is applied
to all segments for examining impacts from the seaward boundary. For determining WWTP impacts,
Equation 9-2 is applied to segments 1-11 and Equation 9-2 is applied to segments 12-13. The final
column in Table 9-3 sums the incremental contributions from each of the sources to provide a prediction
of overall TN concentrations throughout the estuary.
Table 9-3. Calculation spreadsheet for Shipps Creek estuary
Freshwater Inflow
0 = 100
Segment
#
1
2
3
4
5
6
7
8
9
10
11
12
13
000 cmd
Salinity, S,
(PPt)
29
27
25
23
21
19
18
16
14
12
10
5
1
Seawater Salinity
Ss = 30 DDt
Fraction of
Freshwater, f,
0.03
0.10
0.17
0.23
0.30
0.37
0.40
0.47
0.53
0.60
0.67
0.83
0.97
Upstream Load
W = 5,000 e/dav
TNup
(mg/L)
0.002
0.005
0.008
0.012
0.015
0.018
0.020
0.023
0.027
0.030
0.033
0.042
0.048
WWTP Load
W = 500,000 e/dav
1-NwWTP
(mg/L)
0.017
0.050
0.083
0.117
0.150
0.183
0.200
0.233
0.267
0.300
0.333
0.110
0.025
Seawater TN
0.005 DDIII
TNSEA
(mg/L)
0.005
0.005
0.004
0.004
0.004
0.003
0.003
0.003
0.002
0.002
0.002
0.001
0.000
Overall TN
(mg/L)
0.023
0.060
0.096
0.132
0.169
0.205
0.223
0.259
0.296
0.332
0.368
0.152
0.074
9-26
Nutrient Criteria—Estuarine and Coastal Waters
-------
The results of this modeling analysis are shown graphically in Figure 9-5, showing the overall TN
distribution as well as its component. Although the Fraction of Freshwater Method does not require
calibration, it would be worthwhile at this point to confirm that the model predictions of TN throughout
the estuary were consistent with observed data collected over the same time period. Figure 9-5 shows
that the WWTP is the dominant source of nitrogen throughout most of the estuary. Upstream sources are
the dominant component only at the extreme head of the estuary. Nitrogen contributions from the
seaward boundary are small throughout the system. The results in Table 9-2 and Figure 9-4 satisfy the
first objective of this study, which was to determine the contribution of various loading sources to the
overall summer-average total nitrogen concentration. The second objective of the study was to determine
the water quality resulting from a 50% reduction in WWTP TN loads. This was accomplished by
reapplying Equations 9-2 and 9-3 using one-half of the original WWTP loads. Results of this analysis
are shown in Figure 9-6, indicating a decrease in peak TN concentrations from 0.368 to 0.202 mg/L and a
decrease in lower estuary (defined as segments 1-6) average concentrations from 0.111 to 0.064 mg/L.
This nearly 50% reduction in concentrations was expected, because the original analysis had
demonstrated that the WWTP was the dominant loading source to the estuary.
The final objective of the study was to determine the loading reductions necessary to achieve a lower
estuary average concentration of 0.08 mg/L. No single answer exists to this question, because three
separate sources of nitrogen to the estuary contribute to the total concentration. Analysis of the data in
Table 9-3 shows that the incremental contribution of the upstream, WWTP, and seaward sources to lower
estuary average concentrations were 0.010, 0.100, and 0.004 mg/L, respectively. Because the seaward
Total VWVTP
Seaward Upstream
13
11
975
Segment Number
0.50
- 0.45
- 0.40
- 0-35 „_,
O)
h 0.30 £
c
01
0.25 g>
Jus
0.20 1
o
0.15 "~
0.10
0.05
0.00
Figure 9-5. Model results for existing conditions.
Nutrient Criteria—Estuarine and Coastal Waters
9-27
-------
Total
Seaward
-WWTP
Upstream
0.50
0.45
0.40
- 0.35 _
1
0.30 .§.
+J
0.20 •=
I
0.15 "~
0.10
h 0.05
0.00
13
11
975
Segment Number
1
Figure 9-6. Model results for 50% reduction in WWTP load.
boundary nitrogen concentration cannot be controlled, management reductions must be restricted to
either the upstream sources or the WWTP. The WWTP load must be reduced by at least 20% to meet the
target TN concentration of 0.080 mg/L; otherwise, its contribution alone will exceed the target. Beyond
the initial 20% reduction in the WWTP source, further reductions must come either from the WWTP or
upstream sources. The specific allocation of these load reductions among sources is an economic and
social decision that the model is not designed to address. The model is expressly designed, however, to
test alternative proposals of load reductions to determine if they will meet the water quality objective.
For example, a 25% reduction in both upstream and WWTP sources resulted in an average concentration
of 0.087 mg/L (i.e., above the target), but a 30% reduction in WWTP loads coupled with a 40%
reduction in upstream loads was shown to just meet the target.
All model simulations presented here should be viewed with extreme caution, because they are based on
an uncalibrated, screening-level model. The level of uncertainty for these predictions cannot be
quantified and is expected to be quite large. The model results do, however, provide the best possible
estimate describing the relationship between nutrient loads and resulting concentrations, given the
available resources.
Overview of Chesapeake Bay Airshed, Watershed, and Estuary Models
The cross-media models used in the Chesapeake Bay analysis consist of three models: an airshed model,
a watershed model, and a model of the Chesapeake estuary. These models are linked so that the output
of one simulation provides input data for another. The simulation period is the 10-year period of January
9-28
Nutrient Criteria—Estuarine and Coastal Waters
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1, 1985, to December 31, 1994. Versions of these models have been used by the Chesapeake Bay
Program for more than a decade and have been refined and upgraded several times.
AirshedModel
The Chesapeake Bay Program airshed model provides estimates of atmospheric deposition loads of
nitrogen. A product of the EPA National Exposure Research Laboratory in Research Triangle Park,
North Carolina, RADM (pronounced "radum") is an acronym for Regional Acid Deposition Model.
RADM is a three-dimensional model that tracks nutrient emissions across the eastern United States. Two
RADM grids meet various resolution needs. A large grid scale covers the entire RADM domain and
contains 20,000 square cells of 6,400 square kilometers each. A fine grid scale covers the region of the
Chesapeake Bay watershed and has 60,000 cells, each covering 400 square kilometers. The model
domain in the vertical is 15 cells deep, reaching from ground level to the top of the free troposphere. The
depth of the cells increases with altitude. One of the findings of the RADM model is that the Chesapeake
Bay airshed, defined as the area accounting for 75% of the deposition in the watershed, is approximately
5.5 times the size of the watershed.
RADM is used to drive scenarios associated with reductions in atmospheric deposition of nitrogen. A
base condition deposition represents an estimate of the current condition of atmospheric deposition in the
watershed and is developed from a regression of National Atmospheric Deposition Program (NADP)
data. RADM scenarios of atmospheric deposition reductions are incorporated into Watershed Model
scenarios by adjusting the base NADP condition on a segment-by-segment basis with a percent change
prescribed by RADM scenario results. Results from RADM specify loads of wet and dry deposition to
the Chesapeake watershed for the State Implementation Program (SIP) and Limit of Technology
scenarios. Deposition loads are input directly to the land surfaces of the watershed model or to the tidal
water surface of the Chesapeake Bay estuary model package as daily loads of wet deposition (from rain
washout of atmospheric nitrate and ammonia) and 12-year average loads of dry deposition.
Three atmospheric deposition loads were used for the Chesapeake analysis: (1) the base condition of
atmospheric deposition, (2) the estimated atmospheric deposition of nitrogen equivalent to the 1998 SIP
controls of atmospheric deposition, and (3) the estimated atmospheric deposition under full limit of
technology control (see Table 9-4).
Table 9-4. Chesapeake watershed nitrogen deposition under varying management schemes for emissions of
nitrogen atmospheric deposition precursors
Scenario
Base Condition
State Implementation Plan (SIP)
Limit of Technology
TN Deposition (millions of kg/year)
204
178
128
Sources: Chesapeake Bay Program Phase IV Watershed Model and PvADM.
Nutrient Criteria—Estuarine and Coastal Waters 9-29
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Watershed Model
The watershed model simulates nutrient and sediment loads delivered to the Chesapeake Bay from all
areas of the watershed. Land uses of cropland, pasture, urban areas, and forests are simulated on an
hourly time-step, tracing the fate and transport of input nutrient loads from atmospheric deposition,
fertilizers, animal manures, and point sources. The simulation is an overall mass balance of nitrogen and
phosphorus nutrients in the basin, so that the ultimate fate of input nutrients is simulated, either as
incorporation into crop or forest plant material, incorporation into soil, or river runoff. Nitrogen fates
included volatilization into the atmosphere and denitrification. Transport in rivers is simulated to the
tidal waters of the Chesapeake Bay. Sediment is simulated as eroded material washed off land surfaces
and transported to the tidal bay. Scenarios are run for 10 years on a 1-hour time-step, and results are
aggregated into 10-year average loads for comparison among scenarios.
To simulate the delivery of nutrients and sediment to the bay, the watershed was divided into 86 major
model segments, each with an average area of 194,000 hectares. Segmentation, based on three tiers of
criteria, partitioned the watershed into regions of similar characteristics. The first criterion was
segmentation of similar geographic and topographic areas, which were further delineated in terms of soil
type, soil moisture holding capacity, infiltration rates, and uniformity of slope. The second criterion
involved finer segmentation based on spatial patterns of rainfall. Each segment had a bank-full travel
time of about 24-72 hours. The third criterion used to further delineate segments was based on features
of the river reach. River reaches containing a reservoir were separated into a reservoir simulation and a
river simulation of the free-flowing river. For example, the James basin had 11 model segments, 2
represented reservoirs on the James and Appomattox, and the segmentation generally became finer with
closer proximity to tidal waters.
Model segments were located to take advantage of observed data locations, so that a model segment
outlet was located close to monitoring stations. Water quality and discharge data were collected from
Federal and State agencies, universities, and other organizations. More than 150 subsegments were used
at the interface between the watershed and estuary models to accurately deliver flow, nutrient, and
sediment loads to appropriate areas of the estuary. Increased simulation accuracy motivated the division
of basins into multiple segments and into simulation time-steps of an hour, but all scenario results were
reported at the level of the basin and for 10-year average loads.
The watershed model has been in continuous operation at the Chesapeake Bay Program since 1982 and
has had many upgrades and refinements since that time. The watershed model used for this application
was Phase 4.1 based on the HSPF Version 11 code (Hydrologic Simulation Program - Fortan - HSPF).
Version 11 is a widely used public domain model supported by EPA, USGS, and the U.S. Army Corps of
Engineers.
Estuary Model
The Chesapeake Bay Estuary Model Package (CBEMP) is actually several models simulating different
aspects of water quality in the bay and tributaries. A water quality model simulates 22 parameters, or
state variables, as listed in Table 9-5.
9-30 Nutrient Criteria—Estuarine and Coastal Waters
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Table 9-5. Water quality state variables used in CBEMP
Temperature
Salinity
Inorganic suspended sediments
Diatoms
Cyanobacteria (blue-green algae)
Other phytoplankton
Dissolved organic carbon
Labile paniculate organic carbon
Refractory paniculate organic carbon
Ammonium
Nitrate + nitrite
Dissolved organic nitrogen
Labile paniculate organic nitrogen
Refractory paniculate organic nitrogen
Total phosphorus
Dissolved organic phosphorus
Labile paniculate organic phosphorus
Refractory paniculate organic phosphorus
Dissolved oxygen
Chemical oxygen demand
Dissolved silica
Paniculate biogenic silica
Zooplankton were separated into two size classes: microzooplankton (>44 microns) and
mesozooplankton (>202 microns).
Linked to the water quality model is a hydrodynamic model, simulating the hydrodynamics, or water
movement, throughout the tidal estuary. The hydrodynamic model produced three-dimensional
predictions of velocity, diffusion, surface elevation, salinity, and temperature on an intratidal time scale.
The model grid of the hydrodynamic and water quality models consists of more than 10,000 cells.
The modeling process involves simulation of living resource parameters, such as dissolved oxygen,
chlorophyll concentrations, and submerged aquatic vegetation (SAV). Computed parameters are
compared with living resource standards, and an estimation is made of the degree to which computed
conditions benefit the resources of interest (e.g., fish, oysters). In addition, the CBEMP includes the
direct interactive simulation of SAV and water quality. Three phytoplankton groups were simulated.
Over seasonal time scales, the bay sediments are a significant source of dissolved nutrients to the
overlying water column. The role of sediments in the systemwide nutrient budget is especially important
in summer when seasonal low flows diminish riverine nutrient input. In addition, water temperatures
enhance biological processes in the sediments, creating greater sediment oxygen demand. Bay sediments
retain a long-term nutrient load "memory" of several years; that is, sediment nutrient fluxes to the water
column are determined by organic nutrient inputs from several previous years. Therefore, the water
quality model was coupled directly to a predictive benthic-sediment model. These two models interact at
each time-step, with the water quality model delivering settled organic material to the sediment bed, and
the benthic-sediment model calculating the flux of oxygen and nutrients to the water column.
Linked to the CBEMP are the watershed and airshed models, which provide daily input data. Generally,
10-year scenarios are run on 15-minute time-steps with output generated each 10 days. The estuary
Nutrient Criteria—Estuarine and Coastal Waters 9-31
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model has been in operation since 1987, with two major model refinements released since the initial 1987
steady-state model.
Further information on the entire suite of Chesapeake Bay Program models, their documentation, and
application can be found at: http://www.chesapeakebay.net/model.htm.
9-32 Nutrient Criteria—Estuarine and Coastal Waters
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CASE STUDY
SAN FRANCISCO BAY PROGRAM:
MANAGING COASTAL RESOURCES OF THE U.S.
The following case study is extracted from U.S. Geological Survey Fact Sheet FS-053-95
(available online at: http://water.usgs.gov/wid/html/sfb.html)
Coastal ecosystems, such as bays and estuaries, are among our most disturbed natural environments.
These ecosystems also are among our most valuable habitats—estuaries supported U.S. fisheries valued
at $19 billion in 1990. Although many human activities cause change in the coastal zone, they occur
against a background of natural change. Effective coastal-zone management requires that we identify
and understand these separate causes of ecosystem change. With this goal in mind, the United States
Geological Survey (USGS) began in 1968 a broad program of scientific study in San Francisco Bay
(Figure 1). The program is based on a conviction that sustained, multifaceted investigation of one
estuary will produce general lessons to guide the management of natural resources associated with all our
coasts.
The USGS San Francisco Bay Program has produced more than 250 reports, including three books and a
review of the human modifications of the bay. These publications are a source of guidance to resource
managers as they work to understand how human activities (such as water diversion, commercial trade,
and waste inputs) cause change in the coastal zone. The program has been organized around themes.
One of the most important themes is the integrated study of nutrients, toxic substances, and living
resources at lower levels of the food chain—the phytoplankton and bottom dwelling invertebrates. Close
collaboration between chemists and ecologists has helped to explain how plant and animal species of
coastal ecosystems are organized into food chains, how nutrients and toxic contaminants are incorporated
into these food chains, and how the lessons learned from detailed scientific understanding can be applied
to develop effective monitoring programs and rational environmental standards.
Nutrient Enrichment
Human settlement around coastal water bodies has led to increased inputs of nutrients such as nitrogen
and phosphorus. Many estuaries are now among the most intensively fertilized environments on Earth.
Each day, San Francisco Bay receives more than 800 million gallons of municipal wastewater containing
60 tons of nitrogen. In response to these concerns, the USGS developed a biological monitoring
procedure that has been used continuously since 1977 near a waste-treatment facility. Monitoring
continued as wastewater-treatment technologies improved. This is the longest continuous record of
contaminant concentrations in a natural environment of the United States. The transfer of monitoring
procedures developed by the USGS to local agencies and businesses serves as a model of cooperation
between research and regulatory agencies.
Management Questions
Water-quality managers need to know how nutrient inputs cause changes in water quality, the natural
capacity of coastal waters to assimilate added nutrients, the level of waste treatment required to protect
living resources from the harmful effects of nutrient enrichment, and if programs of nutrient reduction
are having beneficial effects.
USGS Contributions
Since 1968, the continuous study of San Francisco Bay by the USGS has given that agency a unique
opportunity to follow ecosystem responses to improved wastewater-treatment methods as mandated by
State and Federal legislation. One result of the implementation of these improved methods has been a
Nutrient Criteria—Estuarine and Coastal Water CS-1
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Da la-Kul I txlioi t ] < >curi 01 is
ituiw North and Soiuh
San Ji-ar»cjsco Bay
Figure 1. San Francisco Bay has been a focus of intensive investigation by the USGS since 1968.
large reduction in the input of ammonia-nitrogen from some municipal wastewater-treatment facilities
(Figure 2).
USGS studies show that in spite of its nutrient enrichment, San Francisco Bay has not been affected by
harmful algal blooms. This seeming paradox is explained partly by the abundant bottom-dwelling
invertebrates (small clams, mussels, crustaceans) that filter the water and remove new algae as fast as
they are produced. Feeding by these animals is a form of natural waste treatment that helps control the
growth of algae in a nutrient-rich environment.
Concepts and measurement techniques from this USGS program are now incorporated into a locally
funded and managed Regional Monitoring Program.
Lessons Learned
• The chemical quality of coastal waters can respond almost immediately to waste-treatment
improvements.
• Responses of biological communities to these chemical changes can take years or even decades.
• Coastal water bodies have differing sensitivities to waste loading. The most cost-effective national
strategy for regulating nutrient inputs will consider these differences among ecosystems.
For Further Information:
Visit the USGS website on San Francisco Bay at: http://sfbav.wr.usgs.gov/
CS-2
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SITE 1
< =
1975
1980 1985
YEAR
1990
1995
Figure 2. Implementation of advanced wastewater treatment in 1979 immediately reduced the input of
ammonia- nitrogen to South San Francisco Bay. In prior decades, the South Bay had repeated
episodes of oxygen depletion and animal die-offs. USGS measurements have shown a
complete cessation of these episodes since 1980. Spawning salmon have recently been
observed in South Bay streams for the first time since the early 1900's. See figure 1 for
location of site.
Nutrient Criteria—Estuarine and Coastal Water
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CASE STUDY
LONG ISLAND SOUND - HYPOXIA
U.S. Environmental Protection Agency
http://www.epa.gov/region01/eco/lis/hypox.html
The Problem
During the summers of 1987-93, from half to two-thirds of the Sound's bottom waters experienced
dissolved oxygen levels below 5 milligrams of oxygen per liter of water (mg/L). Levels of dissolved
oxygen of 5 mg/L and higher are generally accepted as being protective of the Sound's estuarine life. In
1989, a particularly bad summer, more than 500 square miles (40 percent) of the Sound's bottom waters
had dissolved oxygen levels less than 3 mg/L. During most of these years, dissolved oxygen in a portion
of the Sound (up to 50 square miles) fell below 1 mg/L and in 1987 anoxia, the absence of any oxygen,
was recorded in a portion of the Western Narrows.
These low levels of dissolved oxygen cause significant, adverse ecological effects in the bottom water
habitats of the Sound. To date, research shows that the most severe effects (such as mortality) occur
when dissolved oxygen levels fall below 1.5 mg/L at any time and below 3.5 mg/L in the short-term (i.e.,
4 days), but that there are probably mild effects of hypoxia when dissolved oxygen levels fall below 5
mg/L. The levels regularly observed in the Sound during late summer:
• Reduce the abundance and diversity of adult finfish;
• Reduce the growth rate of newly-settled lobsters and perhaps juvenile winter flounder;
• Can kill species that cannot move or move slowly, such as lobsters caught in pots and starfish, and
early life stages of species such as bay anchovy, menhaden, cunner, tautog, and sea robin;
• May reduce the resistance to disease of lobsters and other species; and
• Diminish the habitat value of Long Island Sound.
The Cause of the Problem
Excessive discharges of nitrogen, a nutrient, are the primary cause of hypoxia. Nitrogen fuels the growth
of planktonic algae. The algae die, settle to the bottom of the Sound and decay, using up oxygen in the
process.
Natural stratification of the Sound's waters occurs during the summer when warmer, fresher water
"floats" on the top of cooler, saltier water that is more dense. This natural stratification forms a density
difference between the two layers called a pycnocline. This prevents mixing of surface and bottom
waters.
Oxygen from the atmosphere and photosynthesis keep the surface layer well oxygenated, but the oxygen
cannot pass through the pycnocline into the bottom layer very easily. Decaying algae and other organic
material in the sediment and animal respiration in the bottom layer use up oxygen faster than it is
replenished. Hypoxia develops and usually persists as long as the stratification lasts (usually one to two
months in late summer).
But hypoxia in Long Island Sound is too complex to fully understand using direct observations alone.
Natural variations in weather and other physical factors affect the extent and severity of hypoxia. The
Management Conference has constructed mathematical models in order to understand the relationship
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among natural variations, human-caused pollutant loadings to the Sound, and dissolved oxygen levels in
the Sound. Two models, a water quality model that approximates the biological and chemical processes
of the Sound and a hydrodynamic model that describes physical processes, have been developed.
An intensive field program in Long Island Sound to collect data for the computer models was undertaken
from April 1988 to September 1989. These data were used to calibrate and verify the models to ensure
that they reproduce the important features of the Sound. The water quality model, called LIS 2.0,
provided needed insight into the causes of hypoxia and was the basis for actions to begin to reduce
nitrogen discharges to the Sound. However, because it simulates the movement of the Sound's waters in
only two dimensions (east-west and surface to bottom) and in a simplified manner, the LIS 2.0 model did
not provide the best technical foundation for identifying the total level of reduction in nitrogen loads that
should be attained or the most cost-effective means to achieve targeted reductions. The hydrodynamic
model, developed by the National Oceanic and Atmospheric Administration and completed in July, 1993,
uses tide and current measurements to simulate the water's circulation in three dimensions (east-west,
north-south, surface to bottom). It was coupled to the water quality model, to create LIS 3.0. The LIS 3.0
model provides an advanced tool to relate sources of nitrogen from specific geographic areas to the
hypoxia problem in the western Sound. Because the impact of the nitrogen load from different
management zones can be deter-mined using LIS 3.0, the LISS can assign priorities for management to
ensure that the most the cost-effective options are pursued.
The modeling, combined with field monitoring and laboratory studies, provided a level of detail to
support some clear conclusions about hypoxia in the Sound, its causes, and its solutions. In addition, the
models allowed the LISS to simulate water quality conditions as they were in the past, as they are today,
and as they could be in the future under alternative nitrogen control scenarios.
• The most oxygen that can be dissolved in Long Island Sound at summer water temperatures is
about 7.5 milligrams per liter (mg/L) of water. This is known as the saturation level.
• Oxygen concentrations greater than 5.0 mg/L provide healthy conditions for aquatic life.
Concentrations between 5.0 mg/L and 3.5 mg/L are generally healthy, except for the most sensitive
species. When concentrations fall below 3.5 mg/L, conditions become unhealthy. The most severe
effects occur if concentrations fall below 2.0 mg/L, even for short periods of time.
• The growth of algal blooms in Long Island Sound is dependent upon the availability of nutrients.
These blooms end when the pool of nitrogen available for continued growth is depleted.
• In pre-colonial days, natural, healthy biological activity brought oxygen levels below saturation due
to the natural loadings of organic material and nitrogen, but oxygen levels probably fell below 5
mg/L only in limited areas and for short periods of time.
• Under today's higher nutrient and organic material loading conditions, minimum oxygen levels
average approximately 1.5 mg/L. These levels are associated with severe hypoxia.
• By substantially reducing nitrogen loadings to the Sound, the minimum oxygen levels in the bottom
waters during late summer can be increased to an average of about 3.5 mg/L, thereby significantly
reducing the probability and frequency of severe hypoxia and reducing the area affected by
hypoxia.
• Increases in nitrogen delivered to the Sound could significantly worsen the hypoxia problem,
causing larger areas to have lower oxygen levels for longer periods of time. The probability of
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events like the summer of 1987, when anoxia (no oxygen) became a reality in the Sound, offshore
of Hempstead Harbor, would also increase.
Understanding the components of the load of nitrogen entering the Sound is fundamental to
understanding the plan:
• In 1990, defined as a baseline year by the Management Conference, the total nitrogen load was
90,800 tons per year.
• By 1992, the total nitrogen load had increased to 93,600 tons per year; this increase was anticipated
and was a consequence of terminating ocean disposal of sewage sludge from New York City and
the need to treat some of the sludge at facilities within the basin, reintroducing nitrogen to the
waste stream.
• Of the 93,600 tons per year, approximately 39,900 tons are from natural sources and not subject to
reductions by management activity.
• The remaining 53,700 tons of nitrogen per year are associated with human activities and have the
potential to be reduced through management actions.
• 10,700 tons of nitrogen per year enter the Sound through its boundaries ~ the East River in the west
and The Race in the east; efforts to reduce this substantial western load will come under the
auspices of the New York-New Jersey Harbor Estuary Program.
• 2,200 tons of nitrogen per year enter the Sound from direct atmospheric deposition; the
Management Conference estimates that this load will be reduced to 1,540 tons of nitrogen per year
through implementation of the 1990 Clean Air Act amendments.
• The remaining 40,800 tons of nitrogen per year are a result of human activity coming from point
and nonpoint source discharges in the Sound's drainage basin and are the subject of the plan. Point
source discharges, primarily sewage treatment plants, result in 32,400 tons of nitrogen each year
and nonpoint source discharges, such as agricultural and stormwater runoff, result in 8,400 tons of
nitrogen each year.
The Plan to Solve the Problem
The goal of the hypoxia management plan is to eliminate adverse impacts of hypoxia resulting from
human activities. Achievement of this goal will require very large investments of capital, a long-term
commitment, and the assistance of the New York-New Jersey Harbor Estuary Program. Therefore, the
Management Conference has established interim targets for dissolved oxygen and has outlined a phased
approach to achieving them, using what is known now to support early phases and committing to take
additional steps as increased understanding of the environment will dictate in the future.
Interim Dissolved Oxygen Targets
Using scientific information on the relationship between oxygen levels and ecological effects, the
Management Conference has established interim target levels for oxygen that, if achieved, would
minimize the adverse impacts of hypoxia. In summary, the interim dissolved oxygen targets for the
bottom waters of the Sound are to:
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• Maintain existing dissolved oxygen levels in waters that currently meet State standards;
• Increase dissolved oxygen levels to meet standards in those areas below the State standards but
above 3.5 mg/L; and,
• Increase short-term average dissolved oxygen levels to 3.5 mg/L in those areas currently below 3.5
mg/L, ensuring that dissolved oxygen never goes below 1.5 mg/L at any time.
• There are also interim targets for the surface waters of the Sound.
Ph used Approach
The Management Conference is implementing a phased approach to reducing nitrogen loadings to the
Sound from point and nonpoint source discharges within the Sound's drainage basin.
Phase I, as announced in December of 1990, froze nitrogen loadings to the Sound in critical areas at
1990 levels to prevent hypoxia from worsening.
Phase II, as detailed in the plan, includes significant, low-cost nitrogen reductions that begin the
process of reducing the severity and extent of hypoxia in the Sound.
Phase III will present nitrogen reduction targets to meet the interim targets for dissolved oxygen,
which will prevent most known lethal and sublethal effects of hypoxia on the Sound's estuarine life.
Phase III also will lay out the approach for meeting the nitrogen load reduction targets.
Phase I- The Nitrogen Loading Freeze
Phase I was announced in December 1990. It called for a freeze on point and nonpoint nitrogen loadings
to the Sound in critical areas at 1990 levels. It committed the States and local governments to specific
actions to stop a 300-year trend of ever increasing amounts of nitrogen entering the Sound.
Since 1990, activities have been underway in New York and Connecticut to manage nitrogen from
sources within the New York and Connecticut portions of the drainage basin, starting with adoption of
the Phase I "freeze" on loadings.
Connecticut reacted quickly to obtain $15 million in State funds to ensure that the nitrogen freeze
was implemented. Consent orders are in place to cap the nitrogen loads at the 15 affected facilities.
In New York City, the New York State Department of Environmental Conservation (NYSDEC) and
the city have reached full agreement on sewage treatment permit limits, freezing total nitrogen
loadings at 1990 levels.
In Westchester County, the NYSDEC has issued final permits to the four existing sewage treatment
plants, freezing their aggregate load at the 1990 level. This was done with the full agreement of the
county.
On Long Island, the NYSDEC proposed individual permits that freeze the loads from individual
discharges at 1990 levels; in response, the dischargers proposed establishment of an aggregate
limit. State and local authorities agreed on aggregate load limits for targeted facilities.
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Phase I agreements to control nonpoint sources centered around three categories:
• Use of existing nonpoint source and stormwater management programs to focus on nitrogen control
with the objective of freezing the loads.
• Assessing tributary loads to Long Island Sound to begin planning for their control.
• Assigning priorities for management to coastal subbasins where nitrogen loads were estimated to be
the highest.
Phase II - Low Cost Nitrogen Reductions
For Phase II, the LISS made a commitment in 1994 to reduce nitrogen discharges to the Sound from peak
loadings by approximately 7,550 tons per year. This phase consists of incorporating a variety of low-cost
nitrogen removal technologies at selected sewage treatment plants. The States have moved aggressively
to implement nitrogen control activities, using innovative strategies and seeking the cooperation of local
governments.
In Connecticut, the goal was to achieve a reduction of 850 tons per year in nitrogen loads. The State of
Connecticut has awarded more than $15 million through its State Clean Water Fund to 11 southwestern
sewage treatment plants to test and demonstrate the efficiency of upgrades for nitrogen treatment. In
addition, the first plant in the State designed to denitrify has been constructed in Seymour. As of
December 1997, the load of nitrogen from plants in the Phase II agreement has been reduced by almost
900 tons per year, exceeding the Phase II goal.
The State of New York revised the permits issued to sewage treatment plants, with the consent of local
authorities, to establish nitrogen limits at 1990 levels. The permits include an aggregate load for
facilities within Management Zones 7-11 (New York City, Westchester County, and Long Island). The
New York goal was to reduce nitrogen loadings by 6,700 tons per year from peak loadings from actions
to be completed by 2006. The goal of these actions was to compensate for the increased load due to
sludge treatment and reduce loadings back below 1990 levels. As of 1997, one sewage treatment plant in
Westchester County and four in New York City have implemented nitrogen removal technologies. New
York City is required to implement additional nitrogen removal technologies at the upper East River
sewage treatment plants. As of December 1997, the load of nitrogen from sewage treatment plants in
New York had decreased by 3,000 tons per year from peak loadings. In addition, New York City has
entered into a consent order to provide nitrogen removal at the reconstructed Newtown Creek facility,
scheduled for completion in 2007.
In addition, both States have:
Developed materials and conducted training for treatment plant personnel on nitrogen removal
technologies and procedures.
Required sewage treatment plants to identify in their plans how they will remove nitrogen, if
required to do so.
Required nutrient monitoring at sewage treatment plants to improve understanding of nitrogen
sources and treatment plant capability.
Increased the share of nonpoint source pollution control funds targeted to projects that reduce
nitrogen loads to the Sound.
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• Formulated Coastal Nonpoint Pollution Control Programs to address coastal nonpoint sources of
nitrogen.
• Undertaken demonstration projects that address a variety of nonpoint source control issues and
technologies (e.g., urban runoff treatment by artificial pond/wetland systems, parking lot runoff
treatment, septic system technologies to treat and remove nitrogen, controlling runoff from
agricultural land and from marinas).
As of December 31, 1997, nitrogen loadings to the Sound from point and non-point sources within the
New York and Connecticut portions of the watershed have been reduced as a result of these activities by
3,900 tons per year from peak loadings.
Phase III - Nitrogen Reduction Targets to Eliminate Severe Hypoxia
While steps taken in Phases I and II will help to reduce the extent of hypoxia, additional nitrogen
reduction is needed to restore the health of Long Island Sound. Phase III sets the course by setting
specific nitrogen reduction targets for each of the 11 management zones around the
Sound. An array of environmental and economic considerations were taken into account throughout the
process.
Oxygen Benchmarks
The water quality standard for oxygen in Long Island Sound is 6 mg/L in Connecticut and 5 mg/L in New
York. Modeling indicates that even if maximum nitrogen reduction technologies were implemented, the
water quality standards for oxygen would not be achieved throughout the summer in all areas of the
Sound. To help establish priorities for action, the LISS has identified oxygen conditions that will
minimize adverse impacts on living resources of the Sound.
Two major research efforts, a laboratory study by the EPA's Office of Research and Development and a
field study by the Connecticut Department of Environmental Protection (CTDEP) have provided much of
the information on how low oxygen conditions affect living resources in the Sound. Both studies
corroborated that severe effects occurred whenever levels of oxygen fell below 2.0 mg/L. The field
surveys noted large reductions in the number and types of aquatic life present. The lab experiments
recorded reductions in growth and increases in mortality. In both studies, effects became significant
when oxygen levels fell below 3.5 mg/L, though some effects occurred at levels between 3.5-5.0 mg/L.
As a result, the LISS has determined that unhealthy conditions occur whenever oxygen levels fall below
2.0 mg/L at any-time or remain below 3.5 mg/L over a 24-hour period. Most adverse impacts can be
prevented if oxygen levels exceed these conditions, and they have been used as benchmarks to assess the
relative benefits of alternative management strategies for improving the health of Long Island Sound.
Cost-effectiven ess
LISS managers looked at a range of nitrogen reduction options for the three major sources of nitrogen in
the watershed, sewage treatment plants, industrial facilities, and nonpoint source runoff to determine the
most cost-effective option.
• Sewage Treatment Plants: As nitrogen removal requirements become more stringent, the cost of
controls tends to increase. To identify a cost-effective level of treatment, LISS managers arrayed
the possible nitrogen reduction options for all 70 sewage treatment plants in the 11 management
zones and calculated the average oxygen improvement in the Sound per dollar spent.
Improvements at sewage treatment plants that had better than average cost-effectiveness at
improving oxygen conditions in the Sound were identified. These actions, in total, could achieve a
62 percent reduction in loads, or 122,044 pounds/day.
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• Industrial Facilities: A limited number of industrial facilities directly contribute nitrogen to the
Sound; all are located in Connecticut and contribute an estimated 6,717 pounds per day of nitrogen
to the Sound. Because information on the cost of reducing nitrogen from industrial sources was not
readily available, these facilities were not included in the cost analyses used for sewage treatment
plants. Instead, the cost-effective level of treatment identified for sewage treatment plants, 62
percent, was applied to the industrial sources, resulting in a 4,165 pounds per day reduction for
industrial facilities. This represents an aggressive but cost-effective level of nitrogen control for
these sources.
• Nonpoint Sources: Decisions on controls of nonpoint source runoff must be made in the broader
context of watershed management, since control measures will also help reduce suspended solids,
toxic contaminants, pathogens, and floatable debris. The LISS recommends that aggressive controls
of nonpoint source pollution be implemented for both existing and new development, through both
habitat protection and restoration activities, and structural and nonstructural best management
practices. This effort could result in a 10 percent reduction in the non-point source load from
sources within the New York and Connecticut portions of the watershed, or 2,604 pounds per day.
Adding the potential nitrogen reductions from cost-effective controls on sewage treatment plants,
industrial sources, and nonpoint runoff sources results in a total reduction of 128,813 pounds per day
(23,500 tons per year). The next step is to allocate responsibility for achieving these reductions among
the 11 management zones fairly.
Allocating Responsibility
The cost curve analysis provided an option for allocating nitrogen reductions among the sewage
treatment plants. Sewage treatment plant upgrades with greater than average cost-effectiveness would be
implemented while upgrades with below average cost-effectiveness would not be implemented.
However, the LISS decided that relying on the cost curve analysis alone would not be a fair or even
feasible approach and would not provide the best solution to allocating nitrogen reduction.
There are several reasons for this conclusion. Most importantly, the cost estimates were general and not
uniform in their development. More accurate cost estimates must await detailed facilities planning based
upon a clear definition of the nitrogen discharge limits that will have to be met. In addition, local
concerns and considerations such as the need to purchase land for expansion and to distinguish between
costs for nitrogen removal versus ongoing maintenance, expansions for growth, and secondary upgrade
needs (which were not included in the cost estimates) were not addressed evenly in the cost analysis.
Cost considerations aside, it is necessary for all sewage treatment plants to share the burden of nitrogen
removal. All sewage treatment plants contribute nitrogen to Long Island Sound, albeit with different
effect. All jurisdictions will benefit from improved water quality. Therefore, it is reasonable to expect
all contributors to the problem to con-tribute to the solution.
For those reasons, LISS has assigned each management zone equal responsibility to reduce its share of
the nitrogen load. To achieve a similar level of oxygen improvement from reductions allocated to each
zone by the same percentage, the load reduction target was adjusted slightly to 23,800 tons per year from
the original 23,500 tons per year. The total human-derived load coming from sewage treatment plants,
industrial point sources, and nonpoint sources, including atmospheric depositions within the water-shed,
is 40,650 tons per year. Therefore, the Soundwide nitrogen target is a 58.5 percent reduction in the
human-derived load from point and nonpoint sources in the watershed.
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Phase III Actions
Phase III actions will minimize adverse impacts of hypoxia caused by human activities in
a cost-effective manner, while ensuring that new information is gathered to refine and improve
management over the long term. Using the framework described above, the LISS set a 58.5 percent
reduction target for the enriched load of nitrogen from sources within the New York and Connecticut
portions of the watershed.
Strategies
Attaining the nitrogen reduction targets will require aggressive control of point sources, such as sewage
treatment plants and industrial sources, and nonpoint sources, such as on-site sewage systems and runoff
from roads, parking lots, and construction sites. To achieve the reduction targets, the States, working
with local governments, will select the mix of point and nonpoint source controls to be implemented in
each management zone. Recognizing that each watershed is different, the plan provides the States and
municipalities considerable flexibility in determining how nitrogen reduction actions are carried out
within each zone.
By August 2000, the States will take the following actions:
Develop watershed plans for each management zone that will set the course for achieving the
targets as scheduled.
Consistent with those plans, incorporate limits on the amount of nitrogen that can be discharged
from sewage treatment plants and industrial sources into discharge permits.
Conduct comprehensive nonpoint source management and habitat restoration activities.
Because the total nitrogen load entering the Sound from human sources is dominated by point source
discharges, the plan emphasizes technologies that can be applied to sewage treatment facilities and
industrial discharges.
In order to achieve significant reductions in the nonpoint source nitrogen load, home owners, farmers,
businesses, municipalities, and the States will need to reduce current inputs of nitrogen to the watershed
and restore and preserve the nitrogen removal capabilities of existing natural systems. These reductions
can be achieved using a number of approaches—resource-based land use decisions at the local level,
watershed-wide use of appropriate structural and nonstructural best management practices (e.g.,
stormwater detention ponds, artificial wetlands, streetsweeping, cleaning catch basins), habitat protection
and restoration, and pollution prevention management practices. All approaches will require a concerted
education and outreach effort.
Timing
The planning, financing, and construction of upgrades to sewage treatment plants necessary to achieve
the 58.5 percent reduction target will require sustained effort and commitment over a long period of time.
Therefore, the LISS recommends phasing-in the necessary reductions over 15 years:
40 percent in 5 years,
75 percent in 10 years, and
• 100 percent in 15 years.
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Cost
The Comprehensive Conservation and Management Plan identified that the cost of achieving maximum
nitrogen removal from all point sources would range from $6 to $8 billion ($5.1 to $6.4 billion in New
York State and from $900 million to $1.7 billion in Connecticut). Because of the successful
demonstration of full scale nitrogen removal technologies at sewage treatment plants undertaken as part
of Phase II, the estimated costs of capital improvements at sewage treatment plants have decreased. The
estimated cost of achieving maximum nitrogen removal levels at the 70 treatment plants in New York
and Connecticut is now about $2.5 billion
Because of the cost-effective approach described above, the LISS nitrogen reduction strategy would not
require all treatment plants to meet limit-of-technology reductions. As a result, the incremental capital
cost of achieving the Phase III point source controls was estimated to be $300 million for New York State
and $350 million for Connecticut. These cost estimates have been questioned and will be revised as
more detailed facility planning and design is performed. However, they show clearly that the potential
cost of achieving our goals can be much less than originally estimated.
Nonpoint source controls will be implemented as part of broader watershed and habitat protection efforts.
The cost of controlling nonpoint sources is more difficult to estimate than the cost of point source
controls. Rather than one type of technology applied to a similar source, a variety of strategies can be
applied to control a variety of nonpoint sources of nitrogen. As a result, the costs of achieving nonpoint
nitrogen reductions will be addressed in the zone-by-zone plans developed by the States.
Financing
As recommended in the Comprehensive Conservation and Management Plan, the main source of funding
for these wastewater treatment facility improvements will be the State Revolving Fund programs. The
EPA, through the federal Clean Water Act, provides financing to support State Revolving Fund loan
programs.
Connecticut uses the capitalization grant from EPA to leverage with State bond funds to provide grants
and low interest loans, at 2 percent interest over 20 years, to finance improvements at municipal
facilities. Connecticut provides about $50 million per year in State bonding to supplement the $15
million per year provided under the Clean Water Act. At this capitalization rate, Connecticut should be
able to meet municipal financing needs to implement Phase III nitrogen reductions. During fiscal year
1997, CTDEP awarded $250 million from their Clean Water Fund to finance projects of benefit to Long
Island Sound, including major sewage treatment plant upgrades in Norwalk and Waterbury.
New York State established its State Revolving Fund in the custody of the Environmental Facilities
Corporation. This public corporation benefits local governments in New York State by offering below-
market interest rate loans to municipalities to finance wastewater improvements. Currently, the interest
rate is set at up to one-half of the market rate to be repaid in 20 years. Lower rates of interest, including
zero interest loans, are available for communities that can demonstrate an inability to pay the standard
subsidized rate. Another major source of funding in New York State is the $1.75 billion Clean
Water/Clean Air Bond Act approved by voters in November 1996. The Bond Act targeted $200 million
for Long Island Sound that will be available for sewage treatment upgrades, habitat restoration, nonpoint
source control, and pollution prevention.
The possible funding sources for non-point source controls reflect the diversity of both the sources and
the control options. Grant funding through federal and State water quality management, natural
resources management, and coastal zone management programs is available for nonpoint source
activities. The State Revolving Fund loan program is also available to fund storm water management and
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habitat restoration projects but has not been used to a great extent for these types of activities due to the
magnitude of existing point source funding needs in Connecticut and New York.
Effluent Trading
To provide further flexibility and incentives for maximizing the timeliness and cost-effectiveness of
nitrogen reduction actions, the LISS is investigating the feasibility of allowing effluent trading. Trading,
if employed as part of the nitrogen reduction effort, may be an innovative way to use market forces to
more efficiently meet water quality goals. The LISS is developing a trading proposal and will convene a
public forum for federal, State, and local water quality officials, together with public and private
interests, to evaluate its potential.
Enforcement
The provisions of the federal Clean Water Act provide a vehicle for ensuring that nitrogen reduction
targets are legally enforceable. Section 303(d) of the Act requires the identification of a Total Maximum
Daily Load for pollutants that will result in the attainment of water quality standards. Once a Total
Maximum Daily Load has been established, the act calls for reductions to be allocated to sources so that
the load target is met. New York and Connecticut and EPA will use their authorities to provide an
enforceable foundation for achieving the nitrogen reduction targets. By August, 1998 the States will
propose a Total Maximum Daily Load designed to meet State oxygen standards. The current Long Island
Sound standards were developed with limited data on how low oxygen levels affect aquatic life in Long
Island Sound. EPA is currently developing regional marine oxygen criteria that will provide a more
scientifically valid basis for the development of oxygen standards. Based on this information, the States
may, in the future, modify their oxygen standards. While LISS managers predict significant
improvement in water quality as the nitrogen reduction targets are implemented, the attainment of current
water quality standards at all times and in all areas is not expected. For this reason, the LISS will
continue to assess what other kinds of actions will be needed to bring the Sound into full compliance
with water quality standards.
These actions may include control of nitrogen and carbon sources outside of the Long Island Sound basin
(e.g., tributary import from point and non-point sources north of Connecticut, atmospheric deposition,
boundary import from point and nonpoint sources affecting New York Harbor and The Race).
Alternatives to nitrogen reduction, such as aeration, will need to be considered as a possible means to
achieve water quality standards in remaining areas.
Evaluating Progress
The LISS will track, monitor, and report on progress in meeting the nitrogen reduction targets annually.
In addition, a formal review of the goals and objectives of the program will be performed every 5 years,
coinciding with the progress checkpoints for nitrogen reduction. The review will consider:
• Progress and cost of implementation, including a reevaluation of the knee-of-the-curve analysis
used to establish the Phase III nitrogen reduction targets,
• Improvements in technology, including the results of quality controlled pilot projects,
• The regional dissolved oxygen criteria to be published for comment,
• Water quality standards,
• Refined information on the ecosystem response to nitrogen reductions,
Nutrient Criteria—Estuarine and Coastal Water CS-13
-------
• The results of peer reviewed modeling, and
• Research on the impacts of hypoxia to living resources and their habitats.
Each of these factors will be considered in a balanced manner in the reevaluation process. As a result of
the review, the LISS may recommend improvements that could result in changes in how the overall
program will be implemented.
For More Information:
Mark Tedesco
EPA Long Island Sound Office
888 Washington Blvd.
Stamford, CT 06904-2152
Phone: 203/977-1541
Fax: 203/977-1546
CS-14 Nutrient Criteria—Estuarine and Coastal Water
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CASE STUDY
NP BUDGET FOR NARRAGANSETT BAY
S. V. Smith
University of Hawaii
808-956-8693
http://data.ecology.su.se/MNODE/North%20America/NRB.HTM
Narragansett Bay, Rhode Island (41° 35' N, 71° 20'W), is a relatively well mixed, near-oceanic salinity
estuary on the northeast (Atlantic) coast of the U. S. It occupies an area of 264 km2 (Table 1) and has a
mean depth of 9.7 m. Note that both the area and volume differ from the comments in Nixon et al.
(1995), but seem consistent with Kremer and Nixons (1978) explicit tabulation. Freshwater flow into the
system averages about 8.2 x 106m3/d, from a watershed of 3500 km2. Primary production in the system is
dominated by phytoplankton (29 mol C m"2 yr"1) with a C:N:P ratio of about 112:13:1. The budget
described below is based on data collected primarily in the late 1970's and through much of the 1980's.
Details of this kind of analysis can be found at the LOICZ - Biogeochemical Modelling web site at:
http://data.ecologv.su.se/MNODE/index.htm.
Sector area and volume data are from Kremer and Nixon (1978). Sector nutrient concentrations are
annual averages (based on surface and deep water data) also from Kremer and Nixon. Sector nutrient
masses are calculated as volume x concentration. The sectors at the bay mouth (#5, 8) are used for
"oceanic values."
Nutrient exchange fluxes (Table 2) are calculated using an average 26-day exchange time, as calculated
by Pilson (1985) with a water and salt budget (analogous to procedure in Gordon et al., 1996). The bay
Table 1. Sector areas, volumes, and nutrient concentrations. Data are used to calculate volume-
averaged concentrations for the outer portion of the bay ("ocean") and the bay proper
SECT.
#
1
2
3
4
5
6
7
8
SUM
VOL.106
m3
130
300
115
463
204
222
573
554
2561
AREA
106m2
20.1
44.6
28.5
61.9
20.0
26.0
38.5
24.2
264
bay
ocean
(sees.
#5,8)
DIP
HM
1.8
1.5
1.6
1.4
1.0
1.2
1.0
0.7
1.3
0.8
NH4
HM
12
7
3
4
1
2
2
1
4
1
N03
HM
11
6
6
4
3
5
4
2
5
2
Sum DIN
HM
23
13
9
8
4
7
6
3
9
3
DIP
106 mol
0.23
0.45
0.18
0.65
0.20
0.27
0.57
0.39
NH4
106 mol
1.56
2.10
0.35
1.85
0.27
0.44
1.15
0.39
N03
106 mol
1.43
1.80
0.70
1.85
0.60
1.11
2.29
1.11
SDIN
106 mol
2.99
3.90
1.05
3.70
0.87
1.55
3.44
1.50
Nutrient Criteria—Estuarine and Coastal Water
CS-15
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Table 2. Hydrographic exchange fluxes of nutrients
SOURCE OF
FLUX
residual flow
net exchange flow
total hydrography
DIP
106mol/yr
-3
-18
-21
NH4
106mol/yr
-7
-108
-115
N03
106mol/yr
-11
-108
-119
DIN
106mol/yr
-18
-216
-234
volume divided by the residence time gives a mixing exchange volume of 98.5 x 106 m3/d, while the
residual outflow equals the freshwater inflow (i.e., 8.2 x 106m3/d). It would, in principle, be possible to
time-step through the data (at monthly increments, for example). However, to do that would require
having the flow data to go with the nutrient data. Further, inspection of the graph by Nixon et al. (1995)
of flow data and comparison of Pilson's (1985) flow—residence time regression equation suggests that
the residence time over this range of flow is well approximated by a constant value for the exchange
time. Various authors describe the bay as well mixed, and this is supported by the water composition data
in Kremer and Nixon (1978). We therefore use a 1-box model to perform these calculations, rather than
a vertically stratified model to describe hydrographic fluxes.
In Table 3, all boundary fluxes except hydrography were taken directly from by Nixon et al. (1985).
Hydrographic flux was calculated as above. A Y's (the nonconservative fluxes) are calculated by
difference (as described in Gordon et al., 1996). No data are available for DOP, DON, or for either
inorganic or organic C, so the budget is based on inorganic N and P only. As discussed by Gordon et al.
and consistent with comments in Nixon et al., it seems safe to assume that DOP and DON
nonconservative fluxes do not contribute strongly to the overall nonconservative fluxes in this system.
Rates for the D Y's per unit area are calculated using the bay area of 264 km2 (Table 4). Note that this
area estimate is about 25% lower than the value used by Kremer and Nixon (1978). We have used the
smaller area and volume on the basis that these are the data used to calculate the volume-averaged
concentrations. Net (nfix-denit) is calculated on the assumption that the N:P ratio of D DIP is 13:1, then
D DIN is balanced. Net (p-r) is calculated from the DIP flux, using a C:P ratio of 112:1.
Table 3. Total boundary fluxes of nutrients and inferred internal reactions—
the system budget
Process
atmosphere
rivers
urban runoff
sewage
hydrography
DY
(nfix-denit)
DIP
106mol/yr
0
13
2
9
-21
-3
NH4
106mol/yr
6
113
13
136
-115
-153
N03
106mol/yr
19
177
4
6
-119
-87
DIN
106mol/yr
25
290
17
142
-234
-240
-201
CS-16
Nutrient Criteria—Estuarine and Coastal Water
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Table 4. Nonconservative fluxes of materials and stoichiometrically inferred
biogeochemical pathways
DY
DDINexn
(nfix-denit)
(p-r)
DIP
mmol m 2 yr J
-11
NH4
mmol m 2 yr1
-580
NO3
mmol m 2 yr J
-329
DIN
mmol m 2 yr J
-909
-143
-766
C
mol m 2 yr1
1.2
Nixon et al. (1995) have data with which the present budgetary estimates may be compared: They
estimate DIP and DIN fluxes from the ocean to the bay by a hydrographic budget analogous to values
used here for both influx and efflux, but they do not use this same hydrography to estimate nutrient
fluxes to the ocean. Their inward DIP and DIN fluxes, obtained by time-stepping through the oceanic
nutrient concentration data (bottom water only), are 27 and 115 x 106 mol/yr. The calculations here
(using annual average data) are 29 and 108 x 106 mol/yr. The agreement is within 10%. It should be
close, because both Nixon et al. and the calculations here are performing essentially the same calculation.
Three points for minor disagreement would be that the values here just used a constant exchange rate
(instead of time-varying); values used here were picked data off a graph; and surface and bottom values
were averaged (on the graph, these are effectively identical in the outermost bay sectors).
Instead of using hydrography to estimate outward DIP and DIN flux, those authors estimate DIP and DIN
fluxes from the bay to ocean by difference with other terms in their budget, to close the budget. They get
41-51 x 106 and 240-470 x 106 mol/yr. Again pulling the hydrographic terms apart, the calculations here
yield 50 x 106 and 342 x 106 mol/yr (in both cases, within their range). It is worth noting that if the water
exchange volume is incorrect, it would affect both influx and efflux of nutrients, hence have a relatively
small effect on the difference between influx and efflux. The point here, of course, is that the difference
between influx and efflux is probably more reliable than either of the individual fluxes.
Nixon et al. use a variety of considerations for two different sets of incubation data to assign baywide
denitrification a range of 85-170 x 106 mol/yr (compared to 201 x 106 from the hydrographic budget;
using their high values, agreement is within 20%).
Those authors estimate respiration to consume 8100 to 9200 x 106 mol/yr of organic C. Using their
estimate for primary production (p) of 29 mol C m"2 yr"1 and the DIP-derived estimate for production -
respiration (p-r) of 1.2, r is estimated to be 27.8 mol C m"2 yr"1. Scaling by the bay area, this gives
respiration to be 7340 x 106 mol/yr (within 20% of their lower estimate). If we were to use the are value
given in Nixon et al. (328 km2, instead of the value of 264 km2 from Kremer and Nixon (1978), the
respiration would be 9118 x 106 mol/yr (within their range).
Efforts to control the release of nutrients into Narragansett Bay have recently addressed nitrogen
contributions from Publically Owned Treatment Works (POTWs) throughout the watershed. One
nutrient reduction option currently being pursued is to maximize nitrogen removal from the final effluent
by modifying operating conditions with existing equipment at the facility. Retrofitting existing facilities
will also be considered where appropriate. A second venue involves drafting water quality based permit
limits over the next few years to limit nitrogen in the final effluent of POTWs. Finally, a total maximum
daily load (TMDL) for nitrogen is currently under development for the Providence River upstream of
Narragansett Bay through the NPDES permitting process. A model is being developed that once
calibrated, will set nitrogen load limits for POTWs that discharge to the river.
Nutrient Criteria—Estuarine and Coastal Water
CS-17
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For Further Information:
S. V. Smith
Department of Oceanography,
University of Hawaii
1000 Pope Road Honolulu, Hawaii 96822 USA
email: svsmith@soest.hawaii.edu
phone: 808-956-8693
fax: 808-956-7112
CS-18 Nutrient Criteria—Estuarine and Coastal Water
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CASE STUDY
NUTRIENT MANAGEMENT AND SEAGRASS RESTORATION IN TAMPA BAY, FLORIDA
Holly Greening, Tampa Bay Estuary Program
Abstract: Participants in the Tampa Bay Estuary Program have agreed to adopt nitrogen loading
targets for Tampa Bay based on the water quality and related light requirements of the seagrass
species Thalassia testudinum. Based on modeling results, it appears that light levels can be
maintained at necessary levels by "holding the line" at existing nitrogen loadings. However, this
goal may be difficult to achieve given the 20% increase in the watershed's human population and
associated 7% increase in nitrogen loading that are projected to occur over the next 10-20 years.
To address the long-term management of nitrogen sources, a Nitrogen Management Consortium of
local electric utilities, industries and agricultural interests, as well as local governments and
regulatory agency representatives, has developed a Consortium Action Plan to address the target
load reduction needed to "hold the line" at 1992-1994 levels. To date, implemented and planned
projects collated in the Consortium Action Plan meet and exceed the agreed-upon nitrogen loading
reduction goal.
The Tampa Bay estuary is located on the eastern shore of the Gulf of Mexico in Florida, USA. At more
than 1000 km2, it is Florida's largest open water estuary. More than 2 million people live in the 5700
km2 watershed, with a 20% increase in population projected by 2010. Land use in the watershed is
mixed, with about 40% of the watershed undeveloped, 35% agricultural, 16% residential, and the
remaining commercial and mining.
Major habitats in the Tampa Bay estuary include mangroves, salt marshes and submerged aquatic
vegetation. Each of these habitats has experienced significant areal reductions since the 1950s, due to
physical disturbance (dredge and fill operations) and water quality degradation, particularly impacting
the seagrasses due to loss of light availability. Five species of seagrass are commonly found in Tampa
Bay, with Thalassia testudinum (turtlegrass) and Syringodium filiforme (manatee grass) dominating in
the higher salinity areas and Halodule wrightii (shoalgrass) and Ruppia maritima (widgeon grass) most
commonly found in lower salinities.
The importance of seagrass as a critical habitat and nursery area for fish and invertebrates, and as a food
resource for manatees, sea turtles and other estuarine organisms has been recognized by the Tampa Bay
resource management community for several decades. In 1990, Tampa Bay was accepted into the U.S.
Environmental Protection Agency's (EPA) National Estuary Program. The Tampa Bay National Estuary
Program (TBNEP), a partnership that includes three regulatory agencies and six local governments, has
built on the resource-based approach initiated by earlier bay management efforts. Further, it has
developed water quality models to quantify linkages between nitrogen loadings and bay water quality,
and models that link water quality to seagrass goals.
Recent recommendations from the National Academy of Science National Research Council (NRC)
include those which regional watershed programs might consider in developing nutrient management
strategies. The NRC recommendations are based on the process designed by the Tampa Bay Estuary
Program partners to develop and implement a seagrass protection and restoration management program
for Tampa Bay. Critical elements of the Tampa Bay process are to:
Nutrient Criteria—Estuarine and Coastal Water CS-19
-------
1. Set specific, quantitative seagrass coverage goals for each bay segment.
2. Determine seagrass water quality requirements and appropriate nitrogen loading targets.
3. Define and implement nitrogen management strategies needed to achieve load management targets.
STEP 1. SET QUANTITATIVE RESOURCE MANAGEMENT GOALS
Establishment of clearly defined and measurable goals is crucial for a successful resource management
effort. In 1992, TBNEP adopted an initial goal to increase current Tampa Bay seagrass cover to 95% of
that present in 1950.
Based on digitized aerial photographic images, it was estimated that approximately 16,500 ha of seagrass
existed in Tampa Bay in 1950. At that time, seagrasses grew to depths of 1.5 m to 2 m in most areas of
the bay. By 1992, approximately 10,400 ha of seagrass remained in Tampa Bay, a loss of more than 35%
since the 1950 benchmark period. Some (about 160 ha) of the observed loss occurred as the result of
direct habitat destruction associated with the construction of navigation channels and other dredging and
filling projects within existing seagrass meadows, and is assumed to be nonrestorable through water
quality management actions.
In 1996, the TBNEP adopted a bay-wide minimum seagrass goal of 15,400 ha. This goal represented
95% of the estimated 1950 seagrass cover (minus the nonrestorable areas), and includes the protection of
the existing 10,400 ha plus the restoration of an additional 5,000 ha.
STEP 2. DETERMINE SEAGRASS WATER QUALITY REQUIREMENTS AND APPROPRIATE
NITROGEN LOADING RATES
Once seagrass restoration and protection goals were established by the participants, the next steps
established the environmental requirements necessary to meet agreed-upon goals and subsequent
management actions necessary to meet those requirements.
A. Determine environmental requirements needed to meet the seagrass restoration goal
Recent research indicates that the deep edges ofThalassia testudinum meadows, the primary seagrass
species for which nitrogen loading targets are being set, correspond to the depth at which 20.5% of
subsurface irradiance (the light that penetrates the water surface) reaches the bay bottom on an annual
average basis. The long-term seagrass coverage goal can thus be restated as a water clarity and light
penetration target. Therefore, in order to restore seagrass to near 1950 levels in a given bay segment,
water clarity in that segment should be restored to the point that allows 20.5% of subsurface irradiance to
reach the same depths that were reached in 1950.
B. Determine water clarity necessary to allow adequate light to penetrate to the 1950 seagrass deep edges
Water clarity and light penetration in Tampa Bay are affected by a number of factors, such as
phytoplankton biomass, non-phytoplankton turbidity, and water color. Water color may be an important
cause of light attenuation in some bay segments; however, including color in the regression model did not
produce a significant improvement in the predictive ability of the regression model. Results of the
modeling effort indicate that, on a baywide basis, variation in chlorophyll a concentration is the major
factor affecting variation in average annual water clarity.
C. Determine chlorophyll a concentration targets necessary to maintain water clarity needed to meet the
seagrass light requirement
An empirical regression model was used to estimate chlorophyll a concentrations necessary to maintain
water clarity needed for seagrass growth for each major bay segment. The adopted segment-specific
CS-20 Nutrient Criteria—Estuarine and Coastal Water
-------
annual average chlorophyll a targets (ranging from 4.6 |ig/l to 13.2 |ig/l) are easily measured and tracked
through time, and are used as intermediate measures for assessing success in maintaining water quality
requirements necessary to meet the long-term seagrass goal.
D. Determine nutrient loadings necessary to achieve and maintain the chlorophyll a targets
Water quality conditions in 1992-1994 appear to allow an annual average of more than 20.5% of
subsurface irradiance to reach target depths (i.e., the depths to which seagrasses grew in 1950) in three of
the four largest bay segments. Thus, a management strategy based on "holding the line" at 1992-1994
nitrogen loading rates should be adequate to achieve the seagrass restoration goals in these segments.
This "hold the line" approach, combined with careful monitoring of water quality and seagrass extent,
was adopted by the TBNEP partnership in 1996 as its initial nitrogen load management strategy.
As an additional complicating factor, a successful adherence to the "hold the line" nitrogen loading
strategy may be hindered by the projected population growth in the watershed. A 20% increase in
population, and a 7% increase in annual nitrogen load, are anticipated by the year 2010. Therefore, if the
projected loading increase (a total of 17 U.S. tons per year) is not prevented or precluded by watershed
management actions, the "hold the line" load management strategy will not be achieved.
STEP 3. DEFINE AND IMPLEMENT NITROGEN MANAGEMENT STRATEGIES NEEDED TO
ACHIEVE LOAD MANAGEMENT GOALS
Local government and agency partners in the TBNEP signed an Intergovernmental Agreement (IA) in
1998 pledging to carry out specific actions needed to "hold the line" on nitrogen loadings. The IA
includes the responsibility of each partner for meeting the nitrogen management goals, and a timetable
for achieving them. How those goals are reached will be left up to the individual communities as defined
by them in their Action Plans. The Tampa Bay National Estuary Program was also renamed the Tampa
Bay Estuary Program as part of the progression from the planning phase to implementation of the
adopted Comprehensive Conservation and Management Plan.
To maintain nitrogen loadings at 1992-1994 levels, local government Action Plans address that portion of
the nitrogen target which relates to non-agricultural stormwater runoff and municipal point sources
within their jurisdictions, a total of 6 U.S. tons of nitrogen per year through the year 2010 (Table 1).
To address the remaining 11 U.S. tons of nitrogen of the 17 total per year each year through the year
2010 needed to "hold the line" (attributed to atmospheric deposition, industrial and agricultural sources
and springs), a Nitrogen Management Consortium of local electric utilities, industries and agricultural
interests, as well as the local governments and regulatory agency representatives in the TBEP, was
established (Table 2). The Nitrogen Management Action Plan developed by public and private partners
in the Consortium combines for each bay segment all local government, agency and industry projects that
will contribute to meeting the five year nitrogen management goal. To ensure that each partner was
using similar nitrogen load reduction assumptions for similar projects, guidelines for calculating nitrogen
load reduction credits were developed with the partners, and were used by each of the partners in the
development of their action plans.
The types of nutrient reduction projects included in the Consortium's Nitrogen Management Action Plan
range from traditional nutrient reduction projects such as stormwater upgrades, industrial retrofits and
agricultural best management practices to actions not primarily associated with nutrient reduction, such
as land acquisition and habitat restoration projects. A total of 105 projects submitted by local
governments, agencies and industries are included in the Plan; 95% of these projects address nonpoint
Nutrient Criteria—Estuarine and Coastal Water CS-21
-------
Table 1. Tampa Bay Nitrogen Management Goals
SOURCE
CATEGORY
Old
Tampa Bay
Hillsborough
Bay
Middle
Tampa Bay
Lower
Tampa Bay
Boca Ciega
Bay
TOTAL
%
CUMULATIVE 1995-1999 GOALS FOR NITROGEN REDUCTION/MANAGEMENT
Pinellas
County
0.30
<0.01
<0.01
<0.01
0.85
1.15
1.4
City of
Clearwater
0.20
<0.01
<0.01
<0.01
<0.01
0.20
0.2
City of St.
Petersburg
0.05
<0.01
0.90
<0.01
1.05
2.00
2.4
Hillsborough
County
0.40
4.75
2.50
<0.01
<0.01
7.65
9.1
City of
Tampa
0.10
8.45
<0.01
<0.01
<0.01
8.55
10.2
Manatee
County
<0.01
<0.01
0.50
8.35
<0.01
8.85
10.6
TB
Consortium*
1.05
28.25
7.15
17.00
2.00
55.45
66.1
TOTAL
(reduction in annual load)
(tons)
2.10
41.50
11.05
25.35
3.90
83.85
100.0
CD
l-t-
o
I'
=2.
CD
m
C/J
I— I-
CD
CD
CD
Q.
O
O
Q]
C/J
5T
Q]
l-t-
CD
C/J
* Tampa Bay Nitrogen Management Consortium
O
CO
-------
Table 2. Public and Private Partners of the Tampa Bay Nitrogen Management
Consortium, July 2001
Public Partners:
City of Tampa
City of Clearwater
City of St. Petersburg
Manatee County
Hillsborough County
Pinellas County
Manatee County Agricultural Extension Service
Environmental Protection Commission of
Hillsborough County
Tampa Bay Regional Planning Council
Florida Department of Environmental Protection
Florida Fish and Wildlife Commission/Florida
Marine Research Institute
Southwest Florida Water Management District
U.S. Army Corps of Engineers
U.S. Environmental Protection Agency
Tampa Port Authority
Tampa Bay Estuary Program
Florida Department of Agriculture and Consumer
Services
Private Partners:
Florida Phosphate Council
Florida Power & Light
Tampa Electric Company
Florida Strawberry Growers Association
IMC-Phosphate
Cargill Fertilizer, Inc.
CF Industries, Inc.
Pakhoad Dry Bulk Terminals
Eastern Associated Terminals Company
CSX Transportation
sources and account for 71% of the expected total nitrogen reduction. Half (50%) of the total load
reduction will be achieved through public sector projects, and 50% by industry.
Table 3 summarizes expected reductions from those projects which were completed by the end of 1999.
A total of 134 tons per year reduction in nitrogen loading to Tampa Bay is expected from the completed
projects, which exceeds the 1995-1999 reduction goal of 84 tons per year by 60%. An updated estimate
of nitrogen loadings to the bay from all sources was initiated by TBEP in summer 2001, after which the
effectiveness of the proposed projects in maintaining loads to the bay will be evaluated.
Examples of specific projects and expected nitrogen loading reductions include the following:
Stormwater facilities and upgrades: Stormwater improvements or new facilities include both
public and private examples. Stormwater retrofits using alum injection to urban lakes reduced total
nitrogen (TN) loading by an estimated 6.4 tons per year. Stormwater improvements eliminated an
estimated 2 tons of TN loading per year. Industrial Stormwater improvements at phosphate
Nutrient Criteria—Estuarine and Coastal Water
CS-23
-------
Table 3. Tampa Bay Nitrogen Management Consortium Summary of Goals and Expected
Reductions (cumulative tons TN reduced or precluded/year by the year 2000
Bay Segment
Old Tampa Bay
Hillsborough Bay
Middle Tampa
Bay
Lower Tampa
Bay
Boca Ciega Bay
Total
1995-1999 Nitrogen
Reduction Goal
2.10
41.5
11.1
25.4
3.9
84.0
Expected Reduction: Completed
or Ongoing Projects1
5.1
65.9
21.1
36.2
5.6
133.9
Expected Reduction:
Atmospheric Deposition2
3.6-6.2
13.9-24.0
4.6-7.9
5.7-10.0
1.2-2.1
29.0- 50.2
1 Projects have been completed or are under construction. These summaries do not include reductions expected from atmospheric
deposition reductions.
2 Range of atmospheric deposition reductions expected, based on two methods.
fertilizer factories and transport terminals reduced almost 20 tons TN loading per year by the year
2000.
Land acquisition and protection: Land acquisition and maintenance of natural or low intensity
land uses precludes higher density uses and higher rates of TN loading. Land acquisition precluded
more than 15 tons TN loading per year by the end of 1999.
an
Approved overlay districts requiring additional nutrient control in management areas precluded
estimated 10 tons per year TN loading.
Wastewater reuse: Wastewater reuse programs resulted in a 6.4 ton per year reduction on TN
loading. Conversion of septic systems to sewer reduced TN loading by 1.7 tons per year.
Emissions reduction: Estimated emissions reduction from coal-fired electric generating plants
between 1995-1997 resulted in reductions of NOXemissions of 11,700 - 20,000 tons. To estimate
the reduction of nitrogen deposition which reaches the bay (either by direct deposition to the bay's
surface, or by deposition and transport through the watershed), a 400:1 ratio (NOX emissions units
to nitrogen units entering the bay) is assumed. Expected reductions from atmospheric deposition
thus ranged from 29 to50 tons per year by 1999. To date, emissions reductions have not been
included in the estimated total TN reduction to the bay, pending agreement on estimation methods.
Habitat restoration: Although typically conducted for reasons other than nutrient reduction,
habitat restoration to natural land uses reduces the amount of TN loading per acre in runoff.
Habitat restoration projects have been completed or are underway in all segments of Tampa Bay's
watershed. Estimated TN load reduction from completed habitat restoration projects totaled an
estimated 7 tons per year.
CS-24
Nutrient Criteria—Estuarine and Coastal Water
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Agricultural BMPs: Water use restrictions have promoted the use of microjet or drip irrigation on
row crops (including winter vegetables and strawberries) and in citrus groves. Micro-irrigation has
resulted in potential water savings of approximately 40% or more over conventional systems and an
estimated 25% decrease in fertilizer applied. Nitrogen reduction estimates from these actions total
6.4 TN tons per year.
Education/public involvement: For those projects for which nitrogen load reductions have not
been calculated or measured, but some reductions are expected, the Consortium Action Plan
assumes a 10% reduction estimate until more definitive information is available. These programs
have reduced TN loading by an estimated 2 tons per year.
Industrial upgrades: A phosphate fertilizer mining and manufacturing plant has terminated the use
of ammonia in flot-plants (an element of the fertilizer manufacturing process), resulting in a
reduction of 21 tons per year of nitrogen loading. Other fertilizer manufacturing companies have
upgraded their product conveyor systems, resulting in a TN reduction of more than an estimated 10
tons per year due to control of fertilizer product loss. The termination of discharge by an orange
juice manufacturing plant into a tributary of Tampa Bay has resulted in a reduction of more than 11
tons per year TN loading.
The approach advocated by the TBEP stresses cooperative solutions and flexible strategies to meet
nitrogen management goals. This approach does not prescribe the specific types of projects that must be
included in the Action Plan; Consortium partners have been encouraged to pursue the most cost-effective
options to achieve the agreed-upon goals for nitrogen management. The TBEP will review and revise
nitrogen management goals every five years, or more often if significant new information becomes
available.
SUMMARY
The Tampa Bay management community has agreed that protection and restoration of Tampa Bay living
resources is of primary importance. Through the TBEP process (initiated in 1991), partners have adopted
nitrogen loading targets for Tampa Bay based on the water quality requirements of Thalassia testudinum
and other native seagrass species. A long-term goal has been adopted to achieve 15,400 ha of seagrass in
Tampa Bay, or 95% of that observed in 1950. To reach the long-term seagrass restoration goal, a 7%
increase in nitrogen loading associated with a projected 20% increase in the watershed's human
population over the next 20 years must be offset. Government and agency partners in the Tampa Bay
Estuary Program and private industries and interests participating in the Nitrogen Management
Consortium have identified and implemented specific nitrogen load reduction projects to ensure that
water quality conditions necessary to meet long-term living resource restoration goals for Tampa Bay are
achieved.
For more information:
National Research Council, National Academy of Sciences. 2000. Clean Coastal Waters:
Understanding and Reducing the Effects of Nutrient Pollution. National Academy Press, Washington,
D.C. 405 pages.
Janicki, A. and D. Wade. 1996. Estimating critical nitrogen loads for the Tampa Bay estuary: An
empirically based approach to setting management targets. Technical Publication 06-96 of the Tampa
Bay National Estuary Program. Coastal Environmental, Inc., St. Petersburg, Florida.
Nutrient Criteria—Estuarine and Coastal Water CS-25
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Johansson, J.O.R. and H. Greening. 2000. Seagrass restoration in Tampa Bay: A resource-based
approach to estuarine management. IN: Subtropical and Tropical Seagrass Management Ecology,
Bortone, S. (Ed.). CRC Press, Boca Raton, FL.
Tampa Bay Estuary Program. 1998. Partnership for Progress: The Tampa Bay Nitrogen Management
Consortium Action Plan 1995-1999. Tampa Bay Estuary Program, St. Petersburg, Florida.
CS-26 Nutrient Criteria—Estuarine and Coastal Water
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CASE STUDY
RESTORING CHESAPEAKE BAY WATER QUALITY
Contact: Richard Batiuk, 410/267-5731; batiuk.richard@epa.gov
Original Nutrient Reduction Goal
In 1987, the Chesapeake Bay Program partners set a 40 percent reduction goal for nitrogen and
phosphorus to improve low oxygen conditions in the deep trench of the mainstem bay. The goal was
later defined to apply only to "controllable" sources, and only from the States—Maryland, Virginia,
Pennsylvania—and the District of Columbia are also listed as impaired tidal waters.
All listed impaired waters are scheduled to have a Total Maximum Daily Load or TMDL developed. A
TMDL defines the pollutant load that a waterbody can assimilate without causing violations of water
quality standards and allocates the loading to contributing point sources and nonpoint source categories.
Once a TMDL is established by a State and approved by EPA through regulatory action, it is
implemented through regulatory and nonregulatory programs. A regulatory TMDL covering the entire
64,000 square mile bay watershed will be put in place by 2011 if bay water quality is not restored.
Keeping A Cooperative Approach to Bay Restoration
To avoid potential negative impacts that a regulatory TMDL process might have on the successful,
cooperative efforts being used by the States' tributary strategy programs, the Chesapeake 2000
Agreement lays out a series of commitments directed toward seeking a cooperative solution to restoring
bay water quality by 2010.
The bay watershed partners will define the water quality conditions necessary to support bay living
resources—fish, crabs, oyster, and bay grasses by 2001. These required conditions will be defined
through a series of Chesapeake Bay water quality criteria for dissolved oxygen, water clarity, and
chlorophyll a currently under development.
Important distinct bay and tributary tidal water habitats are being identified and characterized as
designated uses, where the above bay criteria will be applied to fully protect the aquatic living resources.
The States with bay tidal waters—Maryland, Virginia, Delaware, and the District of Columbia—have all
committed to adopting these bay criteria and tidal water designated uses into their individual State water
quality standards by 2003.
Critical to supporting the States' adoption of the bay criteria and refined tidal waters designated uses will
be a baywide Use Attainability Analysis (UAA).
Loading caps on nutrients and sediments needed to meet the bay water quality criteria will be allocated to
major tributary basins and individual States within those basins by December 2001.
Tributary strategies, detailed implementation plans to reach the allocated loading caps will be developed
in cooperation with local watershed stakeholders.
A reevaluation planned for 2005 will provide an opportunity for any necessary mid-course corrections on
the road to restoring bay water quality by 2010.
Nutrient Criteria—Estuarine and Coastal Water CS-27
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Bay Criteria: Defining Restored Bay Water Quality
The Chesapeake 2000 Agreement committed the signatories to the following: "by 2001, define the water
quality conditions necessary to protect aquatic living resources." These water quality conditions are
being defined through the development of Chesapeake Bay specific water quality criteria for dissolved
oxygen, water clarity, and chlorophyll a. Collectively, these three water quality parameters provide the
best and most direct measures of the impacts of too much nutrient and sediment pollution on the bay's
aquatic living resources—fish, crabs, oysters, and underwater bay grasses.
Bay Criteria
Dissolved Oxygen
Fish and other aquatic life require levels of dissolved oxygen to survive. Seasonal algae blooms deplete
dissolved oxygen, potentially rendering deep waters of the bay uninhabitable to certain species, such as
the endangered Atlantic Sturgeon during certain times of the year. Bay dissolved oxygen levels should
be those required by the aquatic communities inhabiting different parts of the bay during different times
of the year, fully reflective of natural conditions.
Chlorophyll a
Measurements of chlorophyll indicate levels of phytoplankton or algal biomass in the water column. Bay
chlorophyll levels should be moderate: not so high as to cause harmful algal blooms that lead to poor
quality food, shading of light in shallow water habitats, and low dissolved oxygen conditions when the
algae die off and sink to the bottom.
Water Clarity
Underwater grasses collectively are an essential component of the bay's living resources habitat.
Decreased water clarity inhibits the growth of underwater bay grasses. Water clarity is adversely affected
by increased sediment loads and algal biomass spurred by excess nutrient inputs to the bay. Bay water
quality conditions should generally provide high water clarity—sunlight penetration—to support
restoration of underwater grasses throughout the bay's extensive shallow water habitats.
Chesapeake Bay Dissolved Oxygen Criteria
Chesapeake Bay Dissolved Oxygen Dynamics
The Chesapeake Bay has a built-in, natural tendency toward reduced dissolved oxygen conditions,
particularly within its deeper waters because of the physical morphology and estuarine circulation. Its
highly productive, shallow waters, coupled with its tendency to retain, recycle, and regenerate the
nutrients delivered from the atmosphere and surrounding watershed set the stage for a nutrient-rich
environment. The mainstem Chesapeake Bay and its major tidal rivers with deep channels coming off
shallower, broad shoal waters, and the significant influx of freshwater flows result in stratification of the
water column, essentially locking off deeper bottom waters from mixing with higher oxygenated surface
waters. Combined together, the retention/efficient recycling of nutrients and water column stratification
lead to severe reductions in dissolved oxygen concentrations during the warmer months of the year,
generally May to September.
Nearshore, shallow waters in the Chesapeake Bay also periodically experience episodes of low to no
dissolved oxygen conditions, in part, resulting from intrusions of bottom water forced onto the shallow
flanks by sustained winds (Carter et al.1978, Tyler 1984, Seilger et al.1985, Malone et al.1986). Diel
cycles of low dissolved oxygen conditions often occur in nonstratified shallow waters where nighttime
water column respiration temporarily depletes dissolved oxygen levels (D'Avanzo and Kremer 1994).
The timing and spatial and volumetric extent of hypoxic and anoxic waters vary from year to year,
largely driven by local weather patterns, timing and magnitude of freshwater river flow and concurrent
CS-28 Nutrient Criteria—Estuarine and Coastal Water
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delivery of nutrients and sediments into tidal waters, and the corresponding springtime phytoplankton
bloom (Officer et al.1984, Seliger et al.1985). In Chesapeake Bay mainstem, the onset of low to no
dissolved oxygen conditions can be as early as April and persist through September, until fall turnover of
the water column. The deeper waters of major tidal tributaries can exhibit hypoxic and anoxic
conditions, with the nature, extent, and magnitude of low dissolved oxygen and the causative factors
varying from river to river.
The scientific underpinnings of these Chesapeake Bay specific criteria have been in the works for
decades. Seasonal low dissolved oxygen conditions in the Chesapeake Bay were first documented in the
1930s (Newcombe and Home 1938). Basic understanding of dissolved oxygen dynamics, critical to
derivation of criteria reflective of ecosystem process, began with the research cruises of the Chesapeake
Bay Institute from the 1950s through the late 1970s. A 5-year multidisciplinary research program
starting in the late 1980s, funded by the Maryland and Virginia Sea Grant Program, yielded significant
advances in understanding of all facets of oxygen dynamics, effects, and ecosystem implications (Smith
et al. 1991). These investigations laid the groundwork for more management-focused applications of the
science.
Chesapeake Bay Dissolved Oxygen Restoration Goal
Published in 1992, the Chesapeake Bay dissolved oxygen restoration goal was developed in response to
the Chesapeake Executive Council's commitment "to develop and adopt guidelines for the protection of
water quality and habitat conditions necessary to support the living resources found in the Chesapeake
Bay system and to use these guidelines." The dissolved oxygen restoration goal consisted of a narrative
statement supported by specific target dissolved oxygen concentrations applied over specified averaging
periods and locations. Dissolved oxygen effects information was compiled for 14 identified target
species1 offish, molluscs, and crustaceans as well as for other supporting benthic and planktonic species
within the bay food web. The target concentrations and their specified temporal averaging and spatial
application were determined from analysis of dissolved oxygen levels that would provide the levels of
protection described within the narrative restoration goal. Best professional judgment was used in areas
where there were gaps in the information base on dissolved oxygen effects available a decade ago.
The original dissolved oxygen restoration goal and its supporting framework made three breakthroughs at
that time of significance to supporting derivation and management application of the Chesapeake Bay
specific dissolved oxygen criteria within this document. The dissolved oxygen target concentrations
varied with vertical depth through the water column as well as horizontally across the expanse of the bay
and its tidal tributaries, directly reflecting variations in required levels of protection for different living
resource habitats. The averaging periods for each target concentration were tailored to specific habitats,
recognizing that short-term exposures to concentrations below the target concentrations were allowable
and still protective of living resources. The dissolved oxygen goal document contained a methodology
through which water quality monitoring data and model-simulated outputs collected over varying
frequencies could be directly assessed in terms of the percentage of time that areas of bottom habitat or
volumes of water column habitat were predicted to meet or exceed the applicable target dissolved oxygen
concentrations.
Regionalizing the EPA Virginian Province Saltwater Dissolved Oxygen Criteria
With the publication of the EPA Ambient Water Quality Criteria for Dissolved Oxygen (Saltwater): Cape
Cod to Cape Hatter as came a decade's worth of systematically developed dissolved oxygen effect data
along with synthesis and close evaluation of several decades of effects data published in the scientific
These target species were from a larger list of commercially, recreationally , and ecologically important species
reported in Habitat Requirements for Chesapeake Bay Living Resources-Second Edition (Funderburk et al. 1991).
Nutrient Criteria—Estuarine and Coastal Water CS-29
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peer reviewed literature (Thursby et al.2000). The approach to derive these dissolved oxygen criteria
combined features of the traditional water quality criteria with a new biological framework. A
mathematical model was used to integrate time (replacing the concept of an averaging period) and
establish protection limits for different life stages (i.e., larvae verus juveniles and adults). Where
practical, data were selected and analyzed in manners consistent with the Guidelines for Deriving
Numerical National Water Quality Criteria for the Protection of Aquatic Organisms and Their Uses
(hereafter referred to as the EPA Guidelines) (Stephan et al.1985).
The EPA Virginian Province dissolved oxygen saltwater criteria document addressed three areas of
protection: (1) juvenile and adult survival, (2) growth effects, and (3) larval recruitment effects. In doing
so, the criteria document segregated effects on juveniles and adults from those on larvae. The survival
data on the sensitivity of the juveniles and adults are handled in a traditional EPA guidelines manner. To
address cumulative effects of low dissolved oxygen on larval recruitment to the juvenile life stage (i.e.,
larval survival as a function of time) a new biological approach was taken. These criteria were derived
using a mathematical model that evaluates the effect of dissolved oxygen conditions on larvae by
tracking the intensity and duration of low dissolved oxygen effects across the larval recruitment season.
Protection of larvae of all species is provided by using low dissolved oxygen effects data on larval stages
of nine sensitive estuarine/coastal organisms.
The Virginian Province saltwater dissolved oxygen criteria document and its underlying effects database
and methodologies were structured to support regional specific derivation of dissolved oxygen criteria
tailored to the species, habitats, and nature of dissolved oxygen exposure regimes of different estuarine,
coastal, and marine waters. The segregation by life stages allows the criteria to be factored into the
refined tidal water designated uses, which themselves, in part, reflect use of different habitats by
different life stages. This segregation by life stage is a significant difference from traditional aquatic life
criteria.
However, the Virginian Province saltwater criteria were not explicitly set up to address natural vertical
variations in dissolved oxygen concentration. If Chesapeake Bay specific criteria were derived through a
strict application of the EPA saltwater criteria methodology, there would not be the flexibility needed to
tailor each set of criteria to the refined tidal water designated uses. The resultant bay criteria would be
driven solely by larval effects data irrespective of depth and season.
The Chesapeake Bay specific criteria were derived through the regional application of the Virginian
Province effects database and application of traditional toxicological and new biological-based criteria
derivation methodologies. Chesapeake Bay specific science was factored directly into each step of the
criteria derivation process. The extensive Virginian Province dissolved oxygen effects database was first
focused down on only Chesapeake Bay species and then supplemented with additional Chesapeake Bay
species effects data from the scientific literature. The Virginian Province larval recruitment model was
modified to better reflect Chesapeake Bay conditions, with its application broadened to include
additional Chesapeake Bay species. Finally, specific steps were taken to factor the requirement to
provide protection of species listed as threatened/endangered in Chesapeake Bay into the bay-specific
criteria.
Current State water quality standards generally require 5 mg/L of dissolved oxygen throughout all of the
bay's waters—from the deep trench near the bay's mouth to the shallows at the head of the bay. Even
though the 5 mg/L standard is baywide, bay region scientists believe natural conditions dictate that in
some sections of the bay, such as the deep channel, bay waters cannot achieve the current 5 mg/L
standard during the warmer months of the year. Additionally, scientists believe other areas, such as
prime migratory fish spawning areas, require higher levels of dissolved oxygen to sustain life during the
CS-30 Nutrient Criteria—Estuarine and Coastal Water
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late winter to early summer timeframe. The amount of oxygen needed in the bay tidal waters depends on
specific needs of the aquatic living resources and where they live and during which time of the year they
live there.
The Chesapeake Bay dissolved oxygen criteria vary significantly across the five proposed tidal water
designated uses to fully reflect the wide array of species living in these different bay habitats (Figure 1).
These working draft dissolved oxygen criteria were developed by the Chesapeake Bay Dissolved Oxygen
Criteria Team, a bay region team composed of scientists, State and federal managers, and technical
stakeholders (Table 1). A draft document describing the Chesapeake Bay Dissolved Oxygen Criteria in
greater detail is available for review and comment at www.chesapeakebay.net. There is a year-long
process and schedule, including three public reviews, leading to publication of these Chesapeake Bay
specific water quality criteria by EPA by June 2002.
Chesapeake Bay Chlorophyll a Criteria
Chlorophyll a is used to measure the abundance and variety of microscopic plants or algae that form the
base of the food chain in the bay. Excessive nutrients can stimulate nuisance algae blooms, resulting in
reduced water clarity, reduced amount of good quality food and depleted oxygen levels in deeper water.
By its very nature, chlorophyll a is both an integrated biological measure of production of the primary
food source of the entire bay food web as well as a critical indicator of water quality through its direct
role in reducing light penetration and fueling bacterial processes leading to low dissolved oxygen levels.
As stated upfront by Harding and Perry (1997), "chlorophyll a is a useful expression of phytoplankton
biomass and is arguably the single most responsive indicator of N [nitrogen] and P [phosphorus]
enrichment in this system [Chesapeake Bay]." Determining the levels of chlorophyll a which are fully
protective of the refined designated uses of the vast tidal waters that compose the Chesapeake Bay and
tributaries must factor in all the different roles chlorophyll a plays in defining a restored, more balanced
Chesapeake Bay ecosystem.
Dissolved Oxygen
Criteria
Minimum Amount of Oxygen (mg/L)
Needed to Survive by Species
Migratory Spawning &
Nursery Areas
Striped Bass: 5-6
Shallow and Open Water
Hard Clams: 5
Deep Water
Deep Channel
LLJ
Figure 1. Dissolved Oxygen Criteria, Chesapeake Bay.
Nutrient Criteria—Estuarine and Coastal Water
CS-31
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Table 1. Working Draft Chesapeake Bay Dissolved Oxygen Criteria (July 3, 2001)
Designated Use
Migratory spawning and
nursery
Shallow/open water
Deeper water
Deep channel
Criteria Concentration/Duration
7 day mean of 6 mg/La
Instantaneous minimum of 5 mg/L
30 day mean of 5 mg/L
7 day mean of 4 mg/L
Instantaneous minimum of 3.5 mg/L
30 day mean of 5 mg/L
7 day mean of 4 mg/L
Instantaneous minimum of 3.5 mg/L
30 day mean of 3 mg/L
Instantaneous minimum of 1.7 mg/L
30 day mean of 5 mg/L
7 day mean of 4 mg/L
Instantaneous minimum of 3.5 mg/L
Instantaneous minimum of 1 mg/L
30 day mean of 5 mg/L
7 day mean of 4 mg/L
Instantaneous minimum of 3.5 mg/L
Temporal Application
February 15th -June 10th
June 11th -February 14th
All year round
April through September
October through March
April through September
October through March
"Applied to tidal fresh waters with long term averaged salinities less than 0.5 parts per thousand.
The derivation of the Chesapeake Bay chlorophyll a criteria were based on the convergence of several
independent lines of evidence—historical observed concentrations, literature values related to trophic
status, direct contributions to light attenuation, and contribution to dissolved oxygen
conditions—collaborating chlorophyll a concentrations derived as a result of characterizing set of
phytoplankton reference communities.
Phytoplankton Reference Community/Food Quality Connection
Estimates of phytoplankton taxon biomasses were derived from the Maryland and Virginia Chesapeake
Bay Monitoring Program phytoplankton count data (1984-1999) and along with other phytoplankton
indicators—chlorophyll a, pheophytin, and primary productivity—were used to investigate differences in
biomass, taxonomic composition, and food value for the range of water quality conditions currently
experienced in the Chesapeake Bay. The biological data were sorted into categories based on season-
and salinity-specific concentrations/levels of three parameters in the associated water quality data:
dissolved inorganic nitrogen, ortho-phosphate and Secchi depth. Relatively small secchi depths and
excess dissolved inorganic nitrogen and excess ortho-phosphate characterized the Poor water quality
categories. Relatively high light levels and algal growth-limiting concentrations of dissolved inorganic
nitrogen and ortho-phosphate characterized the good water quality categories. Mixed water quality
conditions (i.e., one or two water quality parameters qualified as Better but the other(s) did not) and
extreme subsets of the Poor and Better categories (i.e. Worst and Best) were also investigated.
Qualitative and quantitative measures of the phytoplankton community composition and biomass
CS-32
Nutrient Criteria—Estuarine and Coastal Water
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distributions were then evaluated relative to these water quality classifications and implications for food
quality and quantity for filter feeding fish and shellfish.
Historically Observed Concentrations
Several recent in-depth reviews and evaluations of historically observed (1950s to early 1980s) and
current (1984-1998) chlorophyll a concentrations provide a strong basis for collaborating the Chesapeake
Bay specific chlorophyll a criteria (Harding 1994, Harding and Perry 1997, Olson and Lacouture, in
review). Using information from five decades of water quality data provides insights into both
chlorophyll a concentrations that are attainable under a range of otherwise natural conditions
(meteorological, river flow, tidal flushing) as well as concentrations reflective of a healthier bay
ecosystem.
Literature Values Related to Trophic Status
Throughout the scientific literature, there are several defining papers which through synthesis of a wide
array of data from many different aquatic systems center down on ranges of conditions reflective of
different trophic states of water bodies (e.g., Wetzel 1985, Ryding and Rast 1989, Smith et al. 1992).
Chlorophyll a is a principal parameter quantified within these literature reviews. The strength of this
collaborative line of evidence is that information is drawn from diversity of systems across the spectrum
of healthy to clearly eutrophied water bodies. This approach provides insights into common
characteristics associated with trophic status that can not be drawn through the study of a single, although
large, water body like Chesapeake Bay.
Direct Contributions to Light Attenuation
Over the past four decades, the Chesapeake Bay ecosystem has had an extensive, widely distributed
underwater grass community undergo severe declines followed by a decade and a half slow but steady
recovery. The bay management and scientific communities have invested significant resources in the
investigation of this grand natural experiment, learning much about the causes of the decline and
potential solutions for continued, yet accelerated restoration. Two comprehensive technical synthesizes
of this wealth of scientific knowledge and insights have been published which provide direct quantitative
insights into the role of chlorophyll a in the recovery of underwater bay grasses (Batiuk et al.1992,
2000). This collaborative line of evidence draws on the chlorophyll a connection to reductions in light
penetration through the water column.
Contribution to Dissolved Oxygen Conditions
It is well known and documented that algae uneaten by higher trophic levels—zooplankton, oysters and
fish of all kinds—becomes the fuel, through its breakdown by bacteria, for reducing dissolved oxygen
levels. Through an analysis of Chesapeake Bay water quality model simulated outputs from scenarios
which simulated dissolved oxygen conditions which met the dissolved oxygen criteria, the model
simulated chlorophyll a concentrations of desired dissolved oxygen conditions were quantified.
Appropriate chlorophyll a levels vary, depending on the salinity of the water. The proposed criteria for
chlorophyll a are split out from tidal freshwater all the way to very salty—polyhaline—waters. Season
of the year is also important, with spring and summer being the most important times of year that high
chlorophyll a levels can impact living resources in the bay.
These working draft chlorophyll a criteria were developed by the Chesapeake Bay Chlorophyll and
Nutrient Criteria Team, a bay region team composed of scientists, State and federal managers, and
technical stakeholders (Table 2). A draft document describing the Chesapeake Bay Chlorophyll a
Criteria in greater detail is available for review and comment at www.chesapeakebay.net. There is a
Nutrient Criteria—Estuarine and Coastal Water CS-33
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Table 2. Working Draft Chesapeake Bay Chlorophyll a Criteria (July 3, 2001)
Salinity Regime
Tidal Fresh
Oligohaline
Mesohaline
Polyhaline
Chesapeake Bay Chlorophyll Criteria (ug/L)
Spring (March-May)
Median
8
10
6
3
Maximum
12
23
27
7
Summer (July-September)
Median
9
6
7
4
Maximum
16
23
16
9
year-long process and schedule, including three public reviews, leading to publication of these
Chesapeake Bay specific water quality criteria by EPA by June 2002.
Connection to Underwater Bay Grasses
The loss of submerged aquatic vegetation, or SAV, from shallow waters of Chesapeake Bay, which was
first noted in the early 1960s, is a widespread, well-documented problem. Although other factors, such
as climatic events and herbicide toxicity, may have contributed to the decline of SAV in the bay, the
primary causes are eutrophication and associated reductions in light availability. The loss of SAV beds
are of particular concern because these plants create rich animal habitats that support the growth of
diverse fish and invertebrate populations. Similar declines in SAV have been occurring worldwide with
increasing frequency during the last several decades. Many of these declines have been attributed to
excessive nutrient enrichment and decreases in light availability.
Chesapeake Bay Water Clarity Criteria
One of the major features contributing to the high productivity of Chesapeake Bay has been the historical
abundance of SAV. There are over 20 freshwater and marine species of rooted, submerged flowering
plants in bay tidal waters. These underwater grasses provide food for waterfowl and are critical habitat
for shellfish and finfish. SAV also affect nutrient cycling, sediment stability, and water turbidity.
The health and survival of these plant communities in Chesapeake Bay and its tidal tributaries depend on
suitable environmental conditions that define the quality of SAV habitat. Key to the restoration of these
critical habitats and food sources is the return of levels of light penetration in shallow waters necessary to
support the survival, growth, and repropagation of diverse, healthy underwater bay grass communities.
Bay Water Clarity Derivation Approach
Through the combined efforts of the bay's scientific and resource management communities, two
internationally recognized technical syntheses of information supporting quantitative habitat
requirements for Chesapeake Bay SAV have been published in the past decade (Batiuk et al. 1992, Batiuk
et al. 2000). Key findings, the underlying light requirements, and management-oriented diagnostic tools
and restoration targets have been reported in the peer reviewed scientific literature (Dennison et al. 1993,
Kemp et al. in review; Gallegos 2001, Koch 2001, Bergstrom in preparation, Carter and Rybicki in
preparation, Karrh in preparation, Kemp et al. in preparation). These two technical syntheses of
worldwide literature, bay-specific research and field studies, and recent model simulation and data
evaluation provide the scientific foundation for the Chesapeake Bay water clarity criteria described here.
Readers are encouraged to consult these two syntheses and the resultant scientific literature papers for
more in-depth technical details and documentation.
CS-34
Nutrient Criteria—Estuarine and Coastal Water
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The Chesapeake Bay specific water clarity criteria derivation follows four successive stages: first,
determination of water-column-based light requirements for SAV survival and growth, then
quantification of the factors contributing to water column light attenuation. The contributions from
epiphytes to light attenuation at the leaf surface are then factored into methods for estimating total light
attenuation. Finally, a set of minimal light requirements are determined as the actual criteria values.
The draft bay criteria propose that water clarity criteria should apply to areas of the bay that are up to 2
meters deep (approximately 6 feet). Areas where SAV never occurred historically, or where natural
factors prevent its growth (e.g., strong currents, rocky bottoms) would be excluded. The water clarity
criteria reflect the different light requirements for underwater plant communities that inhabit low salinity
versus higher salinity shallow water habitats throughout the bay (Figure 2).
These working draft water clarity criteria were developed by the Chesapeake Bay Water Clarity Criteria
Team, a bay region team composed of scientists, State and federal managers, and technical stakeholders
(Table 3). A draft document describing the Chesapeake Bay Water Clarity Criteria in greater detail is
available for review and comment at www.chesapeakebav.net. There is a year-long process and
schedule, including three public reviews, leading to publication of these Chesapeake Bay specific water
quality criteria by EPA by June 2002.
These Chesapeake Bay criteria will be applied to a series of designated uses which, in turn, reflect key
habitats throughout the bay and its tidal tributaries.
Excessive Nutrient and Sediment Loads
The causes of these water quality impairments—excessive loadings of nitrogen, phosphorus, and
sediment—will be addressed through commitments to determine reductions in loadings needed to
achieve the bay criteria. These required loading reductions will be established as caps on loadings
What's Blocking the Light?
Good Water Clarity
Poor Water Clarity
Percent of sunlight
reaching leaves:
•9% in low salinity
waters
Sediment and other
particles in the water
•15% in high salinity
waters
Algae in the water
Algae on the leaves
equals
Very low percentage
of sunlight reaching
leaves - Bay grasses
grow poorly or die.
Figure 2. Bay Water Clarity.
Nutrient Criteria—Estuarine and Coastal Water
CS-35
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Table 3. Working Draft Chesapeake Bay Water Clarity Criteria (July 3, 2001)
Habitat Category
Tidal fresh shallow water
Oligohaline shallow water
Mesohaline shallow water
Polyhaline shallow water
Criteria Concentration
(percent ambient light)
9%
9%
15%
15%
Temporal Application
April - October
April - October
April - October
March-May, Sept-Nov.
allocated to each tributary basin within the Chesapeake Bay watershed. This approach is consistent with
EPA's regional establishment of ambient concentration-based nutrient criteria, but places more emphasis
on the water quality parameters with a direct impact on aquatic living resources. Through this approach
nutrients and sediments are addressed directly through caps on loading determined through application of
the linked bay airshed-watershed-tidal water quality models and analysis of Chesapeake Bay Monitoring
Program data in place of the development of ambient nutrient and sediment criteria.
Relationship with National Efforts to Develop Nutrient Criteria
At the same time, a parallel effort is currently underway by EPA to develop ecoregion specific numerical
nutrient criteria across the country to meet the objectives of the Clean Water Action Plan. A nutrient
criteria team has been established by EPA-Region III to implement the National Nutrient Strategy issued
by EPA last year for the mid-Atlantic region. The EPA Region III team is focusing its nutrient criteria
development efforts on the free flowing stream, rivers, lakes, and wetlands within the mid-Atlantic
States, not the Chesapeake Bay tidal waters. Whereas the bay criteria are focused on dissolved oxygen,
water clarity and chlorophyll a, the EPA Region III team is developing ambient concentration criteria for
total nitrogen, total phosphorus, chlorophyll, and turbidity.
The advanced scientific understanding of water quality impacts on aquatic bay living resources combined
with the state of the art linked bay airshed-watershed -water quality models enabled the bay watershed
partners to develop criteria for water quality measures directly influencing aquatic resources. The cause
of reduced water quality conditions—too much nitrogen, phosphorus, and sediment—will be addressed
through the establishment of loading caps. The bay models enable the partners to effectively translate the
desired dissolved oxygen, water clarity, and chlorophyll a conditions back into reduced loadings of
nutrients and sediments from the surrounding watershed and airshed. Bay science has shown that it is the
delivered loads of nutrients and sediment, not just the ambient concentrations, that have had an impact on
oxygen, light, and algae levels in the bay tidal waters.
Bay Tidal Water Designated Uses
Because conditions throughout the Chesapeake Bay tidal water habitats differ based on depth, salinity
and season, a uniform baywide water quality standard does not take into account the varying needs of
different plants and animals. As a result, current State water quality standards, which differ between the
four jurisdictions with tidal waters, need to be revised and expanded to account for the natural variability
in conditions found throughout the bay. Each of the three bay criteria will differ from one region of the
bay and its tidal tributaries to another, as determined by the plants and animals residing in that area.
Once the bay criteria and the tidal water designated uses are adopted as State water quality standards,
these tailored set of standards will apply to similar habitats across all jurisdictions.
An area's designated use refers to a waterbody's function—such as fishable or swimable—and takes into
account the use of the water body for public water supply, the protection offish, shellfish, and wildlife,
as well as its recreational, agricultural, industrial and navigational purposes. The existing Maryland,
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Virginia, Delaware, and District of Columbia designated uses for the bay's tidal waters do not fully
reflect the wide variety of different habitats found throughout the bay and its tidal tributaries. Where two
jurisdictions boundaries join, each State has different designated uses for the same waterbodies.
Five refined Chesapeake Bay tidal water designated uses have been established to more fully reflect the
different aquatic living resource communities inhabiting a variety habitats and, therefore, the different
intended aquatic life uses of those tidal habitats (Figure 3).
The Migratory Spawning & Nursery designated use is the propagation and growth of balanced
indigenous populations of ecologically, recreationally, and commercially important anadromous, semi-
anadromous, and tidal fresh resident fish species inhabiting spawning and nursery grounds.
The Shallow Water designated use is the propagation and growth of balanced, indigenous populations of
ecologically, recreationally, and commercially important fish, shellfish and underwater grasses inhabiting
shallow waters habitats.
The Open Water designated use is the propagation and growth of balanced, indigenous populations of
ecologically, recreationally, and commercially important fish, and shellfish species that inhabit open
water habitats.
The Deep Water designated use is the propagation and growth of balanced, indigenous populations of
ecologically, recreationally, and commercially important fish and shellfish species inhabiting deep water
habitats.
The Deep Channel designated use is to provide a refuge for balanced, indigenous populations of
ecologically, recreationally, and commercially important fish species that depend on deep channel
Refined Designated Uses for
Chesapeake Bay and Tidal Tributary Waters
A. Cross Section of Chesapeake Bay or Tidal Tributary
Open Water
Deep Watei
Deep Chanm
B. Oblique View of the "Chesapeake Bay" and its Tidal Tributaries
J-Migratory Finfish
Spawning and
Nursery Habitat
Shallow Water
Habitat
Deep Channe
Figure 3. Refined Designated Uses for Chesapeake Bay Tidal Tributary Waters.
Nutrient Criteria—Estuarine and Coastal Water CS-37
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habitats for overwintering during colder months of the year and the propagation and growth of benthic
infaunal and epifaunal worms and clams that provide food for bottom feeding fish and crabs.
These tidal water designated uses were developed by the Chesapeake Bay Water Quality Standards
Coordinators Team, a bay region team composed of water quality standards coordinators from all six
States, the District of Columbia, EPA Region 2, 3, and headquarters offices. Table 4 shows how refined
tidal water designated uses relate to the bay criteria.
The watershed partners are evaluating the refined tidal water designated uses and the applicable bay
criteria through a baywide use attainability analysis. The final tidal water designated uses will be
adopted by Maryland, Virginia, Delaware, and the District of Columbia, along with the applicable bay
water quality criteria into their State water quality standards by 2003. These refined designated uses will
add more specifics to the existing State designated uses and apply consistently across jurisdictions for
similar habitats.
Baywide Use Attainability Analysis
The Chesapeake 2000 Agreement commits the States with bay tidal waters—Maryland, Virginia, and
Delaware—and the District of Columbia to adopt into their State water quality standards as consistent set
of bay criteria and designated uses across bay tidal habitats. Whenever there is a proposed change in
water quality standards, such as that being undertaken for Chesapeake Bay waters, it is necessary to
assess attainment of the designated uses and underlying criteria. Such an assessment is called a Use
Attainability Analysis or UAA.
A UAA is used by States to justify changes to their water quality standards by assessing the physical,
chemical, biological, economic, or other factors affecting attainment of the designated use. The UAA
describes the scientific attributes of the waterbody, both natural conditions and conditions brought about
by human contribution. If the attributes of the waterbody make attaining the use impossible, or if there
are economic reasons why the use cannot be attained, the UAA is used to clearly document these reasons.
Finally, the UAA describes how the proposed standards will protect existing uses. All six bay watershed
States—New York, Pennsylvania, Maryland, Virginia, West Virginia and Delaware—along with the
District of Columbia and EPA are working together with bay watershed partners to carry out such an
Table 4. Chesapeake Bay Criteria Needed for Protection
of the Proposed Tidal Waters Designated Uses
Migratory
Spawning and
Nursery
Shallow Water
Open Water
Deep Water
Deep Channel
Dissolved
Oxygen
•
•
•
•
•
Chlorophyll a
•
•
•
Water Clarity
•
CS-38
Nutrient Criteria—Estuarine and Coastal Water
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assessment. A use attainment assessment on a scale as large as the 64,000 square mile Chesapeake Bay
watershed has never been carried out.
Adopting Bay Criteria as State Water Quality Standards
Water quality standards combine water quality criteria and designated uses to produce a target numeric
value assigned to a waterbody that, if achieved, will maintain healthy water quality. Through the
Chesapeake 2000 Agreement, Maryland, Virginia, and the District of Columbia are committed to
adopting the new bay criteria-dissolved oxygen, water clarity, and chlorophyll a—along with the refined
tidal water designated uses as State water quality standards. Delaware, which shares bay tidal waters
with Maryland in its portion of the Nanticoke River watershed, has made the same commitment through
the six-State memorandum of agreement. Together, the States and the District must achieve these new
bay-specific water quality standards needed to support restored estuarine ecosystem if the Chesapeake
Bay is to be removed from the list of impaired waters.
Why New State Standards
Existing State water quality standards are applied broadly across each State's tidal waters, without
recognition of the variety of habitats. Each State has different water quality standards applied to the
same tidal waters, whereas the aquatic living resources in bay habitats, which do not recognize these
jurisdictional boundaries, may have the same water quality needs. Currently, dissolved oxygen is the
only numerical water quality standard adopted by all three States and the District of Columbia that
addresses nutrient- and sediment-related water quality pollution problems.
So compliance with existing State water quality standards will not fully protect the living resources in the
bay waters. In some critical habitats of the bay, specifically migratory fish spawning and nursery areas,
existing State water quality standards will not fully protect more sensitive life stages. In other cases,
reaching existing standards is not possible owing to natural conditions found in deeper bay waters during
the warmer months of the year. Existing State water quality standards do not include measures to protect
underwater bay grasses or fully support good quality fish food.
Setting and Allocating New Cap Loads
The Chesapeake 2000 Agreement commits the signatories to determining the nutrient and sediment load
reductions necessary to achieve the water quality conditions that protect aquatic living resources. Those
load reductions will then be assigned or allocated to each major tributary basin in the form of cap loads.
Cap loads are the maximum amounts of pollutants allowed to flow into a waterbody and still ensure
achievement of State water quality standards. In this case, the water quality standards will be the new
bay criteria and refined tidal water designated uses (currently in draft) to be adopted by the States of
Maryland, Virginia, and Delaware, and the District of Columbia into their standards by 2003.
Available Tools and Information
The Chesapeake Bay watershed partners will use the Chesapeake Bay Airshed, Watershed and Estuary
Models, the USGS SPARROW model, along with Chesapeake Bay Monitoring Program data, to help
determine these cap loads for nitrogen, phosphorus, and sediment. These models are mathematical
representations that simulate the real world, interpreting various levels of actions (management
scenarios) to reduce different amounts of pollutant loads. These scenarios are run through the models to
determine how to achieve baywide attainment of the bay water quality criteria for dissolved oxygen,
water clarity, and chlorophyll a as applied to the tidal water designated uses.
Cap Setting and Allocation
These models and other available information will be used to allocate loading caps to the nine major
tributary basins-Susquehanna, Upper Western Shore, Patuxent, Potomac, Rappahannock, York, James,
Nutrient Criteria—Estuarine and Coastal Water CS-39
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Upper Eastern Shore, and Virginia Eastern Shore (Figure 4). Each State and the District will bear a
proportional burden for achieving and maintaining the cap based on their existing pollutant loadings,
progress to date, effectiveness and cost efficiency considerations, and their pollutant loading effects on
different tributaries.
For multijurisdictional waters like the Susquehanna, Potomac, and Eastern Shore basins, the linked
watershed and bay water quality models will be used to further allocate cap load responsibilities to each
State.
Working with their local stakeholders, individual States to further subdivide their major tributary basin
load cap allocations into the 37 State-defined tributary strategy sub-basins.
A comprehensive 2-year schedule has been set up to coordinate the efforts of the six watershed States,
the District of Columbia, and the many other involved bay watershed partners. The schedule also will
ensure direct and continued involvement of local stakeholders and the general public during the entire
cap load setting and allocation process. The Chesapeake Bay Water Quality Steering Committee,
composed of senior managers from the seven watershed jurisdictions, EPA regional and headquarters
offices, Chesapeake Bay Commission, river basin commissions, and involved stakeholders, has the
overall responsibility overseeing and reaching agreement on the cap load allocations. Many of the
subcommittees and workgroups within the Chesapeake Bay Program committee structure will be carrying
out the technical, modeling, data interpretation, economic analysis, policy evaluation, and
communication work in support of setting and allocating the nutrient and sediment cap loads.
Figure 4. The Nine Major Basins.
CS-40 Nutrient Criteria—Estuarine and Coastal Water
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Tributary Strategies: Local Watershed Implementation
The Chesapeake 2000 Agreement commits the bay watershed partners to "complete a public process to
develop and begin implementation of revised Tributary Strategies to achieve and maintain the assigned
loading goals." Tributary strategies are detailed descriptions of planned local actions—riparian forest
buffer replanting, wastewater treatment upgrades, nutrient management on farms, stormwater treatment,
stream restoration, and many others—and a schedule for undertaking those actions necessary to reduce
nutrients and sediment loads from each tributary watershed to reach the assigned loading cap by 2010.
Development of tributary strategies has been a very driven public process with the direct involvement by
local governments, watershed associations, regional organizations, and a wide variety of other interested
local stakeholders. In creating the strategies, the States, and the District of Columbia work closely with
those groups and individuals within each respective watershed who will be directly involved in
implementation strategy. Together, they explore and evaluate a wide variety of point and nonpoint
source pollution control measures. They then draft a strategy using the most effective reduction options
to achieve the cap load allocated to their tributary strategy basin.
The existing Tributary Strategies were designed to achieve the 1987 Bay Agreement goal of a 40 percent
reduction in nutrient loads from controllable sources from 1985 levels. Copies of these existing tributary
strategies are available on-line through the respective Maryland, Virginia, Pennsylvania, and the District
of Columbia tributary strategy web pages.
To restore the tidal water conditions necessary to sustain the bay's fish, crabs, oysters and underwater
grasses will likely require greater reductions in nutrients in many areas then called for by the existing
tributary strategies. In addition, the water clarity conditions needed to restore underwater bay grasses
can not be achieved without significant reductions in sediments loads to the tidal waters. With New
York, Delaware and West Virginia joining as bay watershed partners through a six-State memorandum of
understanding, new tributary strategies will be developed for these States' portions of the bay watershed
not previously addressed under the existing tributary strategies.
The new and revised tributary strategies will now cover all sources of nutrient and sediment pollution,
including air sources, across the entire 64,000 Chesapeake Bay watershed. Tributary strategies will
address nutrient and sediment loading caps allocated to 37 sub-basins across the bay watershed by the
bay watershed partners. Loading reductions required through local stream and river segment regulatory
TMDLs will be directly integrated into each respective Tributary Strategy as part of the overall effort to
effectively "blend" the regulatory TMDL program and cooperative Chesapeake Bay Program.
Information on Local Watersheds
A wide array of information is available on local watersheds within the Chesapeake Bay basin, including
information directly relevant to the overall process for setting and allocating new cap loads through the
Chesapeake Watershed Profiles. Through this point and click information system linking the bay
watershed partners, one can access information on pollution sources, recent modeling results, status of
bay criteria attainment, long-term trends in water quality and living resources, draft cap allocations, and
more from the entire bay basin scale to local watersheds.
Nutrient Criteria—Estuarine and Coastal Water CS-41
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CASE STUDY
A PERSPECTIVE FROM WASHINGTON STATE
Jan Newton, Washington State Department of Ecology; Randy Shuman, King County Department of
Natural Resources; Greg Pelletier, Washington State Department of Ecology
The issue of nutrient control for marine receiving waters in Washington State has its origin in a landmark
case concerning freshwater eutrophication and lake restoration. In the early 1950's, Lake Washington, a
large lake (85 km2) situated near Seattle, Washington, was showing warning signs of ecological
deterioration. Following unregulated dumping of sewage from a growing urban population into the lake,
classic signs of eutrophication were observed, including blooms of Oscillatoria rubescens, reduced water
transparency, and very low nitrogen to phosphorus concentration ratios. The situation was studied
extensively by Dr. W. T. Edmondson, a professor at the University of Washington, who explained that
the changes in the lake were directly attributable to nutrient loading from sewage and wastewater
(Edmondson, 1991). Edmondson made these facts known not only to the scientific community but also
to the public and local government. The case resulted in the diversion of sewage away from the lake and
is a classic example of how scientific observations and understanding were used to shape public policy.
The sewage diversion and subsequent lake recovery were a success, with current day water quality of the
Lake far exceeding that observed during the 1950's-60's.
The solution to this classic case was to divert wastewater from Lk. Washington to nearby Puget Sound, a
large inland sea linked to the Pacific Ocean via the Strait of Juan de Fuca. Studies by Dr. G. C.
Anderson, an oceanographer at the University of Washington, showed that phytoplankton in the Sound in
the vicinity of the outfall proposed to handle the diverted wastewater were limited by light and mixing,
not by nutrients. Thus, diversion of effluent from the Lake to the Sound was not "shifting the problem"
but rather was an ecologically sound solution. This understanding of both limnology and oceanography
laid the conceptual foundation for the formation of a large publicly funded agency (Municipality of
Metropolitan Seattle, or "Metro"). It was proposed that Metro would build a new outfall at West Point,
on the Main Basin of Puget Sound, divert the sewage from Lk. Washington to West Point, and thus
eliminate the nutrient enrichment problem. However, the mandate to create Metro to carry forward these
actions had to be approved in a public election first. Much controversy was associated with the process
chronicled well in Edmondson's book "The Uses of Ecology" (Edmondson, 1991). Among numerous
lines of objection, some public opinion maintained that Puget Sound's ecological health would be
destroyed, despite Anderson's observations. The proposed action required two election attempts before it
was approved in 1958. It is notable that the election passed before deterioration of Lk. Washington water
quality was serious; certainly the local conditions were not as serious as the symptoms seen in lakes in
Europe or the Midwest North America. However, deterioration of Lk. Washington conditions did
continue during the five years before the Metro diversion construction commenced. The persistent,
dense, and obnoxious populations of Oscillatoria galvanized public opinion that the Metro diversion was
necessary. Shortly after construction of the diversion, Lk. Washington water quality improved.
The notoriety of this event and the success of the outfall constructed at West Point to not exhibit
observable biological changes in Puget Sound resulted in widespread lore that "Puget Sound" cannot be
eutrophied because the marine waters are not sensitive to nutrient addition. As Anderson's observations
implied, the reason for the success of West Point outfall owes to the deep, well-mixed waters at the site
which are flushed with a residence time on the order of days. Density-driven stratification is minimal,
the phytoplankton are well mixed, and any depth gradients of oxygen and nutrients do not persist. The
outfall, at 71 m depth, diffuses effluent into water that has naturally high concentrations of nitrate and
this additional nitrogen burden is thought to not significantly contribute to phytoplankton nutrition.
CS-42 Nutrient Criteria—Estuarine and Coastal Water
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Ammonium concentrations in excess of normal Puget Sound background levels are observed sporadically
(King County 2001) near the site and may be associated with the effluent.
This example remains well-known, but it is important to note that not all of the reaches, bays, and inlets
of Puget Sound have the same characteristics of West Point and the Main Basin. Greater Puget Sound is
actually composed of several basins: South Puget Sound, Hood Canal, Whidbey Basin, the Main Basin,
and Admiralty Inlet with its adjoining waters with the Strait of Juan de Fuca. The first three of these
basins have considerable freshwater-induced stratification and are much less well-flushed than the Main
Basin. Residence times of these basins range weeks to months. A study funded by EPA to provide a
general review of the state of knowledge regarding nutrient-phytoplankton relations and quantify the
relative nutrient sensitivity of various areas in the Sound highlighted several areas where nutrients
became deplete and N:P ratios suggested nitrogen limitation (Rensel, 1991). Evidence from C-14 uptake
experiments have shown that enhancements of primary production over ambient rates due to added
nitrogen nutrient can be as high as 300% in South Hood Canal (Newton et al. 1995) and 83% in Budd
Inlet, located in South Puget Sound near Olympia (Newton et al. 1998). Based on environmental
attributes and human growth indicators, only a few places within Puget Sound were judged to be
currently exhibiting signs of eutrophication; however, numerous places, particularly in South Puget,
Hood Canal, and Whidbey basins, were assessed to be highly susceptible to future deterioration from
eutrophication (Bricker et al. 1999).
Perhaps the first place within Puget Sound to gain wide attention for nutrient enrichment effects was
southern Puget Sound, including Budd Inlet, Oakland Bay, Eld Inlet, Henderson Inlet, Case Inlet, and
Carr Inlet. The Washington State Department of Ecology sponsored two studies in the 1980s to evaluate
the acceptability of secondary-treated wastewater discharges to marine waters in southern Puget Sound
(URS, 1985, URS, 1986a). This work identified areas where new or expanded discharges were
unacceptable, based on the potential for eutrophication. A simple screening model based on effluent
dilution and flushing was developed to identify the most sensitive areas.
Wastewater discharge into Budd Inlet was implicated in causing nuisance blooms of phytoplankton and
adding to low dissolved oxygen concentrations noted in the bottom waters at the head of the inlet (URS,
1986b). Studies sponsored by the Washington State Department of Ecology developed the first
numerical models to relate the loading of nitrogen to phytoplankton blooms and dissolved oxygen in
Budd Inlet. This work was the impetus for construction of advanced wastewater treatment systems to
remove dissolved inorganic nitrogen from municipal wastewater from the regional facility that discharges
to Budd Inlet. In 1994, the wastewater entering Budd Inlet was treated for nitrogen removal. This has
resulted in substantially lower ambient nitrogen concentrations (Eisner and Newton 1997). However,
regional growth continues and more capacity to process effluent is needed. Currently, the
Lacey-Olympia-Tumwater-Thurston County Wastewater Management Partnership (LOTT) is considering
numerous alternatives to meet this need, including one proposal to discharge more effluent in winter
when phytoplankton growth is minimal. LOTT undertook a large modeling and observational study: the
Budd Inlet Science Study, which will be used to make the final permitting decisions.
Nutrient-induced increases to phytoplankton production with subsequent drawdown of bottom-water
oxygen has a unique character in Puget Sound relative to other North American systems because of the
influence of upwelled Pacific Ocean water. Upwelling favorable conditions off the Washington coast
lead to upwelling of deep ocean waters. These relatively low-oxygenated waters are transported in,
landwards, at depth through the Strait of Juan de Fuca, as the estuarine flow of Puget Sound waters flow
out, seawards, at the surface. The oceanic waters entering the Puget Sound system through Admiralty
Inlet can have oxygen concentrations as low as 5 mg/L, which then spread into the Main Basin and form
the bottom-water of the other basins as well. Seasonally low deepwater oxygen concentrations can be
Nutrient Criteria—Estuarine and Coastal Water CS-43
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found throughout Puget Sound (Newton et al., 1998; King County, 2001). Upwelling is mostly
favorable in late summer, when productivity-related oxygen deficits are also maximal. The additive
nature of human-caused eutrophication oxygen drawdown to this natural low oxygen quality, results in a
smaller margin of error before deleterious effects would be noted.
The success of discharging nutrients into marine waters via regional Puget Sound wastewater plants,
including the West Point facility, may be at a scale that has reached capacity. The Seattle metropolitan
area continues to grow and new Puget Sound regional wastewater facilities are needed. The King County
Council recently approved that a new facility will be required to meet growth demands and is projected
for completion in 2010. The new facility will also discharge into the Main Basin, to the north of the
West Point facility. King County is currently leading an extensive study to investigate impacts from
nutrient loading on the area. These studies include extensive water quality sampling, experiments on the
susceptibility of the waters to nutrient additions and modeling experiments to predict future impacts.
The story of nutrient control in Puget Sound continues to evolve. Ecologically sound decisions regarding
nutrient management are dependent on two variables: (1) the population producing the effluent, its size,
growth rate, and scale relative to the receiving waters; and (2) the sensitivity of the particular region of
the Sound where the effluent is to be discharged to nutrients. Unfortunately, both of these attributes are
highly variable within the Puget Sound regional area, making nutrient management decisions a challenge
and arguing for the utility of careful scientific studies in companion with planning efforts.
CS-44 Nutrient Criteria—Estuarine and Coastal Water
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APPENDIX A
CONDITIONS FOR BLOOM DEVELOPMENT:
INTERPLAY AMONG BIOGEOCHEMICAL, BIOLOGICAL, AND
PHYSICAL PROCESSES
Overview
In temperate estuaries, the spring bloom typically is dominated by diatoms and occurs when freshwater
delivers adequate amounts of N, P, and Si and other nutrients to the system. In deep estuaries, the spring
freshwater inflow also provides for vertical density stratification where enough of the euphotic zone
resides near the surface to allow phytoplankton to achieve net biomass production (i.e., total
photosynthesis minus respiration is positive). During this time solar insolation and water temperature are
on the increase. For a bloom to develop, other conditions must be met. Physical flushing and local
dispersion of a water parcel must be less than the doubling time for cells. Biological grazing rate must
not be so large as to consume phytoplankton faster than cells' doubling time.
Biogeochemical Processes
Three aspects of N and P biogeochemistry help explain whether N or P dominates nutrient limitation in
estuaries (i.e., relative nitrogen fixation rates, denitrification, and sediment regeneration of P). The first
aspect involves evidence suggesting that nitrogen fixation is less effective in marine than in freshwater
systems in making up nitrogen deficits (Howarth 1988; Schindler 1974). This finding has major
implications for long-term coastal and open-ocean nutrient overenrichment, because N fixation is so
inefficient that any balance in the N versus P limitation occurs in terms of geological time.
The second aspect involves one of the greatest differences in nutrient biogeochemical cycles between
freshwater and marine systems. The superior capacity of freshwater versus marine systems to retain P in
sediments through interactions with iron has profound implications. Nearly all the P deposited in marine
sediments is remineralized annually (Caraco et al. 1990) and depends heavily on the sulfate
concentration, which can be used as a surrogate for salinity. Thus, P in freshwater sediments is bound
more tightly, and proportionally less is released back into the water column. Also, P release from marine
sediments is temperature dependent, and its maximum release during the summer helps explain the
tendency for increasing water column concentrations of P to occur during that season in many estuaries
(Nixon et al. 1980). Estuaries with a well-developed tidal freshwater zone might be expected to be more
P-limited than estuarine systems with small tidal fresh areas.
More details on nutrient cycling in subtropical and tropical marine waters, systems much less studied
than northern temperate estuaries, are provided by Bianchi et al. (1999). The extensive coastal wetland
systems (e.g., marshes and mangroves) that border the Gulf of Mexico provide environments where
chemical transformations and storage of nutrients occur. Also, extensive seagrass meadows apparently
tie up inorganic N and P so that relatively less remains free in the water column. In general, DIN and
PO4 concentrations are much lower in northern Gulf river-dominated estuaries (e.g., Mobile and
Apalachicola Bays) than similar U.S. East Coast systems, presumably because of the lower point sources.
Local groundwater sources are important and water quality managers should be aware of them.
However, local estuarine point sources of N and P may alter nonpoint source patterns (e.g., paper mills
and wastewater treatment facilities) (Livingston 200 la).
Denitrification is the third aspect that plays a role in N limitation in estuaries and coastal waters (Nixon
et al. 1996). Denitrification is the process whereby nitrate is converted to gaseous N2 and N2O and
thereby made unavailable. Denitrification provides a sink for N in estuarine systems. Shelf waters
generally are too deep to provide enough sediment-water column contact for a quantitatively significant
magnitude of denitrification to occur. Bottom-water anoxia limits nitrification and hence denitrification
Nutrient Criteria—Estuarine and Coastal Waters A-1
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in the high-sulfide sediments where nitrification and denitrification are coupled (Jenkins and Kemp
1984). Knowledge of the magnitude of denitrification can help the water quality manager predict the
nutrient overenrichment response of an estuary, because N that is denitrified is largely unavailable to
support primary production.
Biological Processes
The relative importance of biological grazing should be assessed, because when a nutrient problem
occurs it is evidence that enrichment has exceeded the ability of the system to maintain a steady state in
net biomass production at pre-enrichment levels. For example, major changes in the biology of estuaries
in terms of particle filtering capacity (e.g., oysters: Newell 1988) and probably filter feeding finfish (e.g.,
menhaden) can modify phytoplankton primary production, although the quantitative effectiveness of such
cropping is scientifically unsettled. Nutrient overenrichment may drive marine waters toward smaller
algae and other microbes (Jonas 1992) where organic carbon flows more to the microbial loop (Hassett et
al. 1997), versus more direct flow to copepods and higher trophic levels (e.g., finfish). This shift in the
food web may be a significant factor in how estuaries "assimilate" increased nutrient inputs (Roelke et al.
1999; Roelke 2000).
Physical Processes and Factors
Conceptual Framework
Smith (1984) argued that there is no inherent difference in nutrient limitation between lakes and the
ocean. Abundant evidence indicates that phytoplankton net primary production in north temperate lakes
tends to be P limited, and phytoplankton production in the ocean as a whole is potentially moderately P
limited but at higher P concentrations than lakes. This conclusion is supported by the observation that
TN:TP ratios of the surface ocean are usually well in excess of the Redfield ratio (Guildford and Hecky
2000). Local deviations have been detected. In contrast to lakes and oceans, estuaries and coastal shelf
waters tend to be N limited, with some exceptions. Water quality managers may question the reason for
this, as the three case studies described earlier, especially temperate estuaries and the coastal shelf,
appear to be N limited and not P limited. Such an understanding is basic to arguments about cause and
effect and also what ecosystem conditions drive a coastal ecosystem toward N limitation or P limitation.
Smith provided an explanation that still has merit.
Smith posited that the apparent difference in limiting nutrient between lakes and oceans lies, in part, with
the relative rates for material exchange via physical processes of advection (i.e., transport of water and
associated constituents) and eddy diffusion (i.e., local transport of material against a concentration
gradient) and biogeochemical processes of N fixation and fixed N loss. The argument is based on field
experiments in marine embayments with little or no freshwater input, so advective transport of nutrients
from the land simplified nutrient budget development. Smith postulates that if the physical exchange
rates are long (e.g., open ocean), then the system would tend toward P limitation because biogeochemical
adjustment of the N:P availability ratio is short compared with long physical exchange rates (e.g., months
to a year or longer). In other words, nitrogen fixation would balance any losses of nitrogen associated
with phytoplankton sedimentation, but P has no atmospheric reservoir or biochemical mechanism for an
equivalent P fixation to occur. If the physical exchange rates are faster than biochemical rates (e.g.,
nitrogen fixation), then net ecosystem production (and by inference net phytoplankton biomass
production) of organic material may be N limited; if the biochemical rates are faster, then net production
will tend toward P limitation. The ratio of the residence time of the water to the biogeochemical turnover
rate indicates the degree to which the hydrodynamic processes dominate or modify estuarine ecosystems
(Day et al. 1989). This is an example of the importance of scaling critical processes (Harris 1986).
Smith's conceptual model should apply to estuaries and shelf waters. Smith's data, analysis, and
synthesis and other empirical data support N limitation in estuaries and coastal shelf environments.
A-2 Nutrient Criteria—Estuarine and Coastal Waters
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The concept of scale is another element of the conceptual framework. Physical processes that modify the
expression of bloom dynamics will best be detected at the ecological level (Figures A-l, A-2). In this
context, Geyer et al. (2000) cite many examples supporting the observation that "at virtually every
spatial scale, within every component of estuarine ecosystems, physical processes influence the
distribution and fate of chemicals (sic including nutrients) and organisms." Physical processes are
involved in the delivery of nutrients to the biota in estuaries and coastal shelf waters, and also
fundamentally influence advective and dispersive processes that transport and retain dissolved and
particulate material, including nutrients and plankton in estuaries and on the coastal shelves. The roles of
physical processes influencing net biomass production of phytoplankton are explored in the main text in
more detail.
Cycles/day
f/p Time
«1 (seconds)
Low
Hgh
Horizontal Vertical
Space metrics
104
10'
phenomena'
104 H
10' -
10* _
Physology
'equilibrium'
10' -
10-'-
10*m -
10'm -
1 m -
10'm -
10'm-
Coaree
Fine
Figure A-l. Scales of phytoplankton ecology. Horizontal and vertical scales
are determined by the respective diffusion coefficients Kh and KL. The time
scales for the algae are determined by the scales of growth (shaded band). The
processes of importance at each scale are noted. In the past it was always
assumed (wrongly) that physiological processes were at equilibrium and that
climatological variability could be ignored. Source: Harris 1986.
Nutrient Criteria—Estuarine and Coastal Waters
A-3
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Figure A-2. The left panel shows the distribution of chlorophyll—an indicator of algal biomass—along the east
coast of the U.S. from Boston to South Carolina as measured from the ocean color satellite SeaWIFs. Note the
higher chlorphyll levels closer to shore, and the much higher levels in enclosed bays, such as Pamlico Sound
(latitude 35°) and Chesapeake Bay (mouth at 37° latitude). The above panel shows chlorophyll distributions within
Chesapeake Bay in more detail, as measured during a phytoplankton bloom. Both images were taken in April 1998.
Source: Howarth et al. 2000.
A-4
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APPENDIX B
ADDITIONAL INFORMATION ON THE ROLE OF TEMPERATURE AND LIGHT
ON ESTUARINE AND COASTAL MARINE PHYTOPLANKTON
Availability in temperature and light values can be plotted as hydroclimographs (e.g., polygons) to
picture their relative seasonal change around the coasts. A temperature increase will lower the density of
seawater and contribute to density stratification. In particular, rapid changes in water temperature
influence the rate of biological metabolic processes, including algal growth rates (Eppley 1972). Some
species exhibit various degrees of thermal adaptation when temperature changes are gradual.
Speculation suggests that if sea temperatures continue to rise as a function of the "greenhouse effect,"
estuarine biotic communities may change over the next several decades as they approach thermal limits.
For example, although the seagrasses Halodule (shoalgrass) and Zostera (eelgrass) now overlap in Core
Sound, NC, a northward migration of Halodule and a retreat of Zostera may occur if water temperatures
rise faster than populations can adapt. Such relationships and their potential alteration probably can be
documented for other biotic groups along other coasts. For example, if temperature now limits regular
flowering of the seagrass Thalassia (turtle grass) along the northern Gulf of Mexico, then increased
flowering may occur if temperatures warm. Such conjectures notwithstanding, however, little
information is available to help assess the consequences of a potential interaction between an increased
temperature rise and increased nutrient supply on seagrasses.
Light has a fundamental role in aquatic primary production and is essential in the development of models
to estimate phytoplankton primary production (Behrenfeld and Falkowski 1997) and submerged aquatic
vegetation (Dennison et al. 1993). Many concepts in aquatic ecology are based on the light gradient
(Huisman 1999) (e.g., diel plankton vertical migration, benthic animals migrating out of sediments, depth
of euphotic zone, and mixing depth). Phytoplankton growth, nutrient relationships, light, and other
physical processes interact in a feedback system. Although light may be adequate and all other
requirements met for their formation, blooms may not form if dispersive processes are greater than algal
cell doubling time (Kierstead and Slobodkin 1953; Lucas et al. 1999). Tidal ranges greater than
approximately 2.0 m apparently disperse phytoplankton faster than cell doubling time, even if nutrient
conditions would be supportive of a bloom (Monbet 1992).
Both the vertical distribution of phytoplankton abundance and community composition are frequently
changing in the water column. Swimming through use of a flagellum, especially by dinoflagellates, and
changes in cell density through physiological mechanisms allow modest vertical mobility against weak
mixing forces. If the mixing depth is substantially greater than the euphotic zone depth (e.g., a depth
where approximately 1% of surface insolation occurs), then phytoplankton spend too much time in an
inadequate light environment and net primary production is limited (Figure B-l) (Huisman et al. 1999).
The compensation depth is where water column phytoplankton photosynthesis and respiration are in
balance, and this often approximates the 1% insolation depth, or about two times the Secchi disc depth
(Parsons and Takahashi 1973). For example, in the lower Delaware Bay the upper-mixed layer often
corresponds to the bay bottom, the result of which is that phytoplankton spend too much time below the
compensation depth and hence low biomass production occurs (Pennock 1985). In systems where the
dominant support of the food web is derived from photosynthesis, net phytoplankton production must be
large enough to support the microbial loop (Azam et al. 1983) and higher trophic levels.
Nutrient Criteria—Estuarine and Coastal Waters B-1
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Production and respiration
Q.
0)
Q
a e
^Phytoplankton production
^ day'1)
Compensation depth
• Phytoplankton respiration
(m* day-1)
Limit of mixed layer
Figure B-l. Diagram illustrating theoretical distribution of
phytoplankton production and phytoplankton respiration. After
Sverdrup (1953). Source: Mann and Lazier (1996).
In a homogeneous medium, light decreases exponentially with depth and can be represented by the
negative exponential equation:
~- kz
where Iz is light quantity at depth z, I0 is the light quantity at the water surface, and k represents the
vertical light extinction coefficient; the extinction coefficient is more easily calculated as the base Iog10:
0.434 kz = log I0 - log Iz). The light gradient also often extends longitudinally down estuaries, especially
those dominated by large volumes of sediment-bearing freshwater. In highly turbid estuaries, deepest
light penetration shifts toward the orange end of the spectrum (Champ et al. 1980). The euphotic zone
depth generally increases from the tidal head to the coast. Where turbidity is at its maximum level, a
localized sharp decrease in euphotic zone depth is typical (Flemer 1970; Pennock and Sharp 1986).
Regions with the greatest turbidity typically are light-limited or almost so. In the turbid upper
Chesapeake Bay, riverine loading supplied nearly 90% of the particulate organic carbon, but in the
clearer waters of the mid-bay primary production dominated supply at 97% (Biggs and Flemer 1972);
Smith and Kemp (1995) have updated their original estimate and suggest a lower percentage. However,
much of the allochthonous organic matter may not be biologically available. In waters of high humic
material content (e.g., Charlotte Harbor, FL), light attenuation can be severe (Dixon and Kirkpatrick
1999). The interaction between turbidity caused by humic materials and nonchlorophyll-bearing
B-2
Nutrient Criteria—Estuarine and Coastal Waters
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participates complicates the direct application of the Secchi disc as a measure of nutrient overenrichment.
Spectral radiometers can to some extent partition the various components of light extinction and are the
preferred tools. Modern algal pigment diagnostic tools (e.g., HPLC) can compare water quality
responses to varying nutrient and other pollutant inputs across various coastal system types (Jeffery et al.
1997). It is possible that turbidity may mask the impending development of undesirable algae.
The vertical extinction coefficient for estuaries shows wide seasonal variations. Values frequently
exceed 0.1 m"1). The extinction in open-ocean waters is often estimated by the relation B.C. = 1.7/
Secchi depth in meters. Holmes (1970) and Keefe et al. (1976) both arrived independently at a constant
of approximately 1.46 instead of 1.7 for turbid estuarine and nearshore coastal waters. Walker (1980)
suggested that the original Poole and Adkins (1929) Secchi disc constant on average gives results
approximately 17% too high and suggested a value of 1.45. These corrections should be made and, more
importantly, it is useful for the constant to be checked for each estuary. For more quantitative work, a
quantum light meter that measures over a spectral range of 400-700 nM is preferred. Commercial
products (e.g., www.licor.com) are now available that can measure the spectral photon flux over a range
of interest to aquatic scientists.
The extinction coefficient can be broken down into several components. The total light attentuation, KT
= Kw + Kc + Kd + Kp (Lorenzen 1972; Kirk 1983; Bledsoe and Phlips 2000; Koenings and Edmondson
1991) can be resolved for the effects of water, chlorophyll a, dissolved substances, and nonalgal
particulate matter. In many estuaries, Kd may contribute between 5% and 50% of the KT. The Kw
usually can be ignored because it is such a minor component. In blackwater estuaries receiving high
loads of humic materials the Kd may dominant KT. In systems such as the "turbidity maximum zone" in
upper Chesapeake Bay, Kp may be the dominant component. The EPA Chesapeake Bay Program has
sponsored research to calibrate KT components applicable to SAV beds (www.chesapeakebay.net; search
the publications database for "Chesapeake Bay Submerged Aquatic Water Quality and Habitat-Based
Requirements and Restorations Goals: A Second Technical Synthesis."
Among some coastal ecosystems, light (i.e., mean photic depth) and nutrient loading appear to be equally
good predictors of phytoplankton primary production (see Figure Ib in Cloern 1999). This observation
strengthens the proposition that phytoplankton production in these systems can be limited by other
resources and processes in addition to nutrient loading. Pennock and Sharp (1994) suggest that the
Delaware River Estuary functions analogously as a chemostat during the summer. They point out that
high N supplies from upstream advect continually through the brackish-water region into the lower
estuary, where high primary production occurs from remineralized N and the advected supply and the
phytoplankton biomass is limited by grazing. Their conclusion is especially significant because evidence
also suggests that bioassay experiments that isolate the water may lead to misidentification of nutrient
limitation. The flushing component in the bay provides the physical analogue to a chemostat, where the
nutrient supply and adequate light support high phytoplankton biomass production and grazing and
flushing maintain a potentially steady-state phytoplankton biomass, with fluctuations due primarily to
physical forcing factors.
Nutrient Criteria—Estuarine and Coastal Waters B-3
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APPENDIX C
ADDITIONAL INFORMATION ON FLUSHING IN ESTUARIES
by Edward H. Dettmann, U.S. EPA, Office of Research and Development, National
Health and Environmental Effects Research Laboratory, Atlantic Ecology Division
A variety of terms such as residence time, flushing time, transit time, turnover time, and age are used to
describe time scales for transport and removal of materials that enter waterbodies. Use of these terms in
the literature is often inconsistent and sometimes imprecise, so that care must be exercised to determine
the meaning of terms being used, to avoid misinterpretation or incorrect comparisons of data. This
appendix contains definitions of terms and discussions of factors that affect flushing and empirical and
modeling methods for estimating residence times in estuaries. An understanding of residence times is
especially important when estimating not only system responses to nutrients, but also the lag phase
between management and system improvements.
Definitions of Residence Times
The freshwater residence time (ifw) is the average amount of time that freshwater, or a conservative tracer
introduced with freshwater inputs, resides in the estuary before exiting. It is the mean transit time for a
molecule of water or conservative tracer to progress from the point of input to the seaward boundary.
This is the definition of residence time most often given in the literature, and is generally the most useful
for analysis of eutrophication in estuaries, as most nutrients are usually introduced with freshwater.
Another commonly used concept is that of the mean amount of time required for water (or a
homogeneously-distributed conservative tracer) that is in the estuary at a given time (regardless of
source) to leave the estuary. This is here called "estuary residence time"(Te), although Takeoka (1984)
and Zimmerman (1976) use the term "residence time."
The values of ifw and ie may differ. Consider the case of an elongated unstratified estuary, with the
seaward boundary at one end and a river as the single source of freshwater located at the opposite end. In
this case, the travel distance from the river mouth to the seaward boundary is longer than the average
travel distance from other parts of the estuary, so that ifw > ie. Conversely, if the river enters the estuary
near the seaward boundary, ie may exceed ifw. For cases intermediate between these two, the difference
between these two residence times becomes smaller, and the times may be equal. Multiple freshwater
inputs and other factors such as stratification and estuary shape may complicate matters, but the same
principles apply.
These concepts, the related concepts of turnover time and age, and the relationships among these
measures are discussed in detail by Bolin and Rodhe (1973) and Zimmerman (1976), and summarized by
Takeoka (1984). Miller and McPherson (1991) discuss another concept, "pulse residence time," for water
or a conservative constituent introduced as a pulse, that may be useful in some circumstances.
Factors That Influence Residence Time
Water residence times in estuaries are influenced by any factor that affects water movement, including
freshwater inflow rates, tides, wind, mixing, stratification, and system topography. Because many of
these factors are variable, residence times are also not static. This variability requires attention to the
appropriate time interval over which the residence time should be expressed, and the representativeness
of the conditions under which a given measurement is made. A long-term (seasonal or annual) average
residence time is often most appropriate for analysis of the effects of nutrients.
Nutrient Criteria—Estuarine and Coastal Waters C-1
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Freshwater Forcing
Residence time may be quite sensitive to freshwater inflow rate, with larger flow rates associated with
smaller residence times. This is illustrated by three Rhode Island estuaries. For Narragansett Bay, ifw
varies between approximately 10 and 40 days; ifw at the long-term mean freshwater inflow rate is 26 days
(Pilson 1985). For the Seekonk River, ifw varies between 0.4 and 2.9 days, with a value of 1.2 days for
the mean freshwater inflow rate, and for the Providence River, ifw varies between 0.7 and 6.3 days, and is
2.5 days at average inflow (Asselin and Spaulding 1993).
Tidal Forcing
Tides can be a major factor controlling estuary-ocean exchange of water and therefore water residence
time. Important factors are the tidal range, tidal frequency (diurnal vs. semidiurnal), and estuary depth.
Tidal ranges in U.S. coastal waters range from centimeters to more than 5.5 m. For a given estuary,
residence times may vary over the spring-neap tide cycle.
Win d Forcing
Wind may substantially influence estuarine circulation, and therefore water residence time. In a study of
two small shallow estuaries, Geyer (1997) found that wind direction (offshore vs. onshore) had a
substantial effect on salinity structure and nontidal flow in one estuary, and ifw varied by a factor of
approximately 3 (from 0.8 to 2.7 days) in response to the sea-breeze cycle. Measurements in a nearby
and similarly oriented estuary during the same time period found that whereas wind direction influenced
salinity distributions, a constriction at the estuary mouth limited estuary-offshore exchange, so that there
was no significant relationship between residence time and wind stress. Estuaries bordering the Gulf of
Mexico are shallow, highly susceptible to wind forcing, and considered meteorologically dominated
(Solis and Powell 1999; Ward 1980).
Determination of Residence Time
A number of empirical and computational methods are used to measure or estimate water residence
times.
Empirical Measurements
Empirical measurements of water residence time depend on measurement of the distribution or dynamics
of tracers (generally freshwater or introduced dye) in estuaries.
Bolin and Rodhe (1973) show that the transit time of a tracer through a reservoir is given by the turnover
time (TO) , i.e.
where M0 is the total mass of a constituent in the reservoir and F0 is the total flux through the
reservoir. This is the basis for the freshwater replacement method for calculating the mean freshwater
residence time in an estuary. The ifw is calculated as
v=i
Ujw
where Vfw is the volume of freshwater in the estuary and Qfw is the input rate of freshwater. This ratio
gives the time required for the inflowing freshwater to replace the freshwater already in the estuary. The
volume of freshwater in the estuary is calculated as the amount of freshwater that must be mixed with
C-2 Nutrient Criteria—Estuarine and Coastal Waters
-------
seawater having salinity (Ss) equal to that entering at the seaward boundary to yield a volume Ve equal to
that of the estuary, with salinity equal to the mean salinity of the estuary (Se). Vfw may be calculated as
ve
Examples of this method are given by Pilson (1985), Geyer (1997), and Solis and Powell (1999).
Fluorescent dye (usually Rhodamine WT) is also used to determine water residence times. To measure
freshwater residence time, dye is introduced continuously into the inflowing freshwater at a rate
proportional to the freshwater flow rate. Dye concentrations are surveyed at (high or low) slack tide.
Dye input is terminated when the mean concentration of dye in the estuary reaches equilibrium
(approximately 3 times the freshwater residence time). Further periodic surveys are conducted at the
same tide phase to monitor the dye content of the estuary. The average concentration of dye is usually
found to follow a decreasing exponential with time,
C(0 = C e
-kt
f.
o
where C0 is the initial average concentration of dye in the estuary and C(t) is the concentration at time t.
This function is fit to the data, and the mean residence time of dye in the reservoir, a surrogate for the
mean freshwater residence time, is then ifw = 1/k. Alternatively, as the concentration one residence time
after termination of dye input is C = C0 e"1, the time required to attain this concentration is sometimes
taken as the residence time. The estuary residence time may be measured similarly. In this case, the dye
is distributed as uniformly as possible throughout the estuary in a rapid application, and the change in
dye content is monitored as above. The residence time determined from the rate of change of tracer
concentration is sometimes called the e-folding time. Because of cost and logistical considerations, dye
studies are most often done on relatively small estuaries. Examples of such studies are described by
Callaway (1981), Dettmann et al. (1989), and Geyer et al. (1997).
Models
A wide range of models, ranging from simple to complex, has been used to calculate water residence
time in estuaries. The simplest of these is the tidal prism model, which estimates residence time (in
number of tidal cycles) as the ratio V/P, where V is the estuary volume (usually expressed at high tide)
and P is the volume of the tidal prism. The model is based on the assumption that water entering the
estuary on the flood tide is thoroughly mixed throughout the estuary within a tidal cycle. Most estuaries
do not meet this requirement and, for all but very small estuaries, this model generally underestimates
estuary and freshwater residence time, sometimes by manyfold.
Ketchum (1951) modified this simple model by dividing the estuary into segments, each having a length
that corresponds to the local tidal excursion. Water exchanges between adjacent segments during each
tidal cycle are calculated, and complete mixing is assumed to occur only within each segment. The
segmented tidal prism model requires data for freshwater inputs, tidal range, and estuary topography. The
simple and segmented versions of the tidal prism model are reviewed by Pritchard (1952) and Dyer
(1973), and modified by Dyer and Taylor (1973). For most (but not all) estuaries to which the segmented
model has been applied, it has given good results (Dyer 1973). Beyond permitting calculation of
residence time, segmented tidal prism models permit calculation of the distributions of fresh- and
Nutrient Criteria—Estuarine and Coastal Waters C-3
-------
saltwater and other water quality constituents along the estuary, but do not address the effects of wind on
flushing.
Another approach to calculating residence times is the box model. Box models also segment the estuary,
assume complete mixing within segments, and calculate diffusive and advective exchanges between
adjacent segments. These models require data for estuary topography and for freshwater inflows and
salinity distribution in the estuary for each set of conditions to which the model is applied. As actual
salinity data are used, the effects of wind and tide are implicitly taken into account. Applications of such
models are described by Brown and Arellano (1980), Dettmann et al. (1992), Hagy et al. (2000), and
Miller and McPherson (1991). Such models also allow calculation of advective and diffusive exchanges
among segments and spatial distribution of water quality constituents.
Numerical computer models of hydrodynamics and constituent transport are still more complex. These
models are used to simulate movement of water and water quality constituents at fine spatial resolution,
and may also be used to calculate freshwater or estuary residence times. These models generally allow
consideration of wind stress as well as tidal and freshwater forcing. A few examples of such models are
described by Brooks et al. (1999), Signell (1992), and Signell and Butman (1992).
C-4 Nutrient Criteria—Estuarine and Coastal Waters
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APPENDIX D
NOAA SCHEME FOR DETERMINING ESTUARINE SUSCEPTIBILITY
(SOURCE: BRICKER ET AL. 1999)
The following provides an overview of the NOAA scheme for determining the overall human potential
for causing nutrient enrichment of estuaries.
Dilution and Flushing Potential
The length of time that nutrients spend in an estuary potentially affects their opportunity to contribute to
overenrichment. The time is a function of dilution potential and flushing rate. The analysis uses
physical and hydrologic data to define separately (1) a dilution rating and (2) a flushing rating. In both
cases, the higher rating, the greater the capacity to dilute or flush nutrient loads (see Tables D-l and
D-2).
Figure D-l combines dilution potential and flushing potential. By combining dilution and flushing
components, an export potential (EXP) is determined. Estuaries in the upper left portion of the matrix
generally have a high EXP that suggests an ability to dilute and flush nutrient loads. Estuaries in the
lower right portion of the matrix have the opposite capacity, making them more susceptible to nutrient
input.
Nutrient Inputs
NOAA used the USGS Sparrow Model (spatially referenced regressions of contaminant transport on
watershed attributes) and other information to estimate nutrient loads and measure nitrogen pressure on
estuaries.
Determination of Overall Human-based Nutrient Pressure
A matrix was used to compare the susceptibility to nutrient retention and the level of N inputs to rank the
overall expression of human influence on eutrophic conditions in the estuary (Figure D-l). Experts at the
National Assessment Workshop reviewed and, where appropriate, modified the assessments based on
higher quality data available for some estuaries; expert knowledge also played a role.
Nutrient Criteria—Estuarine and Coastal Waters D-1
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Table D-l. Decision rules for dilution potential
Type
A
B
C
IF: vertical stratification
Vertically homogenous
•all year
•throughout estuary
Minor vertical stratification
•navigation channels
•upper estuary
Vertically stratified
•most of year
•most of estuary
THEN: dilution volume
I/ VU.Lestualy
l/VOLestualy
1/VOLM
(fwf=freshwater fraction)
IF: dilution value
icr13
icr12
1CT11
icr10
JQ-09
Dilution potential
High
Moderate
Low
No. estuaries
30
63
45
Note: This analysis assumes that a larger portion of the water column is potentially available to dilute nutrient loads in a
vertically homogeneous estuary than in a vertically stratified system. The assumption is that for stratified systems, nutrients are
most often retained in the upper portion (freshwater fraction) of the water column. In contrast, downward transport (more
complete mixing) is likely in vertically homogeneous systems. Type B estuaries are generally vertically homogeneous, although
stratification is observed (confined) in narrow navigation channels or the extreme upper reaches of an estuary. In this case,
nutrients are assumed to be diluted throughout the entire water column.
Source: Bricker et al. 1999.
Table D-2. Decision rules for flushing potential
Type
1
2
3
4
5
6
7
8
Tide range (ft) Freshwater inflow/estuary volume
Macro (>6) and Large or moderate (10°° to 1CT02)
Macro (>6) and Small (1CT03, 1CT04)
Meso(>2.5) and Large (10°°, 1CT01)
Meso(>2.5) and Moderate (1CT02)
Meso(>2.5) and Small (1CT03, 1CT04)
Micro (<2.5) and Large (10°°, 1CT01)
Micro (<2. 5) and Moderate (1CT02)
Micro (<2.5) and Small (1CT03, 1CT04)
Flushing potential
High
Moderate
High
Moderate
Low
High
Moderate
Low
No. estuaries
12
21
15
16
26
4
13
31
Note: This analysis assumes that a greater capacity to flush nutrient loads exists for estuaries that have large tide and freshwater
influences.
Source: Bricker et al. 1999.
D-2
Nutrient Criteria—Estuarine and Coastal Waters
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ESTUARINE EXPORT POTENTIAL AND SUSCEPTIBILITY
DILUTION Potential
High Moderate
Low
o
o
Q_
O
X
CO
13
Low
Susceptibility
Low
Susceptibility
Moderate
Susceptibility
Low
Susceptibility
Moderate
Susceptibility
High
Susceptibility
Moderate
Susceptibility
High
Susceptibility
High
Susceptibility
HIGH EXP. Estuary has capacity to dilute
and flush nutrients
MODERATE EXP. Estuary has capacity
to either dilute or flush nutrients
LOW EXP. Estuary does not have capacity to
dilute or flush nutrients
OVERALL LEVEL OF HUMAN INFLUENCE
MODERATE
'. Even low nutrient additions
I may result in problem
> symptoms in these
* estuaries.
MODERATE LOW
Symptoms observed in the
estuary are minimally to
moderately related to
nutrient inputs.
LOW
Symptoms observed in the
estuary are likely
predominantly naturally
related or caused by human
factors other than nutrient
additions.
MODERATE HIGH
Symptoms observed in the
estuary are moderately to
highly related to nutrient
additions.
MODERATE
'. Symptoms observed in the
<- estuary are moderately
' related to nutrient inputs.
LOW
Symptoms observed in the
estuary are predominantly
naturally related or caused
by factors other than nutrient
additions.
HIGH
Symptoms observed in the
estuary are probably closely
related to nutrient additions.
MODERATE HIGH
Symptoms observed in the
estuary are moderately to
highly related to nutrient
additions.
MODERATE LOW
Symptoms observed in the
estuary may be naturally
related or the high level of
nutrient additions may cause
problems despite low
susceptibility.
Low Nutrient Input
Moderate Nutrient Input
High Nutrient Input
Figure D-l. Estuarine Export Potential Susceptibility.
Nutrient Criteria—Estuarine and Coastal Waters
D-3
-------
APPENDIX E
COMPARATIVE SYSTEMS EMPIRICAL MODELING APPROACH:
THE EMPIRICAL REGRESSION METHOD TO DETERMINE NUTRIENT
LOAD-ECOLOGICAL RESPONSE RELATIONSHIPS FOR
ESTUARINE AND COASTAL WATERS
By James Lattimer, USEPA, Office of Research and Development, National
Health and Environmental Effects Research Laboratory, Atlantic Ecology Division
The empirical regression method can be used to determine the response of estuarine systems to nutrient
loading. This method can take two forms: single system and comparative-systems approaches.
In single system approaches, ecological responses and nutrient concentrations are measured over time in
a single system. This allows for the development of load-response models for the individual systems that
are robust because many of the controlling factors, such as, for example, physiographic setting, primary
productivity base, and hypsography, are relatively constant. This approach has been used to develop
models of the response of primary productivity to nitrogen load (Figures E-la,b) (Boynton et al. 1995).
However, this approach is only applicable to the system in which it was developed and thus is not
considered to be very useful in the development of widely applicable nutrient load-response
relationships. Single system empirical studies, however, may be useful in providing data on processes
useful for numerical model development.
The alternative approach, using the space for time paradigm (Picket,1988), posits that relationships
between nutrient inputs and ecologically meaningful estuarine responses, using multiple systems, have
predictive capability, at least for the category of systems used in the model development. This allows for
a wide range in nutrient loading and estuarine types to be included. The comparative-systems empirical
approach has been used to determine, for example, relationships between nutrient inputs and fish yields
(Lee and Jones 1981; Nixon 1992), benthic biomass, production and abundances (Josefson and
Rasmussen 2000), summer ammonia flux (Boynton et al., 1995), chlorophyll-a concentration (Boynton et
al. 1996; Boynton and Kemp 2000; Monbet 1992), primary productivity (Nixon etal, 1996), and the
dominant source of primary productivity (Nixon et al. in press). In many of these cases, important
environmental factors such as flushing time and depth, are used to normalize the nutrient loading in a
similar way as Vollenweider (Vollenweider 1976) to yield more precise relationships.
The comparative-systems empirical approach has been successfully used in a regulatory framework to
develop total maximum loads for 30 subestuaries within Buzzards Bay in Massachusetts (Costa et al.
1999). Using a citizens' monitoring network, many important water quality variables were measured
during summer sampling periods over a five to seven year period. Nitrogen inputs were estimated from a
land-use model modified from Valiela (Valiela et al., 1997) augmented with literature data on point
source and atmospheric inputs. In this study, nitrogen load-response relationships were derived for
nitrogen concentration, chlorophyll-a, secchi depth, dissolved oxygen concentration, and eelgrass habitat
ratio. Nitrogen loadings were normalized to account for volume and flushing time of each of the systems
to improve the precision of the empirical models. An example of the types of relationships determined
using this approach are given in Figure E-2. The regional estuarine management program used this
method to adopt total maximum annual loads (TMALs) for nitrogen. Specifically, "...the nitrogen
management strategy represents a linking of estimates of watershed nitrogen loading...to empirical
observations of ecosystem response in a wide variety of Buzzards Bay embayments" (Costa 2000).
The comparative-systems empirical approach does not explicitly consider the processes that produce the
observed phenomena; however, factors known to affect behavior are determined in order to reduce the
Nutrient Criteria—Estuarine and Coastal Waters E-1
-------
uncertainty in the models. The smaller the variance in the load-response relationship, the more
compelling the association. One important question for management is to determine what level of
variance is sufficient to convince stakeholders to accept nutrient limits and associated monetary
expenditures.
This approach is largely based on statistical associations and is therefore restricted to prediction within
the class of systems used in the model development. Applicability improves, however, by the inclusion
of systems that encompass a wide range in loading and responses. The comparative-systems empirical
approach allows for the direct measurement of important endpoints (e.g., hypoxia, SAV loss, biomass)
obtained in the environment. The endpoints are what are important to the general public. So by
providing a mathematical relationships between the stressor and important endpoints, managers can
convince the public of the importance of regulatory action.
CM
Alt. Eutrophication Index vs Loading
4 yr mean
^ 100
o ™ •
-5 90 -
X 80
_g 70 -
_C 60
C 50 :
.2 40 '
+-> .
o 30 "
IE 20 -
§- 10-
is n
3 0
•
ff
I
1 i
i 1
T ' 1 i
|
f ,
|
i
i
\
\
• -i
,\
HI 1 10 100 1000 100
mg N Am3 AVr
Figure E-la. Scatter plots showing correlation between nitrogen loading, expressed
using the volume Vollenweider-term flushing scale, and 92-98 mean +/- std. errors of the
Alternate Eutrophication Index scoring (without oxygen scores).
E-2
Nutrient Criteria—Estuarine and Coastal Waters
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Eelgrass habitat cover vs. Loadings
1.6 i
1.4
o
ID 1.2
1 1.0 -
|o.8
1 O-6
1 0.4-
LLJ
0.2
n n
Mgh
F
Quh
Waquoit Bay, _-— •
1951,1971,1985,
1 WiH
PoH
JhH •
RBH
wSf%
990 \
^
As
wt;
:H
t
«»WaRAppl
n »
» / A,
5ApB y
> »nR vV®K
• n*
nogansett, 1971
ponogansett, 1984
SIR
1 10 100 1000 10000
N loading (mg/mA3 per Vr)
Figure E-lb. Ratio of eelgrass habitat area to potential habitat area versus nitrogen
loading, expressed using the volume Vollenweider-term flushing scale (from Costa et al.
1999).
Nutrient Criteria—Estuarine and Coastal Waters
E-3
-------
Primary Production vs N-Load
1200
1000
800
CJI
£
Z
S
c"
.
§ §00
«
I 400
200
Malnstem Bay
(1971-1976)
(1985-1990)
(A)
y= -
f2 = 0.54
100
200
300
400
Total Nitrogin Losd, |iM N m'z h'1
(average of year+year-1)
Figure E-2. Plot of annual TN loading rates versus phytoplankton primary production rates at a
single station in the Chesapeake Bay from 1971-76 and 1985-90. Source: Boynton et al., 1995.
E-4
Nutrient Criteria—Estuarine and Coastal Waters
-------
APPENDIX F
SELECTED THEORETICAL APPROACHES TO CLASSIFICATION OF
ESTUARIES AND COASTAL WATERS
Several theoretical schemes have been presented that may facilitate classification of estuaries and coastal
waters if the need arises for more theoretical approaches. In most cases, these schemes have potential
future value but are not likely to be immediately useful. The information-thermodynamic approach
offered by Ulanowicz would seem to provide useful insights to development of an eutrophication index
and may be worth an earlier consideration.
Functional Attributes
Odum and Copeland (1974) proposed a classification scheme based on the idea that an ecosystem is a
balance between energies that build structure and order, or ordering energies, and energies that cause
loss of structure and order, or disordering energies. Although this approach is of theoretical interest, the
data to apply it are still largely unavailable. There are conceptual difficulties. Energy sources and
stresses are not always mutually exclusive categories. Day et al. (1989) give several examples of cases
where moving water can be either an energy source or a stress, depending on the situation. Moderate
currents are a source of energy for seagrass meadows, because they transport organic matter and
inorganic nutrients to beds and remove metabolic wastes. If currents are too strong, however, they can
disrupt beds and act as disordering energies.
Theoretic Indices
This approach is data intensive and has potential as appropriate estuary data are generated. Ulanowicz
(1986, 1997) described an approach based on flow analysis and information theory. The idea is that an
ecosystem can be characterized in terms of growth and organization. Growth is defined as an increase in
system activity or total system throughput (analogous to total system energy flow). Organization is
equated with the mutual information inherent in the trophic flow structure. To apply this approach, one
would need to obtain energy flow values (or organic carbon-based equivalents) among trophic
compartments. Following Ulanowicz's formulation, one could develop an information index of
eutrophication. Compiling such would entail an assessment of the growth and organization status of a
eutrophic or nutrient-enriched system compared with a current reference system, a minimally impaired
system, or an estimate of pre-eutrophication values from historical data. The ratio of growth and
organization of a nutrient-enriched system to its reference value would express the degree of impairment.
Such information could help guide restoration priorities. The downside is that the approach requires the
availability of a rich ecological database.
Nutrient Criteria—Estuarine and Coastal Waters F-1
-------
APPENDIX G
EXAMPLES OF NUTRIENT CONCENTRATION RANGES AND RELATED
HYDROGRAPHIC DATA FOR SELECTED ESTUARIES AND COASTAL WATERS IN
THE CONTIGUOUS STATES OF THE UNITED STATES
Atlantic Coast Systems
BOSTON HARBOR, MA (Source: Kelly 1998)
WAQUOIT BAY, MA (Source: Valiela et al. 1992)
SEEKONK-PROVIDENCE RIVER REGION OF NARRAGANSETT BAY, RI
(Source: Doeringetal. 1990)
WESTERN LONG ISLAND SOUND AND HUDSON-RARITAN ESTUARY, NY/NJ
(Source: O'Shea and Brosnan 2000)
HUDSON RIVER, NY (Source: Lampman et al. 1999)
DELAWARE RIVER ESTUARY (Source: Lebo and Sharp 1993)
MARYLAND COASTAL BAYS (Source: Boynton et al. 1996)
CHESAPEAKE BAY, MD/VA (Source: Magnien et al. 1992)
CHESAPEAKE BAY (Harding and Perry 1997)
YORK RIVER ESTUARY, CHESAPEAKE BAY, VA (Source: Sin et al. 1999)
NEUSE RIVER ESTUARY, NC (Source: Rudek et al. 1991)
CAPE FEAR RIVER ESTUARY, NC (Source: Mallin et al. 1999)
COASTAL GEORGIA (Source: Hopkinson and Wetzel 1982)
Gulf of Mexico Estuaries
GULF OF MEXICO CASE STUDIES (Source: Bianchi et al. 1999)
GALVESTON BAY, TX (Source: Santschi 1995)
Pacific Coast Systems
EMBAYMENT OF PUGET SOUND, WA (Source: Bernhard and Peele 1997)
Nutrient Criteria—Estuarine and Coastal Waters G-1
-------
Boston Harbor
42.45-
42.40.-
42.35-
42.30-
42.25-
42.20-
-71.10 -71.05 -71.00 -70.95 -70.90 -70.85 -70.80 -70.75
The study area (N latitude and W longitude) in Boston Harbor
and western Massachusetts Bay. The boundary between the
Harbor and the Bay is defined by the solid line from Deer
Island to Hull combined with the dashed lines showing the 2
high-resolution transects, the spatial limits of data for box
modeling are depicted. Dots position some watercolumn
monitoring stations that were sampled during 1994. Stations
prefixed with 'N' surround the future offshore outfall, which is
centered between Stns N20P and N16P about 15 km from
Deer Island. Data from 2 lines of Bay staions (NO IP to N10P
and N04P to N07P, each representing about 10 km distance)
were used to provide approximate concentrations for the future
tidal source region for the Harbor.
G-2
Nutrient Criteria—Estuarine and Coastal Waters
-------
Survey Transect
W9402 North
South
W9403 North
South
W9404 North
South
W9405 North
South
W9407 North
South
W9409 North
South
W9411 North
South
W9412 North
South
W9413 North
South
"From high-resolution,
Bay C0 (to
BA" NH4
(nr1) (uM)
2.23
1.89
1.23
1.07
0.85
1.11
1.53
1.46
1.55
nd
2.04
1.50
1.51
1.78
2.13
nd
1.94
1.86
3
2
1
0.75
2
1
1.5
1.5
1.5
nd
5
5
7
4
6
6
2
2
in situ sampling.
calculate ocean loading)
DIN TN PO4 SiO4
(pM) (uM) (uM) (uM)
10.5
9.5
2
1.75
3
2
3
3
2
nd
6
6
11
7
8
8
3.5
3.5
20
18
nd
nd
16
14
nd
nd
14
nd
nd
nd
21
16
nd
nd
nd
nd
0.8
0.8
0.3
0.3
0.3
0.25
0.5
0.5
0.7
nd
1
1
1.1
1
1
1
0.5
0.5
8
8
1
0.75
1.5
1
2.75
2.75
1.5
nd
4.5
4.5
6
5
3
3
2.75
2.75
Converted to TSS as described
Harbor Q
BA" NH4
(nr>) (uM)
2.34
1.78
1.61
1.32
1.29
1.30
1.84
2.00
2.15
nd
3.26
1.64
2.07
1.90
2.66
nd
2.06
2.06
in the text
4
3.5
nd
nd
3
2.5
nd
nd
2.25
2
2.1
1.4
10
11
2.25
nd
15
18
(to calculate Harbor output)
DIN TN PO4 SiO4
(uM) (uM) (pM) (uM)
14
12.5
nd
nd
4.5
3.25
nd
nd
2.6
3
nd
nd
15
16
nd
nd
20.5
20.5
28.5
19
nd
nd
25
18
13
11
17.5
14.5
11
10.5
26
24
20
2.1
29.7
29.7
0.8 12
1 9.5
nd nd
nd nd
0.3 3
0.25 1.4
0.5 nd
0.8 nd
0.6 1.5
0.8 2.5
0.7 nd
0.55 nd
1.3 8
1.3 7
1.5 nd
1.5 nd
1.7 nd
1.7 nd
Bay (C0) and Harbor (Q) concentration data used in calculating Harbor-Bay exchange. BA: beam attenuation from
transmissometer readings; nd: not determined.
Nutrient Criteria—Estuarine and Coastal Waters
G-3
-------
Waquoit Bay
co
£ 6
CM
2
Light
Jun
Jul Aug
1988
Sep
Time course of light, temperature,
oxygen, ammonium, phosphate, and
chlorophyll in water of Waquoit Bay,
summer 1988. Modified from Costa et al.
(in press).
G-4
Nutrient Criteria—Estuarine and Coastal Waters
-------
O2 Cmg P)
o
0,2
0,8
at
0
i r~~T
- after 5
cloudy
days
Vertical profiles of oxygen in Childs River during dawn and midafternoon
of a sunny day (circles), and during afternoon after three cloudy days in a
row. Data from C. D'Avanzo.
Nutrient Criteria—Estuarine and Coastal Waters
G-5
-------
Seekonk—Providence River Region,
Narragansett Bay
Station locations, Solid lines delimit boxes used
in modeling effort. BR=Blackstone River, BV=
Blackstone Valley Sewage Treatment Plant
(STP), TM= Ten Mile River, MR Moshassuck
River, WR=Woonasquatucket River, FP=Field's
Point STP, EP=East Providence STP, PR=
Pawtuxet River
G-6
Nutrient Criteria—Estuarine and Coastal Waters
-------
SURFACE SALINITY
30-
24-
18-
L
L12-
6-
o-
0
SURFACE AMMONIA
80-1
10 15 20 25
SURFACE NITRATE + NITRITE
140-'
120-
T
10 15 20
SURFACE PHOSPHATE
25
5 10 15
SURFACE SILICATE
20
25
12-1
9-
o
* 3-
o-
X
— -.
H
^
0 5 10 15 20 2.
120-
^ 80-
cn
0)
0
140-
o-
5
,
1
\
X-\ Jl ,
^HI^
0 5 10 15 20 2f
SURFACE PARTICULATE NITROGEN
SURFACE PARTICULATE CARBON
0.5-
0.4-
-0.3-
01
E0.2-
0.1-
1-
5 10 15 20
DISTANCE KM
25
5 10 15 20
DISTANCE KM
25
Mean and range of salinity and nutrient concentrations observed in
surface waters (depth=1.0m) during the six cruises, versus distance
from the Main Street Bridge, Pawtucket, Rhode Island at the head
of the Seekonk River.
Nutrient Criteria—Estuarine and Coastal Waters
G-7
-------
BOTTOM SALINITY
24-
"5.16-
BOTTOM AMMONIA
10 15 20 25
BOTTOM NITRATE + NITRITE
80
60-
^401
^20H
o~V
o
30-
20-
cH
5 10 15 20
BOTTOM PHOSPHATE
25
10
15
20
25
BOTTOM SILICATE
*'-
"^-^
V
-
n
^
80-
60-
^ 40-
o
^20-
n-
\
^
1 1
10
15
20
BOTTOM PARTICULATE NITROGEN
25 0 5 10 15 20
BOTTOM PARTICULATE CARBON
25
0.5-
0.4-
-0.3-
cn
£0.2-
0.1
o.o-
-
'*^
•^4-
D 5 10 15 20
2
3-
2-
cn
E
o-
3
x
\\
\liL___l
D 5 10 15 20 2f
DISTANCE KM
DISTANCE KM
Mean and range of salinity and nutrient concentrations observed in bottom
waters (1.0 m from bottom) during the six cruises, versus distance from the
Main Street Bridge, Pawtucket, Rhode Island at the head of the Seekonk
River.
G-8
Nutrient Criteria—Estuarine and Coastal Waters
-------
Western Long Island Sound
Hudson-Raritan Estuary
LOCATION OF WPCP's AND SAMPLING STATIONS IN NY HARBOR
m
m
m
ra
CD
co
m
NYC WPCPs and Dewatering
Facilities (in Bold)
H North River (6.18 nvs-1)*
Wards Island (9.81 m-S')
Bowery Bay (5.48 m»s-<)
Hunts Point (5.61 nrs')
Tallman Island (2.58 m> s-<)
Newtown Creek (11.65 nv s-1)
Red Hook (1.62 nvs1)
Owls Head (5.08nvs1)
Coney Island (4.64 nvs1)
26th Ward (2.89 nvs')
Jamaica (3.59 m> s-1)
Rockaway (0.87 rrvs-1)
Port Richmond (1.79 m3 s')
Oak wood Beach (1.27 rrrs-1)
Flows represent 1/97 - 12/97
daily averages (total flows)
Other WPCP's >0.44 m's'.
Approximate outfall locations
WESTCHESTER
COUNTS
Hudson River
NEW JERSEY
Newark Bay —
Arthur Kill
ATLANTIC OCEAN
Location of NYC-DEP
Harbor Survey monitoring
stations
Location of New York City Department of Environment Protection's Harbor Survey Water Quality
Monitoring Stations. Also depicted are the 14 New York City Water Pollution Control Plants and the !
New York City sludge dewatering facilities (shaded boxes). Sandy Hook, lower center, is located at
approximately 40°30'N, 74°W.
Nutrient Criteria—Estuarine and Coastal Waters
G-9
-------
WATER QUALITY INDICATORS FOR SUMMER 1999
Bottom Minimum
Dissolved Oxygen
Raritan
Bay
1.20-2.99
3.00-3.99
4.00 - 4.99
.'-I- 5.00 - 5.71
O "Of
^^ measured
Total Dissolved
Inorganic Nitrogen
NYSDEC STDS: DO > 3.0 mg L-' = SD (fish survival);
DO > 4.0 mg L-1 = I (fishing); DO > 5.0 mg L-1 = SB (bathing)
Secchi
Transparency
Key water quality indicators for summer (June-September) of 1999. Depicted are: bottom minimum
dissolved oxygen (upper left) in mg I"1; summer average dissolved inorganic nitrogen
[NH3-N+(NO3+NO2)-N] (upper right) in mg I"1; summer average Secchi transparency (lower left) in m;
summer average total phosphorus (lower right) in mg I"1.
G-10
Nutrient Criteria—Estuarine and Coastal Waters
-------
Dissolved Inorganic Nitrogen Trends
Summer Means, 1989-1999
Station E9
Stepping Stones, Long Island Sound
Station E9
Stepping Stones, Long Island Sound
y y=-17.4 ug NH4-N L '1yr'1*x + 34fl10
*_J
f
-P 0.4
*"^
o
z
g" 0.0
n ! I i 5 n n i
1989 1991 1993 1995 1997 1999
Station J3
Canarsie, Jamaica Bay
o>
\ 0.4
1989 1991 1993 1995 1997 1999
Station K5A
Raritan River, Raritan Bay
'-1 0.8
01
^ 0.4
X
Z
n n
i
y»-21
, — I -f .
1
1 ug
NH4
1
-N L "\r'1'x+42580
^ -L
I ir T
1989 1991 1993 1995 1997 1999
Year
1989 1991 1993 1995 1997 1999
Station J3
Canarsie, Jamaica Bay
_§
jL °-4
cT
y«12.0 ug NO,j-N L yr *x-23640
1989 1991 1993 1995 1997 1999
Station K5A
Raritan River, Raritan Bay
'_!
CD
"-1
5. °-4
cT
O
— n n
' M '
"*- _L
T
1
A
r
-inl:
1989 1991 1993 1995 1997 1999
Year
Whisker=Mean±SD
1989-1999 summertime (June-September) average ambient dissolved inorganic nitrogen [dissolved
ammonium-nitrogen (NH4-N) and dissolved nitrate- and nitrite-nitrogen (NO3)-N] concentrations (mg I"1 for
western Long Island Sound, Raritan Bay, and Jamaica Bay stations. Trends in summertime average
concentrations are noted where significant (p < 0.05).
Nutrient Criteria—Estuarine and Coastal Waters
G-11
-------
Phosphorus Trends
Summer Means, 1989-1999
Station E9
Stepping Stone, Long Island Sound
o>
0.4
0.3
0.2
0.1
5 0.0
1989 1991 1993 1995 1997 1999
Station J3
Canarsie, Jamaica Bay
'-> 0.4
D>
E
M 0.3
3
t_
O
J= 0.2
a
0
f °-1
ra
T T
1 I £ _j
— x — T
1 ,
: -i — "
I
r— "~;
1 _J
-1
'. —
—
'
• y»9.5 ug TP L"1 yr"1*n-18640
Q 0.0
1- 1989 1991 1993 1995 1997 1999
Station E9
Stepping Stone, Long Island Sound
0.2
o
O.
0.0
1989 1991 1993 1995 1997 1999
Station J3
Canarsie, Jamaica Bay
0.2
0)
_E
a. 0.1
O*
a.
0.0
y=5.3 ug PO,-P L yr 'x-10510
1989 1991 1993 1995 1997 1999
Station K5A
Raritan River, Raritan Bay
Station K5A
Raritan River, Raritan Bay
-1 0.4
o>
E
Phosphorus (
ppp
1* K) W
n
** n ft
jj I , ]
«-,
OJ
E
a. 0.1
cT
Q.
n n
l|i!
[
'
II ;
Q U.U w.w ~ . • • — •
hi 1989 1991 1993 1995 1997 1999 1989 1991 1993 1995 1997 1999
Year Year
Whisker=Mean±SD
1989-1999 summertime (June-September) average ambient total phosphorus (TP) and dissolved
orthophoshpate (PO4-P) concentrations (mg I"1) for western Long Island Sound, Raritan Bay, and Jamaica
Bay stations. Trends in summertime average concentrations are noted where significant (p < 0.05).
G-12
Nutrient Criteria—Estuarine and Coastal Waters
-------
Hudson River, NY
RJ
ingston
/c
oughkeepsie
N.Y.C.
KmO)
Map of the tidal freshwater Hudson River, its watershed
and major tributaries. The inset shows the location of the
watershed of the watershed of the Hudson, primarily in
eastern New York State. The main panel shows the
watershed of the Hudson. The upper (non-tidal) tributaries
are in white and their surrounding watershed is shaded.
The remaining, unshaded, section of the watershed
delivers water to the tidal Hudson River (heavy black line)
primarily through 7 tributaries. These tributaries (creeks)
are labeled: K for Kinderhook; Cfor Catskill; RJ for
Roeloff Janson Kill; E for Esopis; R/W for
Rondout/Wallkill; WP for Wappingers; Cr for Croton.
The first 5 of these tributaries enter within the budgeting
reach (labeled) between km 225 and 125. The section of
the river that was sampled during our transect sampling
(labeled sampling stretch) is also shown.
Nutrient Criteria—Estuarine and Coastal Waters
G-13
-------
80-
2
3 60-
c
OJ
o>
O 40-
20-
1
^p*
T
n = 1130
-3
8
0
-T
NH4 NOa TN PO4 TP
Box and whisker plots for forms of N and P for the
entire data set of the Hudson. The data include all
samples taken from January 1992 to December 1996
and combine seasonal and spatial variation. Shown are
the medians, upper and lower quartiles and 90%
inclusion lines. The number of samples for each
analysis is labeled near the boxes.
G-14
Nutrient Criteria—Estuarine and Coastal Waters
-------
TJ
3
12
6
240
Albany
180 120
River km
river flow
Spatial variation in selected variables for a
representative transect taken in early September
1996. Transects run km 240 (Albany
downstream to km 40. For each variable
samples were taken every 2 to 4 km. Units are,
conductivity—(note log scale); pH—normal pH
units; chlorophyll a— I'1; turbidity— NTU; N
and P—uM. For both N and P the per line
represents total N and toal P. "Other N" is TN
minus DIN (NH4 plus NO3); For P, "Other P" is
TP minus phosphate.
Nutrient Criteria—Estuarine and Coastal Waters
G-15
-------
Delaware River Estuary
200
The Delaware Estuary with insert showing relation to
the mid-Atlantic coast of the United States. Stations
(1-9, 11-26) were located along the main axis of the
estuary, every 5-15 km. Station 10 was located in the
Schuylkill River (not shown). The river is tidal from
the mouth of the bay up to the fall line at Trenton,
New Jersey (240 km). Salinity increases from 0.1%
at station 14 (130 km) to 30% at the bay mouth. The
majority of TP inputs to the river are clustered
between stations 6-9.
G-16
Nutrient Criteria—Estuarine and Coastal Waters
-------
Parameter
Distance
Salinity
PH
Dissolved oxygen
Nitrate
Total nitrogen
Dissolved organic
carbon
Paniculate carbon
Secchi depth
Phytoplankton
production
Units
km
ppt
NBS Scale
Percent of
saturation
MM
MM
MM
MM
cm
mniol C
m-2d-'
2
197
<0.1
7.1
87
74
98
253
36
129
54
7
161
<0.1
6.8
73
88
126
283
39
124
40
14
127
0.1
6.9
76
121
163
287
52
79
23
16
101
1.5
7.1
82
122
169
285
138
33
12
20
66
11.2
7.7
96
72
94
277
69
64
47
23
39
20.4
8.0
105
29
41
208
54
120
102
RSD
4
16
35
27
25
47
36
119
Figure 16Average biological, chemical, and physical parameters for Delaware Estuary 1986-1988. Values
are the nonweighted mean of 17 to 24 discrete samples taken at each location (Lebo et al. 1990). Data are
shown for locations near the Delaware River (station 2), Philadelphia (station 7), beginning of the salinity
gradient (station 14), turbidity maximum (station 16), downstream of the turbidity maximum (station 20), and
in the lower bay (station 23). In addition, average relative standard deviation (RSD) for all stations is shown.
The average standard deviation for station location and salinity was 1.1 km and 2.1%, respectively.
Nutrient Criteria—Estuarine and Coastal Waters G-17
-------
Maryland Coastal Bays
N
15km
Assawoman
Bay
V
Upper St. Martin
River
Bishopville Creek
Lower St. Martin:
River
Turville Creek
Trappe Creek
Newport Bay
Fenwick
Island
Isle of Wight
Bay
- Ocean City, MD
Sinepuxent Bay
Atlantic
Ocean
North Chincoteague
Bay
• Assateague Island
South Chincoteague
Bay
Map of the Maryland coastal bays complex indicating the boundaries of the watershed and a
subsystems for which nitrogen inputs were estimated.
G-18
Nutrient Criteria—Estuarine and Coastal Waters
-------
Q_
O
O
St. Martin Turville Isle of Wight Assawoman
River Creek Bay Bay
80
70
lT 60
I 50
40
30
20
10
0
Lower Bays
[2 1975
m 1983
Trappe
Creek
Newport Chincoteague Sinepuxent
Bay Bay Bay
Summer average chlorophyll a concentrations for representative regions of the Maryland coastal bays
based on samples collected during 1975, 1983, and 1991. Data are from Fang et al. (1977a,b), Maryland
Department of Health and Mental Hygiene (1985), and National Park Service (1991).
Nutrient Criteria—Estuarine and Coastal Waters
G-19
-------
Chesapeake Bay
Magnien et al.
Chesapeake Bay
Region
Map of Chesapeake Bay region showing tributaries, sampling sites, and major wastewater
treatment plants for systems examined or referenced in this paper. The major wastewater
treatment plants are as follows: 1) Western Branch, 2) Arlington, 3) Blue Plains, 4)
Alexandria, 5) Piscataway, 6) Little Hunting Creek, and 7) Lower Potomac.
G-20
Nutrient Criteria—Estuarine and Coastal Waters
-------
WINTER O SPRING * SUMMER ° FALL
Salinity dilution plots of nitrate plus nitrite,
ammonium, total nitrogen, orthophosphate, and total
phosphorus for surface samples at all Chesapeake
Bay Mainstream stations. Each point represents a
single station at which median values were calculated
for each seasonal time frame (see Fig. 2) over the
entire 195-1989 period. All stations identified in Fig.
1 were plotted in longitudinal order starting at the
head of the estuary. The circled point for each
seasonal plot is the upper mesohaline plankton station
at the Annapolis Bay Bridge.
Nutrient Criteria—Estuarine and Coastal Waters
G-21
-------
Concentrations of ammonium (uM) in Back River, and ammonium concentrations, primary production
estimates (mg C m2 d"1) and temperature (°C) for the Patapsco River(Baltimore Harbor). Values presented
are surface mixed layer medians by month from continuous monitoring at a frequency of one to two times per
month over the period 1985-1989. Station locations are centrally located in each system (see Fig. 1)
Month
January
February
March
April
May
June
July
August
September
October
November
December
Back River
NH4+
257.0
414.0
112.4
439.0
151.7
35.7
82.1
87.5
107.1
103.5
232.0
317.7
Patapsco River
NH4+
8.9
48.2
24.3
33.6
11.7
9.7
13.0
30.3
18.8
21.3
15.4
13.1
Primary Production
299
1,384
640
663
1,520
1,840
2,263
1,950
1,744
953
751
930
Temperature
3.2
3.0
6.2
12.3
16.8
24.3
27.2
26.6
23.8
17.7
11.3
6.5
G-22
Nutrient Criteria—Estuarine and Coastal Waters
-------
Chesapeake Bay
Harding and Perry
Chesapeake Bay showing locations of the 6 regions.
Nutrient Criteria—Estuarine and Coastal Waters
G-23
-------
Regions in the Chesapeake Bay by latitude and geographic location.
Region
I
II
m
IV
V
VI
Latitude range
36.95-37.40NN
37.41-37.80NN
37.81-38.40NN
38.41-38. 80NN
38.81-39.10NN
39.11-39.66NN
Geographic location
Mouth of Bay to Mobjack Bay
Mobjack Bay to Rappahannock River
Rappahannock River to Patuxent River
Patuxent River to South River/ Annapolis
South River/ Annapolis to Bay
Bridge/Magothy River
Bay Bridge/Magothy River to Susquehanna
Flats
G-24
Nutrient Criteria—Estuarine and Coastal Waters
-------
Year
1985
1986
1987
1988
1989
1990
Region
I
II
III
IV
V
VI
I
II
III
IV
V
VI
I
II
III
IV
V
VI
I
II
III
IV
V
VI
I
II
III
IV
V
VI
I
II
III
IV
V
VI
n
227
254
149
142
124
106
240
243
161
141
143
118
265
249
173
148
159
128
251
241
160
124
126
109
243
234
252
116
230
223
218
234
130
129
146
118
LS mean
5.17
5.77
7.19
7.84
10.5.
7.17
5.21
6.79
7.42
7.81
10.6
6.74
8.91
10.6
9.86
10.5.
12.1
7.60
5.05
8.78
10.3
8.42
10.9
6.04
5.98
7.51
8.39
8.93
10.9
5.69
6.31
10.4
9.90
9.49
10.8
4.76
95% LCI
4.63
5.22
6.32
6.88
9.17
6.15
4.68
6.13
6.56
6.85
9.39
5.82
8.12
9.61
8.79
9.26
10.8
6.62
4.55
7.95
9.17
7.33
9.53
5.18
5.40
7.08
7.62
7.74
9.85
5.10
5.67
9.41
8.66
8.30
9.53
4.08
95'% L'CI
5.7.5
6.38
8.17
8.92
12.0
8.3-1
5.78
7.50
8.3S
8.8S'
12.1
7.78
9.78
11.6
11.1
11.8
13.7
8.71
5.60
9.67
ll.fi
9.66
12.4
7.1W
6.fiU
7.95
9.23
10.3
12.0
6.33
7.02
11.4
11.3
10.8
12.2
5.54
Summary statistics on surface chlorophyll
concentrations (mg m"3) from Monitoring Program
cruises on the Chesapeake Bay, 1985-1990.
Nutrient Criteria—Estuarine and Coastal Waters
G-25
-------
York River
77.20
I
77.00
I
76.80
I
YORK RIVER SYSTEM
The Environmental Protection Agency Chesapeake Bay Monitoring stations in the
York River estuary.
G-26
Nutrient Criteria—Estuarine and Coastal Waters
-------
A) Tidal freshwater station (TF4.2)
B) Transition station (RET4.1)
40
35
30
25
20
15
10
5
0
• Chlorophyll a
Primary Production
I \
40
35
30
25
20
15
10
o
JFMAMJJASONOJ
C) Transition station (RET4.3)
JFMAMJJASONDJ
D) Estuary station (LE4.1)
JFMAMJJASONDJ
E) Estuary station (LE4.2)
JFMAMJJASONDJ
F) Estuary station (LE4.3)
_o
6
\ i V-Hi
i-.
JFMAMJJASONDJ
JFMAMJJASONDJ
Month
Seasonal distributions of chlorophyll a and primary production in the York River system;
monthly means and standard errors were calculated from the 10 years data (1985-1994)
for chlorophyll a and from 7 years data (1988-1994) for primary production. Dashed line
at 10 ug 11"1 indicates our criterion for algal blooms and primary production shown in Fig
3F was measured at WE4.2.
Nutrient Criteria—Estuarine and Coastal Waters
G-27
-------
& 5
O
O
in 3
250 -
_ 20°
~ 150
CO
n 100
0
250 -
-= 150
m
3 100
G)
H)
10 15 20 25
1990
1991
1994
0 5 10 15 20 25
Salinity (psu)
14
§;:
Z 8
O™ 6
0
14 -
O"
12 -
10 -
e
z
E 4
m 2
o.o -
2.0 -
2 1.0-
F>
25
f— 20
en
I"
"o. 10
S
o s-
10 15 20 25
10 15 20 25
10 15 20 25
10 15 20 2
Salinity (psu)
Salinity dilution curves of DO, ammonium, orthophosphate, silicate, nitrite + nitrate,
chlorophyll a and light attenuation coefficient in the water column of the York River system
for low (1991), mean (1990) and high (1994)flow years during the summer-fall period (June
through October).
G-28
Nutrient Criteria—Estuarine and Coastal Waters
-------
Neuse River Estuary, NC
*3? ^ #s
J?l £f *P ATL
°
APE
HATTERAS
^*
I ^-f-w^x if > rr»
r^V^p
.Z^&fo/f'
ATLANTIC
OCEAN
O lO 2O 3O *O
8OGU61
SOUND
'-BEAUFORT
Sampling locations (Stns 1, 5, and 6) in the lower Neuse River Estuary,
North Carolina, USA.
Nutrient Criteria—Estuarine and Coastal Waters
G-29
-------
V*)
/! 1
Neuse River mean monthly flows (x 1000 ft3 s"1 or x 28.3
m"3 s"1), at the US Geological Survey gauging station in
Kinston North Carolina. Dashed line represents 60 yr
average flow (J.D. Bales, US Dept of Interior, Geological
Survey, Water Resources Division, Raleigh, North
Carolina). (Middle) Surface measurements of salinity at
Stns 1, 5, and 6 (see Fig. 1). Continuous line represents
means of stations measure. (Bottom) Monthly rainfall
totals at the Institute of Marine Sciences, Morehead City,
North Carolina. Dashed line represents average monthly
precipitation for the southern section of the Albermarle-
Pamilico estuarine system area (118.5 mm mo"1) (H.
Porter, University of North Carolina Institute of Marine
Sciences, Morehead City).
G-30
Nutrient Criteria—Estuarine and Coastal Waters
-------
2
12
10
IL
12
10
8
6-
2--
DJFMAMJJASONDJFMAMJJASONDjFMAMJJAS0
NH4
DjFMAMJJASONDjFMAMJJASONDjFMAMJJA SO
P0:
oSTA 6
•STA 5
«STA 1
ND IFMAMJJASOND iFMAMJJASOND IFMAMJJASON
1987° 1988 u 1989 1990
Nitrate, ammonium, and phosphate concentrations in
surface waters at Stns 1,5, and 6 (see Fig. 1).
Continuous line represents means among stations.
Nutrient Criteria—Estuarine and Coastal Waters
G-31
-------
30
25
20
"°15
_i
x
JO
20
o ,5
10
40
?
2
° 30-
j.
5 20-)
z
O
10-
OjfMAMJJASONOjFhAMJaASONDjFMAMJJASO
OjFMAMJ JASONO JFMAMJ JASONDJFMAMJ JASO
!
/!
1 6
l\
*fr
-H-
NOJFMAMJJASONDJFMAMJJASONOJFMAMJJASON
1987
1988
1989
1990
(Top and middle) Chlorophyll a concentration of selected
nutrient addition treatments minus controls, averaged over
the 4 d of eachbioassay. N( 14.3) and N(28.6) respectively
indicate addition of 14.3 and 28.6 um NO3- P(3.2) indicates
addition of 3.2 um PO43". Error bars not visible are smaller
than symbol. (Bottom) Dissolved inorganic nitrogen:
dissolved inorganic phophorus (DIN:DIP) ratio (by atoms).
No error bars plotted.
G-32
Nutrient Criteria—Estuarine and Coastal Waters
-------
6--
3--
DjFMAWJJASONojFMAMJJASONDjFMAMJJASO
DjFMAMJ JASOND JFMAMJ JASON D JFMAMJ JASO
NO jFMAMJJASONO jFMAMJJASONDjFMAMJJ ASCN
1987 1988 1989 1990
(Top and middle) 14C assimilation of selected
nutrient addition treatments minus xontrols,
averaged over the 4 d of each bioassay. Symbols
as in Fig. 4 (Bottom Nitrate, ammonium, and
phosphate concentrations in surface waters at Stn 6
(see Fig. 1). Data are compiled from Fig.3 and
presented here for comparative puposes.
Nutrient Criteria—Estuarine and Coastal Waters
G-33
-------
Cape Fear River Estuary
Greensboro
NAV
.Fayetteville
^^V
Lock and Dam #1
Sampling stations along the Cape Fear River Estuary, North Carolina,
United States. The lower estuary is centered in 33°56'N, 77°58'W.
G-34
Nutrient Criteria—Estuarine and Coastal Waters
-------
40 T
JAODFAJAODFAJAODFAJAOD
JAODFAJAODFAJAODFAJAOD
120 i
5-100
JAODFAJAODFAJAODFAJAOD
JAODFAJAODFAJAODFAJAOD
1995 1996 1997 1998
Physical parameters for selected stations in the Caper Fear River
Estuary, June 1995-November 1998.
Nutrient Criteria—Estuarine and Coastal Waters
G-35
-------
JAODFAJAODFAJAODFAJAOD
JAODFAJAODFAJAODFAJAOD
JAODFAJAODFAJAODFAJAOD
—»—NAV
--X--M54
—•—M35
• -O--M18
JAODFAJAODFAJAODFAJAOD
Inorganic nutrient concentrations for selected stations in the Cape
Fear River Estuary, Junel995- November 1998.
G-36
Nutrient Criteria—Estuarine and Coastal Waters
-------
Coastal Georgia
Nutrient exchange (|_ig-atom m2 d"1) across the sediment/water interface and initial nutrient
concentrations (|ag-atom I"1) in the overlying water of the nearshore environment. Values
given as release (+) or uptake (-) of the nutrient by the sediment.
Constituent
O2
NH4+
NO2-
NO3
DON
PO4=
OOP
2 Inorganic nitrogen
SN
2 Inorganic phosphorus
SP
Initial Concentration p,M
6.5
1.1
0.11
0.16
24.3
1.4
0.0
1.26
25.56
1.4
1.4
Nutrient Criteria—Estuarine and Coastal Waters
G-37
-------
Gulf of Mexico Estuaries
GULF OF MEXICO ESTUARIES J \ -
CARIBBEAN SKA
Map showing the distribution of estuaries in the Gulf of Mexico.
G-38
Nutrient Criteria—Estuarine and Coastal Waters
-------
Site
Florida Bay, FL
Tampa Bay, FL
Apalachicola Bay, FL
Mobile Bay, AL
Fourleague Bay, LA
Nueces River Estuary,
TX
Celestun Lagoon,
Mexico
Chelum Laggon,
Mexico
Dzilam Lagoon,
Mexico
Rio Lagartos Lagoon,
Mexico
Area
(km2)
1800
896
593
1060
56
538
28
13.6
9.4
96
Volume
(106 m3)
3200
3490
1600
3200
72.8
1290
33
16.3
11.2
76.8
Mean
Depth
(m)
2.0
3.9
2.7
3.0
1.3
2.4
1.2
1.0
0.8
0.8
Maximum
Length
(km)
70.0
61.1
12.9
48.5
16.0
28.2
20.7
14.7
12.9
80.0
Maximum
Width
(km)
40.0
16.1
33.8
40.0
3.5
12.9
2.1
1.8
1.6
1.5
Carbonate system low
freshwater input
Marine dominated estuary
River dominated with micro-
tides
River dominated with
seasonal coastal plume
Shallow, wetland dominated
salinity variable usually low.
Event driven estuary, low
river flows
Freshwater spring-fed
lagoons.
Table 5-1. Physical Characteristics of Gulf Estuaries Examined in Case Studies.
Nutrient Criteria—Estuarine and Coastal Waters
G-39
-------
Site
Florida Bay
Eastern Region
Central Region
Western Region
Tampa Bay
Apalachicola Bay
Mobile Bay
Upper Bay
Mid Bay
Lower Bay
Fourleague Bay
Nueces River Estuary
Nueces Bay
Corpus Christ! Bay
Celestun
Chelem
Dzilam
Rio Lagartos
Salinity (g liter"1)
29 (0.2-45)
33 (9-63)
35 (25-51)
— (20-35)
16.6 (0-37.3)
5.2 (0-19)
9.1 (0-23)
15.3(0-32)
—
18.8 (3-30)
32.1 (28.6-37.9)
25 (5-37)
36 (27-43)
31 (30-37)
57 (20-100)
N03 (pM)
0.62 (0.01-10.0)
0.25 (0.01-5.70)
0.12 (0.01-7.25)
—
7.6 (1-35.7)
6.6 (0-40)
4.1 (0-19)
3.3 (0-18)
42
4.6 (0.5-23)
3.2 (1.5-11.3)
4.82 (0.9-1.5)
1.89 (1-6)
7.07 (1-10)
0.7 (0.2-5)
NH4 (fiM)
3.19 (0.03-82.1)
5.31 (0.01-120)
0.12 (0.01-7.25)
—
1.4 (0.1-7.1)
3.2 (0-17)
2.4 (0-13)
1.9 (0-12)
2.5
4.63 (0.5-19)
7.82 (1-90)
7.31 (1-38)
2.5 (1-15)
8.5 (2-21)
P04 (pM)
0.03 (0.01-0.51)
0.04 (0.01-0.84)
0.03 (0.01-0.39)
—
0.16 (0.1-0.62)
0.58 (0.05-1.4)
0.46 (0.05-1.7)
0.38 (0.05-1.6)
0.8
1.99 (0.6-4.6)
0.82 (0.02-7.2)
0.41 (0.1-6)
1.45 (0.2-8.1)
1.55 (0.3-11)
SiO2 (iiM)
16.0 (0.18-122)
65.6 (0.06-109)
18.6(0.13-57.1)
—
51.8 (3.0-136)
02.1 (13-131)
59.7(8-110)
40.5 (2-104)
—
38.2 (10.5-60.8)
54.4 (5-220)
36.8 (4-50)
163 (12-210)
24.7 (5-75)
Annual Mean and Range of Salinity and Nutrient Concentrations for Gulf Estuaries Examined in the Case Studies.
C Itir * dt '}
.137.1)
2,3
1.2 (0.1-6)
1.2 (0.21-2.1)
0.42 (0.3
O.S3
0 74 0.0S4-O.S75)
771
842
683
661
1440000-2520)
480
420 {40-/30)
S40 (40-700}
370
525
153 (109-182)
153(14-286)
215(H-255)
Annual mean and range of Chlorophyll-a Primary Production of Gulf Estuaries Examined in the Case Studies.
G-40
Nutrient Criteria—Estuarine and Coastal Waters
-------
Galveston Bay
Ortho-phosphate (nig P I ') 0.21 + 0.001 0.22 i 0.01 0 42 t 002 007 L O.IXHt 0.07 ± 0.007
Total phosphate (ing I1 I ') 0 Ifi t on? 1)27 r u 01 ti.il t 002 O.lOi .t 0,8113 0.10 -t 0.009
Nitrate (mg NI ') 0.10 * 0.025 008 i 001 0.43 t 0.05 Q.ii.l ± 0.007 0.04 t 0.01
Ammonia (ing N 1 ') 006 i 0.01 0.08 t 001 U.20 t 002 0 OS t 0.02 0 Ob t. U.fll
NitraiefmgN I-') - 0.035 ±. 0.01 0.13 ±0.02
Total Kjddahl nitrogett - 1,4 i 0.06
Salinity (o%) 9 i ± i 3 17 1 -t Of. 15 3 » 1 1 15.9 t o.B
Total suspended solids (nig I"')" 15 t 1.7 20 i 1 i 277 ± S 7
Total organic carbon (pig C I ''} — 7,4 t 0 5 44 ± 0.7
CIilorophyli-u{,iigl"'j 3.8 .t 09 g i I 146 t 2.4 3 0 . 1.4
" (.'ailed 'total residue concentration* in TWC database. The anihinclit averages are for illustrative purposes only, and do nol claim iruc sta-
tistical meaning; Morgan's t'uint has highci salinily due lo ihe fast trunsporl ol' ieawaier in ihe ship channel, anJ lower freshvvaier input from
Ihe San Jacinio River.
Average Concentrations of Nutrients and Other Chemical Parameters in Galveston Bay, Calculated from the
TWC Database (± lo).
Nutrient Criteria—Estuarine and Coastal Waters G-41
-------
Puget Sound
Padilla Bay
49°N-
48°N-
47°N-
T
124°W
Map showing location of Padilla Bay in relation to Puget Sound (left) and the study site (X) in
Padilla Bay.
G-42
Nutrient Criteria—Estuarine and Coastal Waters
-------
I
V)
V,
z
35
30
25
20
15
10
June
—•— Nitrate
—•— Silicate
- -» - - Ammonium
- -A--Phosphate
July
I 11 I 11 I [I I 11 I 11 I 11 I 11 I 11 I II I I
September October
o t
3 JS
a.
o
PH
2 Z
i
1 •=
August
Sampling Date
Dissolved inorganic nutrient concentration in surface water at the
study site from June to October 1992.
18 i-
16
14
g 12
Z
S 10
8
1 4
^ Chlorophyll a
June
July
3 «J
I
a,
g
_o
U
August September October
Sampling Date
DIN:PO43" ratios and chlorophyll a concentrations in surface water at
the study site from June to October 1992. DIN is NO3" + NO2" + NH4+
PO43" is soluble reactive phosphate. The dotted line represents the
Redfield ratio of 16:1.
Nutrient Criteria—Estuarine and Coastal Waters
G-43
-------
APPENDIX H
PRELIMINARY STATEMENT OF PROPOSED NEAR COASTAL MARINE
NUTRIENT SAMPLING AND REFERENCE CONDITION
DEVELOPMENT PROCEDURE
Synopsis of the National Nutrient Criteria Program, Coastal Marine Sampling Design Planning Meeting,
4-5 June 2001, USEPA Environmental Science Center, Ft. Meade, MD.
Near Coastal Marine Nutrient Sampling and Reference Condition Development Committee Members:
Barry Burgan, USEPA; John Fox, USEPA; Laura Gabanski, USEPA; Jeroen Gerritsen, Tetra Tech Inc.;
George Gibson, USEPA*; William Muir, USEPA; Kent Price, Univ. Del; Don Pryor, NOAA; Greg
Smith, GLEC, Inc.; Val Smith, Univ. Kansas; and Jack Word, MEC Analytical Systems, Inc.
*Contact: USEPA Laboratory, 701 Mapes Road, Ft. Meade, Md. 20755-5350. Phone: (410) 305-2618
E-mail: Gibson.George@EPA.gov.
Background and Purpose
Cultural eutrophication is an established water quality management concept and concern reaching as far
back as the 1600s in America (Capper, Power and Shivers 1983). However, extensive public recognition
of this form of pollution in coastal waters is relatively recent. The publication "Eutrophication, Causes,
Consequences, and Correctives" (NAS 1970) is often perceived as the technological beginning of
American nutrient pollution awareness and is centered on the understanding and abatement of this
problem primarily in freshwater lakes and reservoirs. We have since come to better understand the
problem in streams, rivers and estuaries with the publicity and public involvement in the Chesapeake Bay
studies of the 1980s. Vollenweider, Marchetti, and Viviani published "Marine Coastal Eutrophication"
in March of 1992 and this volume may be considered the coastal equivalent of the land mark NAS
freshwater publication a decade earlier.
In response to this growing awareness, the EPA National Nutrient Criteria Program is preparing technical
guidance for nutrient reference condition determination and related criteria development to be used by
States and Tribes in the reduction of cultural eutrophication of the Nations' surface waters. This report
concentrates on coastal marine waters and the effort to identify relatively natural nutrient water quality
conditions, which can be used as a benchmark to evaluate cultural eutrophication or overenrichment. A
preliminary literature investigation and data search indicate that insufficient data exist to derive the
reference condition information suitable for the needs of the Program without going to primary data
collection.
Because data gathering is likely to be a preliminary concern as well as an ongoing requirement, this
meeting of coastal marine nutrient research and management specialists was called to design a standard
protocol that the EPA can recommend for use in U.S. marine coastal waters. Coastal marine waters are
defined as those waters within 20 miles of shore along the East, West, and Gulf coasts of the United
States as well as Alaska, Hawaii, and the U.S. Trust Territories. Emphasis is on the three mile limit State
waters, although interest may devolve to the 12- mile U.S. limit as well. Nutrient loading from cultural
land run off sources are not presently expected to be a serious problem beyond this limit. The general
design and protocol are applied here to a case study example, the coastal waters of the mid-Atlantic Bight
from New Jersey to the Virginia Capes. Many of the design elements, in particular the number, size and
placement of spatial elements (strata and cells; see below) would need to be modified for specific
applications in different parts of the U.S. coastline.
Nutrient Criteria—Estuarine and Coastal Waters H-1
-------
A coastal transect of fixed stations exists for most of the Mid-Atlantic Bight, which has been used by
EPA and NOAA for several years to collect nitrogen, phosphorus, Secchi depth (SD), and chlorophyll-a
(Chl-a) data. This procedure and data base will be the prototype presented and discussed to develop the
recommended protocol.
Objective
To determine a simple, cost effective, scientifically defensible and standardized method to sample for
marine enrichment variables to use in determining reference condition for nutrient criteria derivation.
Premise of the National Coastal Nutrient Criteria Program
Offshore marine and onshore, near-coastal sites removed from point and estuarine discharges can be
identified as reference sites reflecting the least culturally impacted nutrient water quality of a region.
"Region" in this case is a geographically similar portion of the coastline such as the Mid-Atlantic Bight.
Such regions can also be described from the coastal portion of the Level III nutrient ecoregion map of the
continental United States (which is consistent with the rest of the National Program and is similar to the
ORD Provinces used by EMAP).
Nutrient water quality is established from representative sampling of the coastal waters at these a priori
reference sites. The other elements of nutrient criteria, i.e., historical trends, modeling of the data for
additional insights, and attention to the consequences down-current of any proposed nutrient criteria, and
assessment of all of this information by a Regional Technical Assistance Group (RTAG) are applied to
the initial reference condition values to develop nutrient criteria for total phosphorus (TP), total nitrogen
(TN), Chl-a, and SD.
These criteria can then be used by States and Tribes to manage and monitor the nutrient quality of their
coastal marine waters. While this concept was developed and initiated by the USEPA beginning with the
Biological Criteria Program in 1989 (EPA-440/5-90-004, EPA-440/5-91-005) and further refined and
applied to nutrients by the EPA National Nutrient Criteria Program in 1995 (EPA 822-R-96-004, EPA
822-R-98-002), the idea has also been independently developed by the Swedish Environmental
Protection Agency using the reference condition approach on a regional basis and employing the same
indicator variables (Report number 5052, 2000).
Importance
All other waters of the continent drain to the coast and these coastal marine areas are the recipient of any
nutrients not intercepted from that cumulative runoff. Globally, conditions of many coastal areas have
shown several-fold increased levels of nitrogen and phosphorus since industrialization (Smith 1998,
Smith, Tilman and Nekola 1999), and preliminary assessments of empirical data collected from the Mid-
Atlantic Bight between New Jersey and North Carolina by USEPA Region III since 1987 have suggested
an upward trend in the concentration of both dissolved inorganic nitrogen (DIN) and dissolved inorganic
phosphorus (DIP) at stations approximately 1 to 5 nm offshore (Muir, pers com, 2001). An extensive
baseline of region specific coastal nutrient data, regularly and consistently collected, is needed to
establish criteria against which future loading conditions may be compared. Potential predictive models
may also be developed from this information relating algal booms and other biological responses to these
nutrient levels in coastal waters.
The Transect Sampling Design
As illustrated below (Figure H-l, of coastal sampling stations between Atlantic City, NJ and Kitty Hawk,
NC) a series of transect sampling stations have been located between one and five miles from shore along
a designated coast line to measure ambient water quality as reflected in TN, TP, Chl-a and SD. Data
collected from those stations determined to be remote from significant cultural impacts such as sewage
H-2 Nutrient Criteria—Estuarine and Coastal Waters
-------
Figure H-l. Example of an initial coastal monitoring project conducted off the mid-Atlantic
shelf just north of Delaware Bay and south of the mouth of the Chesapeake Bay to
obtain nutrient reference condition values. Open circles are reference sites and the
inshore references are compared to the offshore ones for confirmation.
Philadelphia, PA
Baltimore, MD
Washington, D.C. •
J16 -J16Z
J17 -J17y
>J18 -J18y -JlSz
>J19
•J20
1D11 -CDlly -CDllz
•CD12
•CD13 -CD13y -CD13z
•CD14 *CD14y -CD14z
°CD15
•CD16z
Virginia Beach, VA
•C14z
•GG4z
Nutrient Criteria—Estuarine and Coastal Waters
H-3
-------
discharges, industrial activities, major port facilities, or estuarine discharges constitute the reference
condition. The values measured can be compared to marine conditions further offshore (e.g., 20-25 nm)
reflecting, at least for TN and TP, the unimpacted condition.
This comparison, together with attention to prevailing currents including upwellings, would establish the
seasonal values for TN, TP, Chl-a, and SD. Salinity would also be an important variable to document
local constancy of the waters and to avoid discharge plumes from rivers and estuaries that are not part of
the defined coastal waters of concern. These estuarine and riverine waters should have their own criteria.
Data have been collected from the surface 1 meter, mid depth, and the bottom 1 meter of the water
column at each station. Measurements are made on a seasonal basis, essentially mid-summer and mid-
winter and with sufficient data, criteria can be established for each season. While the protocol as
described above has been in use for about twenty years, most of those data are for dissolved inorganic
nitrogen and phosphorus rather than TN and TP, Chl-a, and SD. The TN, TP and offshore stations are a
recent addition and only two summers of data are available which is reported here.
Data Collection
Operations are during daylight hours only in order to include SD measurements with surface nutrient
collections and to maintain a consistent nutrient depth profile relative to photo periodicity. Additional
data collected with a CTD includes salinity, pH, temperature, depth, conductivity, and DO.
Data sampling points presently include the discharge plumes of estuaries, rivers, and point dischargers to
monitor impacts and design management plans, but not as part of the proposed coastal reference
condition sampling system per se.
The Pilot Project
The area approximately between Atlantic City, NJ and Kitty Hawk, NC has been studied by EPA Region
III for about 17 years and includes a mix of nutrient, chemical, biological, and physical measurements.
This data has been processed and will provide trend information about the area.
For the last two summers and one winter, nutrient data have been collected from this area in the manner
described above. This is the initial basis for a reference condition determination and is presented in
Figures H-2 a-d, below.
If the pilot project is judged successful, it is expected that the process, training, and funds for similar
equipment will be provided to each of the coastal EPA Regions for comparable operations to develop
their ecoregional coastal nutrient criteria. This area more than any other because of the proximity of
State and Federal waters will lend itself to joint data gathering and criteria development.
Methods and Materials
Forty-nine sampling stations are located along the 200 nautical mile (nm) stretch of coastal waters.
There are twenty stations roughly 10 miles apart situated one to five miles offshore (total of 39
consisting of either single stations or sets of two or three), there are five intermediate (ten miles off
shore) stations, and there are eight stations located about 20 nm offshore (Figure H-l).
Sampling was conducted from the OSV Peter W. Anderson using a Sea-Bird brand CTD and rosette
sampler with 30 L Niskin bottles to produce a continuous water column profile and discrete water
samples from the surface one meter, the mid-depth, and the bottom one meter of the water column at each
station. One liter of sample was filtered using a Millipore Corporation apparatus and 0.7 um fiberglass
filters (Whatman GF/F). Ten ml sub-samples each of water were taken for TN and TP analysis using the
H-4 Nutrient Criteria—Estuarine and Coastal Waters
-------
Figure H-2a. Two-year summer nutrient survey results using a sampling design as illustrated in
Figure 1. Potential reference condition for summer conditions is 0.025 mg/L TP
(NM = nautical miles).
0.05
0.04
0.03
0.02
0.01
l
l
Chesapeake Bay
i i i
i
i
J16 J17 J18 J19 CD11 J20 CD12 CD13 CD14 CD15CD16 GG1 CD17 CD20 CC12 CC13 C14 GG2 GG3 GG4
Reference Sites
Plume Impacted Sites
Referen. Sites
Rume Impacted Sites Reference Sites
-B- 1999 5 NM offshore data
2000 5 NM offshore data
1999 & 2000 20 NM offshore data
Nutrient Criteria—Estuarine and Coastal Waters
H-5
-------
Figure H-2b. Two-year summer nutrient survey results using a sampling design as illustrated in
Figure 1. Potential reference condition for summer conditions is 0.175 mg/L TN
(NM = nautical miles).
0.35
0.3
0.25
0.2
0.15
0.1
J16 J17 J18 J19 CD11 J20 CD12 CD13CD14 CD15 CD16 GG1 CD17 CD20CC12 CC13 C14 GG2 GG3 GG4
Reference Sites
-B- 1999 5 NM offshore data
Rume Impacted Sites I I Referen. Sites
Rume Impacted Sites
Reference Sites
-*- 2000 5 NM offshore data
-X- 1 999 & 2000 20 NM offshore data
H-6
Nutrient Criteria—Estuarine and Coastal Waters
-------
Figure H-2c. Two-year summer nutrient survey results using a sampling design as illustrated in
Figure 1. Potential reference condition for summer conditions is 0.09 jig/L
chlorophyll-a (NM = nautical miles).
° 2
^
6 1
Delaw are Bay
| Chesapeake Bay |
J16 J17 J18 J19 CD11 J20 CD12 CD13 CD14 CD15 CD16 GG1 CD17 CD20 CC12 CC13 C14 GG2 GG3 GG4
Reference Sites
Plume Impacted Sites
Referen. Sites
-B- 1999 5 NM offshore data -*- 2000 5 NM offshore data
Plume Impacted Sites
* 1 999 & 2000 20
Reference Sites
NM off shore data
Nutrient Criteria—Estuarine and Coastal Waters
H-7
-------
Figure H-2d. Two-year summer nutrient survey results using a sampling design as illustrated in
Figure 1. Secchi depth data are incomplete because of missing observations during
night-time operations (NM = nautical miles).
16
14
12
10
g
f 8
CD
T3
•— £•
4
— O
Delaw are Bay
-2
Chesapeake Bay
J16 J17 J18 J19 CD11 J20 CD12 CD13 CD14 CD15 CD16 GG1 CD17 CD20CC12 CC13 C14 GG2 GG3 GG4
Reference Sites
Rume Impacted Sites
Referen. Sites
Rume Impacted Sites I Reference Sites I
1999 5 NM off shore data o 2000 5 NM offshore data
H-8
Nutrient Criteria—Estuarine and Coastal Waters
-------
Standard Methods persulfate digestion method. All samples were frozen on board for later analysis at the
University of Maryland Chesapeake Biological Laboratory. Secchi depth was determined using a 35 cm
diameter white Secchi disc on 0.5 m marked line.
Results
Preliminary data analyses indicated that the groupings of two or three stations (originally intended to
reflect progressive offshore encroachment of nutrient runoff) are not significantly different. Similarly,
initial assessment of the surface, mid-depth, and bottom water column samples showed no significant
differences for the nutrient criteria indicators except for DO. Salinity did show the expected surface to
bottom variations expected of near coastal waters. Consequently horizontal and vertical sampling
distances for each station were combined to produce the nutrient results presented here.
The following figures illustrate the results of two summer surveys in 1999 and 2000. Data indicated by
an asterisk represents the mean of combined two year data from each 20 nm offshore station. Gaps in
the Secchi depth results are because of the night time intervals during 24-hour operations. The stations
located in the vicinity of the estuarine discharges of Delaware and Chesapeake Bays are indicated by the
appropriately labeled open boxes, i.e. J18 through CD12 for the Delaware and CD17 through CC13 for
the Chesapeake.
Discussion
The survey technique appears to faithfully reflect the nutrient conditions of the coastline in that
higher nutrients were found in the waters just off each estuary. Further, when surface station
salinity was plotted, a mirror image of the nutrient data results was presented indicating a
freshwater correlation with higher nutrient concentrations. The biological response of the waters
was also evident from the correlation of Chi-a and SD data with the TN and TP levels.
Given the variability in only two years of data, it is interesting that the offshore stations appear to
be relatively consistent in all three nutrient parameters. The incongruity of the TP and TN data
at GG1 for 1999 and for chlorophyll-a at CD16 for 2000 is also noted, but unexplained.
The two consecutive sampling years demonstrated similar trends among the stations indicating
no major weather variability during this period and also suggesting that the proposed reference
sites respond comparably to the discharge sites to inter-annual climatological events.
• The stations that consist of sets of two or three sites within a few miles of one another were
determined upon review of the data to have very little distinction. This suggests that
perturbation, if any, originating from the coastal land mass has had an impact over all three sites.
It appears that the Delaware plume drifts more southward from the mouth of its bay than does the
Chesapeake. This may reflect the lower discharge volume of the Delaware and the influence of
the Hudson River discharge and Longshore current, both displacing the plume southward.
Similarly, the slight northward offset of both the Delaware and Chesapeake plumes in 1999
relative to 2000 may be a response to changing coastal current dynamics from year to year.
• There was no significant difference between the eight offshore sites (20 nm offshore) and their
inshore counterparts. Further comparison of offshore stations to those counterparts within the
estuarine discharge plumes will help determine the sensitivity of the comparison. The
intermediate 10 nm stations do not appear to add substantial information and can be discontinued
except in the vicinity of discharge plumes to help define these margins.
Nutrient Criteria—Estuarine and Coastal Waters H-9
-------
While the data is limited, it is encouraging that the offshore controls are comparable to the
expected inshore reference sites. Candidate inshore reference sites are initially selected on the
basis of an apparent physical absence of local cultural impact, i.e. tributary discharges, municipal
discharges, ports or marinas, or other commercial enterprises. Because of the potentially high
variability in the existing data, a concurrence between offshore and inshore reference stations is
not necessarily confirmation of the quality of the inshore reference sites, but when consistent
with observed physical indications, this information adds confidence to the selection.
Conversely, significant differences would be cause for suspecting the inshore site selection as of
reference quality.
• In this regard, an interesting trend was noted among the three groups of reference sites located
north of Delaware Bay; between Delaware and Chesapeake Bay; and south of Chesapeake Bay
(Figures 2a-d). The mean ambient concentrations of TP, TN and Chlorophyll-a at these
reference sites trend downward from north to south. The same trend appears in the eight
stations 20 nm offshore presumably reflecting a broad scale process affecting this area. This
further supports the importance of establishing relatively close spaced reference sites when
preparing coastal marine criteria The mean TP values for the nearshore reference sites at the
northern terminus of the transect of stations is significantly higher than those at the southern end
(p = 0.0006) even though the region is presumed to be geologically homogeneous.
• Secchi depth data were inconclusive because not enough data points were generated as a
consequence of 24-hour sampling when Secchi depth could not be determined during night-time
hours. Additional future sampling will be conducted during daylight so all parameters may be
evaluated.
Committee Discussion of the Prototype Methodology
Fixed Station Sampling vs Stratified Random Sampling
Inferences derived from fixed-station and fixed-transect sampling, while common in oceanographic
research and monitoring, are potentially confounded by unintentional and unknown biases, and by the
inability to extend statistical inferences to the entire sample space desired. Alternatively, fixed stations
tend to reduce the amount of unknown physical variability associated with interpreting climatic factors
upon a given site such as when attempting to assess hurricanes, upwelling or acid deposition effects on a
particular coastline. Although there is no reason to suspect that the existing stations of the mid-Atlantic
coastal nutrient study are biased, the design group thought that the design should allow data inference to
the entire coastal sampling space. It therefore proposes a change to a probability-based design, of equal
sampling cost, to avoid the potential pitfalls of a fixed-station design.
The sample space for this project is open marine waters of the U.S. coastal zone, with emphasis on state
waters within the 3 nautical mile state limit. Sampling will be carried out in three sampling strata for
which nutrient conditions are to be estimated:
1. Reference areas within the 3-mile limit, outside the influence of major estuary plumes (e.g.,
Hudson River, Delaware Bay, Chesapeake Bay);
2. Nutrient influenced areas within the 3-mile limit, affected by the estuary plumes and other
discharges; and
3. Offshore waters beyond the state 3-mile limit.
H-10 Nutrient Criteria—Estuarine and Coastal Waters
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These three regions will define the sampling strata. The sampling design will be to define longshore
"cells" in each of these three zones. During each sampling event, one site will be selected randomly
within each sampling cell. The overall design can be described as stratified-systematic-random; where
the strata are the three areas defined by estuary influence and distance from shore, the systematic
component is the cells that define each stratum, and the random component is the random sampling
location selected on each cell during each sampling event (Figure H-3).
The first task in developing the sampling plan will be to define areas of presumed estuarine influence,
from existing physical oceanographic research on water movement, in this case in the mid-Atlantic
coastal area, and from existing water chemistry data showing elevated nutrients and other constituents
such as salinity and conductivity in the estuary plumes. Because plumes will vary with estuarine
discharge, Gulf Stream eddies, and other events, precise definition of the plume area is not possible.
Instead, plume areas should be defined as where the estuarine influence is likely to occur. An
understanding of local estuarine hydrology will help in this regard, but extensive and expensive physical
investigations should not be a prerequisite for the determination of likely plume influence.
Coastal reference site determinations and reference condition derivation should be, at least initially,
established by EPA. Data collection would be continued as an EPA function because the offshore
stations are the purview of the Federal government and because the related reference sites can be
incorporated in the effort.
Estuarine and riverine plume monitoring are more likely accomplished by the coastal States using
existing budgets and vessels already at their disposal. A coordinated effort relating State and Federal
sampling and data exchange should be promoted. The nutrient quality of the discharge plumes can then
be expeditiously compared to the proximal reference condition(s) to assess impact upon the near coastal
marine waters.
Regional coastal characteristics will determine both the reference site cell structures and placement and
the estuarine or riverine plume sampling designs.
Sampling Times and Depths
The committee recommended that sampling be conducted during the period of optimal marine vegetative
growth. In temperate areas this is Spring and Summer, generally May, June and July, August. Other
interval options depending on locale might be wet and dry periods during the growing season.
As a cost effective approach producing scientifically valid results, sampling depth is recommended as
always at the surface, i.e., top meter of water accompanied by either a composite sample from the
remainder of the water column or sampling from just below the thermocline and at one meter above the
bottom. Some members of the committee advocate the surface, mid-depth, and bottom sampling
technique where the mid-depth sample is usually below the thermocline and/or where inshore waters
often fail to demonstrate a thermocline.
Variables
The primary four variables of TN, TP, Chlorophyll-a, and SD are recommended because they are the best
early indicators of causal and biological response indicators to nutrient loadings. Other measures of
clarity or transparency may also be used, but Secchi depth should always be included because such a
large body of information is already available in this form; it is inexpensive and reliable; and continued
Secchi depth measurements provide a continuity with much historical data. In an independent
Nutrient Criteria—Estuarine and Coastal Waters H-11
-------
Figure H-3. Illustration of stratified random grid sample design for reference condition cells as
related to an estuarine discharge. Nutrient quality and spatial extent of the grid
can be compared to a value derived from the measurements in reference sites (cells)
A-E. Values at 10 and 20 nm stations are a check against reference values found at
A-E.
1-3 NM
A
(X)
10 NM
(X)
20 NM
C
D
(X)
10 NM
(X)
20 NM
H-12
Nutrient Criteria—Estuarine and Coastal Waters
-------
investigation, the Swedish Environmental Protection Agency selected the same four primary variables
(Report 5052, 2000).
Other recommended variables are: dissolved oxygen as an important secondary response variable almost
always measured by investigators because of the significance of respiration to the biological community,
and planktonic species composition as a refined diagnostic indicator of the nature and extent of
enrichment.
Geographic Application of the Protocol
Members of the committee are familiar with both the East, West, and Gulf coasts of the continental U.S.
and conclude that the method described above, with allowances for regional modifications such as
relative distance from shore to shelf break and the magnitude of upwelling, can be successfully applied in
all three coastal environments to identify reference conditions for criteria development.
Summary Conclusions
1. The basic protocol as described, but modified to include a probability-based sampling design, a
variation on the surface, mid-depth, and bottom sampling profile, and sampling emphasis on
twice during the growing season, is a scientifically defensible and broadly applicable method for
establishing regional reference conditions to support coastal marine nutrient criteria
development.
2. The TP, TN, Chlorophyll-a, and perhaps Secchi depth variables are responsive and together with
salinity measurements are descriptive of both estuarine discharge plumes and near coastal
reference quality waters. These measurements can be used to assess the concentration and area
changes of discharge plumes overtime.
3. The comparison of marine nutrient water quality to inshore reference sites is valuable as
confirmation of "natural" reference conditions and as a graphic descriptor of the reference
concept for the public. However, for cost effectiveness, these off shore stations need not be
monitored every time the inshore reference sites are sampled.
4. But it is important to note that the distinction of cultural from inherent nutrient discharges by
estuaries and other tributaries to coastal waters is not possible just by comparison to reference
conditions. The other elements of nutrient criteria development should be incorporated to help
make this distinction. The reference conditions and criteria will reveal exceedences of the
"natural" background levels to be preferred and the relative extent and magnitude of the problem,
but source identification and cause and effect studies will be required for an effective
management response to this identified concern.
Nutrient Criteria—Estuarine and Coastal Waters H-13
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