903R87002
United States Environmental Protection Agency
CBP/TRS 16/87
December 1987
Perspectives on
The Chesapeake Bay:
Recent Advances in
Estuarine Sciences
U.S. Environmental Protection Agency
Region III Information Resource
Center (3PM52)
841 Chestnut Street
Philadelphia, PA 19107 .^^
QH
541.5
.E8
P4G
1987
Chesapeake
Bay
Program
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U.S. Environmental Protection Agency
Region HI information Resource
Center (3PM52)
841 Chestnut Street
Philadelphia, PA 19107
PERSPECTIVES ON THE CHESAPEAKE BAY:
ADVANCES IN ESTUARINE SCIENCES
Maurice P. Lynch
Scientific Editor
Elizabeth C. Krome
Technical Editor
December 1987
Chesapeake Research Consortium
P.O.Box 1120
Gloucester Point, Virginia 23062
CRC Publication No. 127
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Published with funds provided by
the Chesapeake Bay Program, EPA.
DISCLAIMER
Mention of trade names or commercial products
does not constitute endorsement
or recommendation for use.
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ACKNOWLEDGEMENTS
Members of the Steering Committee gave guidance; they and
other reviewers contributed helpful and thorough comments on the
draft chapters. Janice Meadows assisted authors in literature searches.
Ruth Hershner managed electronic data files. Pam Owens assisted in
word processing; Susan Myers, G. Glynn Rountree, and Alan B.
Krome helped in proofreading. Throughout the project Karen L.
McDonald provided help and support.
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STEERING COMMITTEE
Dr. Maurice P. Lynch, Chair
Chesapeake Research Consortium
P.O. Box 1120
Gloucester Point, Virginia 23062
Dr. Richard Jachowski
U.S. Fish and Wildlife Service
Branch of Migratory Bird Research
Patuxent Wildlife Research Center
Laurel, Maryland 20708
Dr. Dennis Burton
The Johns Hopkins University
Applied Physics Lab
4800 Atwell Road
Shady Side, Maryland 20764
Dr. William Rickards
Virginia Graduate Marine Science Consortium
Madison House—170 Rugby Road
University of Virginia
Charlottesville, Virginia 22903
Ms. Gail MacKiernan
Maryland Sea Grant College Program
H. J. Patterson Hall
University of Maryland
College Park, Maryland 20742
Dr. Jonathan Garber
University of Maryland
Chesapeake Biological Lab
Box 38
Solomons, Maryland 20688
Dr. Thomas Osborn
The Johns Hopkins University
Suite 315, The Rotunda
711 West 40th Street
Baltimore, Maryland 21211
Dr. Kent Mountford
U.S. Environmental Protection Agency
Chesapeake Bay Liaison Office
410 Severn Avenue
Annapolis, Maryland 21403
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TABLE OF CONTENTS
Preface vii
EXECUTIVE SUMMARY
Introduction 1
Chapter 1 2
Chapter 2 4
Chapter3 5
Chapter 4 6
Chapters 7
Conclusion 9
Chapter 1. Benthic-Pelagic Coupling in the Chesapeake Bay
Jonathan Garber 11
Chapter 2. Factors Driving Changes in the Pelagic Trophic Structure of Estuaries,
with Implications for the Chesapeake Bay
Peter G. Verity 35
Chapter 3. Physical Processes that Control Circulation and Mixing
in Estuarine Systems
E. C. Itsweire andO. M. Phillips 57
Chapter 4. Genetics and the Conservation of Estuarine Species
Laura Adamkewicz, Robert W. Chapman, and Dennis A. Powers 75
Chapter 5. Chemical and Physical Processes Influencing Unavailability
of Toxics in Estuaries
James G. Sanders and Gerhardt F. Riedel 87
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PREFACE
The Chesapeake Bay Program began in the early
1970's as a research effort to uncover and define
problems that had resulted in widely observed
declines in harvestable resources and water quality in
this great estuary. The Environmental Protection
Agency published extensive reports on the results of
this effort, which culminated in the Governors' Con-
ference and the Chesapeake Bay Agreement of 1983.
The Chesapeake Bay Liaison Office was created
as a result of that Conference and Agreement. Today
the office is charged with coordinating extensive state
and federal efforts to reverse the long downward trend
and to restore and protect the Bay's living resources.
The Office directs its resources and staff energy to
implementing the Restoration and Protection Plan
adopted by the Chesapeake Bay Executive Council in
1985. Our role was confirmed in a new basin-wide
agreement signed December 15, 1987.
In this effort, however, far more needs to be
known about the Chesapeake Bay, both to elucidate
its problems and to inform managers making deci-
sions about those problems. Accordingly, the
Chesapeake Bay Program commissioned the Chesa-
peake Research Consortium to sponsor thorough
literature searches in areas where management and the
scientific community agreed our knowledge and
understanding were incomplete.
This volume is directed at defining our needs in
five major topic areas of inquiry:
1. Estuarine benthic-pelagic coupling and flow of
materials
2. Changes in estuarine pelagic-trophic structure
3. Physical processes controlling estuarine
circulation and mixing
4. The role of genetics in conservation of Bay
plant and animal species
5. Estuarine chemical and physical processes and
bioavailability of toxics
These areas were delineated by a steering
committee of Bay scientists and science administra-
tors (elsewhere acknowledged) convened on behalf of
EPA by the Chesapeake Research Consortium. A full
chapter is devoted to each topic area, with summaries
of the available literature (both gray and refereed) and
recommendations about direction and urgency for
future work.
We have provided an executive summary, which
follows this introduction. The summary is an impor-
tant capsule view of the overall product, but I strongly
commend to you the full text of each chapter. These
writings represent some of the best and most compre-
hensive thinking available today, and especially in
their "recommendations" sections, they point to areas
of needed inquiry that are vital to the success of the
Restoration and Protection Program. This publication
and a major research conference in March 1988 will
contribute to the foundation of a comprehensive—and
annually updated—research plan mandated for
delivery in July 1988 by the new Bay Agreement.
The perceptive reader will note that these topics
do not directly include the important areas of living
resource habitat requirements and modeling of the
Bay system (some would say ecosystem modeling).
While both these areas are referred to in the text of the
following chapters, our substantial efforts in these
areas will be found in the following documents:
-Habitat Requirements for Chesapeake Bay
Living Resources, Report of the Living Re-
sources Task Force, Chesapeake Bay Program
Liaison Office, Annapolis, MD 21403, 63 pp
plus appendices
-A Steady State Coupled Hydrodynamic Water
Quality Model of the Eutrophication and Anoxia
Process in Chesapeake Bay (Assignment #40,
EPA Contract 68-03-3319) Prepared by Hydro-
Qual, Inc., for Battelle Ocean Sciences,
Duxbury,MA 21404, in 2 Volumes
This volume is intended as a starting point for
dialogue in the Bay community on research needs. It
has built upon several admirable past efforts including
those of the Consortium, the National Oceanic and
Atmospheric Administration, resource agencies in the
states, and EPA itself. We expect this publication and
the related research conference will help us define
what research is necessary to advance the restoration
and protection of the Bay's resources. All of us—
government, scientific community, and concerned
citizens—must cooperate to solve the dilemma of how
we fund research in a world of limited budgets,
without impeding the restoration and protection effort.
We look forward to the help and advice which
you, the audience of this document, will bring us.
KENT MOUNTFORD, PnD
CHESAPEAKE BAY PROGRAM
ENVIRONMENTAL PROTECTION AGENCY
Vll
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Executive Summary
INTRODUCTION
Scientific knowledge and/or understanding is not
static. The scientific community is continually
advancing its understanding of fundamental principles
or processes that will eventually have impact on
decisions of societal importance. Our knowledge of
the complex Chesapeake Bay system is no exception.
The scientific community has made marked
progress in understanding fundamental processes
underlying management approaches to the restoration
of the Chesapeake Bay. Many of the individual
studies have not, however, been synthesized with
other studies in a context that disseminates their
results to scientists or technical experts from other
disciplines. In addition, as is to be expected in the
study of complex systems or processes, the results
raise additional questions.
This publication reviews five areas of estuarine
research that we believe have critical importance to
the protection and restoration of the Chesapeake Bay.
The particular areas were selected because of their
broad implications beyond the immediate disciplines
in which the studies are conducted.
The Chesapeake Bay Program has already
recognized the critical limitations in our knowledge of
two of the areas: BENTHIC AND PELAGIC COUPLING IN
THE CHESAPEAKE BAY and FACTORS DRIVING
CHANGES IN THE PELAGIC TROPHIC STRUCTURES OE
ESTUARIES, WITH IMPLICATIONS FOR THE CHESAPEAKE
BAY, by initiating plans to develop new information
on these areas for use in the next generation of
models.
We have known for several decades that the life
histories and distribution of living resources within
the estuary are intimately tied to estuarine circulation,
as is the distribution of toxic compounds introduced
into these systems. It is only recently, however, that
we have recognized that our general knowledge of
circulation has not been sufficient to address the ques-
tions raised about many specific issues. It is hoped
that the review of the topic: PHYSICAL PROCESSES
THAT CONTROL CIRCULATION AND MIXING IN ESTUAR-
INE SYSTEMS will assist in elucidating processes
dependent on circulation and will stimulate much-
needed additional research in this area.
The principal underlying goals of the Chesapeake
Bay Program are to restore, protect, and ultimately
manage the living resources of the Bay. The main
thrust at present is to provide appropriate water
quality to sustain these resources. A lesser thrust has
been to develop a better assessment of stocks of
commercial, recreational, and ecological importance
to the system. Explosive advances in methodology
applicable to genetic studies are now being applied to
estuarine species.
The implications of the fourth review topic:
GENETICS AND THE CONSERVATION OF ESTUARINE
SPECIES may open new approaches to restoration and
manipulation of stocks and must be considered in any
emergent plans for living resource management.
The EPA-funded Chesapeake Bay Study provided
us with the first complete benchmark inventory of
toxic compounds in Chesapeake Bay. Additional
efforts since the termination of the formal study have
extended the inventory up the tributaries. Identifica-
tion of toxic "hot spots" and our ability to measure
compounds at ever lower concentrations have led to
the realization that the simple presence of a compound
may not be sufficient to define a toxic problem. Our
final paper: CHEMICAL AND PHYSICAL PROCESSES
INFLUENCING BIOAVAILABILITY OK Toxics IN ESTUARIES
addresses many of the complex phenomena that must
be considered when trying to evaluate the impact of
specific compounds or classes of compounds.
The following capsule summaries of the full
reviews only touch upon the full ramifications of each
topic. The reader is encouraged to read the reviews.
The authors were requested to write the reviews for an
audience of scientists in their own disciplines,
scientists who require knowledge of the particular
processes for understanding other phenomena, and
technically conversant resource managers wishing to
incorporate state-of-the-art scientific knowledge in
their long-term planning. The authors were also
requested to briefly touch on potential management
implications and additional research needs related to
their topics.
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Executive Summary
Summary of Chapter 1:
Benthic-Pelagic Coupling in the Chesapeake Bay
Although the importance of benthic-pelagic
interactions in aquatic systems has been recognized
for half a century, it is evident from this review that
research in the field has been active for little more
than a decade. Scientific studies in the Chesapeake
Bay and elsewhere have provided initial descriptions
of the complex patterns and mechanisms involved in
connecting ecological processes in the water with
those in the sediment. While many unresolved
questions remain, a few preliminary conclusions can
be drawn from this review.
Summary of Major Findings
Various methods have been employed for
measuring both particle deposition and benthic fluxes
of oxygen, ammonium, phosphate, and other
metabolites across the sediment-water interface. It is
encouraging to see that particle deposition rates
estimated by geochemical (and paleobotanical) tracer
techniques tend to converge with rates obtained from
sediment-trap deployments, even though the two
approaches measure this process on very different
time scales. Direct measurements of benthic fluxes of
nutrients and oxygen tend to agree well with observa-
tions made using in situ chambers and those involving
intact sediment cores. Indirect estimates of nutrient
flux based on diffusion modeling of pore-water
concentrations are, however, often 2- to 10-fold lower
than rates obtained from direct measurements,
especially for productive systems with active macro-
fauna burrowing in sediments.
Seasonal cycles of particle deposition in the
Chesapeake Bay and other coastal systems closely
follow trends in phytoplankton production. Generally,
40-60% of the organic production by phytoplankton
settles to the benthos. Much of the total mass of
deposited material, however, appears to be terrigenous
inorganic sediment, suggesting an important linkage
between biotic processes and sedimentological
transport.
The importance of particle deposition and burial
in the overall input-output balance of nutrients
(nitrogen, phosphorus, and silicon) for the Chesapeake
Bay has recently been questioned. Whereas earlier
reports indicated that most of the nutrient inputs from
the Bay's watershed were retained and buried in its
sediments, a revised analysis indicates that this may
not be the case. This unresolved question has
significant management implications for control of
nutrient wastes.
Sediment oxygen consumption is an important
sink for oxygen pools in the Chesapeake Bay and
other estuaries. Over 50% of the total oxygen
consumption in Bay regions less than 10 m deep
occurs through benthic processes, which remove 10-
20% of the water column oxygen pools per day in late
spring. Nitrogen and phosphorus regenerated by
benthic processes can satisfy about 20-50% of the
phytoplankton demand in the main Bay and 50-100%
in the shallower tributaries. These relative rates are
similar to those reported for other coastal systems.
Recent work in the Bay region has also indicated that
benthic-pelagic interactions may provide a mechanism
for temporary retention of nutrients delivered from the
watershed in winter to support production in spring
and summer.
Research efforts have identified a wide variety of
physical, chemical and biological factors affecting
benthic fluxes of nutrients and oxygen. Temperature
and deposition of paniculate organic matter are
primary variables regulating benthic fluxes. Fluxes of
nitrate and oxygen into sediments are directly related
to concentrations and turbulent mixing in the overly-
ing water. Oxygen concentration (especially as it
approaches zero) also exerts profound, but as yet
poorly described, effects on most benthic processes.
Similarly, burrowing and feeding activities of macro-
fauna appear to significantly influence benthic fluxes,
although exact mechanisms remain to be explained.
Relevance to Restoration and Protection Activities
Many of the processes involved in benthic-
pelagic coupling will directly influence the outcome
of proposed management actions to restore and
protect water quality and living resources in the
Chesapeake Bay. Several relevant questions are given
below.
- Considerable evidence directly connects nutrient
enrichment and increased anoxia in Bay waters. How
is this relationship affected by deposition of plankton
production to the bottom and subsequent oxygen
consumption by benthic processes?
- How important is the nutrient regeneration
capacity stored in Bay sediments, and to what extent
would this result in continued large releases of
nutrients even after inputs from the watershed were
reduced, thereby delaying benefits of clean-up action?
- To what extent has the relative importance of
benthic-pelagic processes such as denitrification and
burial, which act as virtual sinks for nutrient inputs to
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Executive Summary
the Bay, been diminished due to eutrophication, and
will management strategies restore these natural
mechanisms of nutrient waste assimilation to effec-
tively accelerate the benefits of clean-up actions?
- What is the role of benthic-pelagic coupling in
the fate and transport of toxic substances entering the
Bay?
In addition, if water quality modeling is to be a
key component in management of Bay resources, it is
particularly important that benthic-pelagic coupling
processes be sufficiently well described to allow
careful model calibration. The factors regulating
these processes must also be understood and incorpo-
rated into such models to enable accurate projection
and assessment of proposed strategies (e.g., 40%
reduction in nutrient inputs) for restoration and
protection of Bay water quality.
Future Directions
Conceptualization, calibration, and validation of
rigorous mathematical water quality models will
require an expanded scientific information base
including improved descriptions of: (1) seasonal and
spatial trends in benthic fluxes and metabolic pro-
cesses; (2) the relationships among plankton produc-
tion, particle deposition and benthic oxygen and
nutrient fluxes; (3) specific metabolic processes (e.g.,
denitrification and sulfate reduction) that affect
benthic fluxes of oxygen and nutrients; and (4)
physical, chemical, and biological factors that regulate
benthic fluxes. Data should be collected to resolve
questions about Bay-wide budgets of nutrient inputs
and outputs. In addition, the predictive capabilities of
mathematical water quality models could be greatly
supplemented by conducting parallel studies with
estuarine mesocosms ("living models") to assess the
potential outcome of major management decisions.
Ultimately, for these data collection activities to
contribute to the development of effective Bay
management strategies, an administrative mechanism
must be established to communicate scientific
findings to the managers and their constituents.
Conversely, management questions and priorities
must be articulated to the scientific community so that
study designs and approaches can be adjusted to meet
the management needs. The ongoing Bay monitoring t
program may offer a model for efficient bi-directional
communication.
—W. MICHAEL KEMP
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Executive Summary
Summary of Chapter 2
Factors Driving Changes in the Pelagic Trophic Structure of Estuaries,
with Implications for the Chesapeake Bay
Processes governing the relative abundance of
organisms ultimately determine the composition of
biological communities and thus how energy and
material is transferred among components of the
ecosystem. In marine and estuarine environments, the
pelagic food web plays a dominant role in these
transfers. For this reason, changes in the composition
of this trophic network, or in the direction or magni-
tude of flows, could have significant impacts on
fishery yields, water quality, or other factors of
concern to scientists and managers. This chapter
reviews the factors that may influence the composi-
tion of the trophic network.
Summary of Major Findings
The biological success of any organism reflects
both its physiological tolerances and controlling
factors such as food supply. If growth is limited by
quality or quantity of food, then population size may
be regulated by factors controlling prey organisms in
lower trophic levels; this has been termed "bottom-
up" or source control. If nutrition is adequate,
however, population size may be limited by predators
in the next higher trophic level; this is "top-down" or
sink control. The available evidence suggests that
pelagic trophic structure in estuaries is controlled by a
combination of these processes, the relative impor-
tance of which constantly changes in response to
environmental fluctuations.
Estuaries such as the Chesapeake Bay are
characterized by environmental variability of greater
magnitude and frequency than other aquatic habitats;
factors such as light, salinity, and nutrient availability
change over time scales ranging from seconds to
decades or more. This great variability is reflected in
the system's biological components as well. Species
composition and relative abundance typically change
throughout the year, and, although repeating seasonal
signals are strong and often important indicators, year-
to-year variability may be significant. As a general
rule of thumb, the effects of environmental change
can be scaled to body size; that is, the smaller the
organism and the faster its growth rate, the shorter the
relevant time scale for response. Thus changes in
phytoplankton community structure may occur within
a few days, whereas those of fish communities may
extend over years or decades. One of the most
difficult tasks for managers is to separate changes due
to natural variability from those due to human activity
(which are potentially controllable), and to predict
response to natural or anthropogenic environmental
perturbation.
In most estuarine systems, phytoplankton are the
dominant primary producers and the principal source
of food for both the zooplankton-fish trophic chain
and the bacterial-based microbial food web. For this
reason, changes at the phytoplankton level can have
wide-ranging effects on pelagic trophic structure.
Elevated primary production stimulates increased
microbial decomposition, with accompanying
demands on dissolved oxygen. Resulting hypoxia and
anoxia restricts habitat for plankton, fish, and shell-
fish, with consequent mortalities in many species, and
enhances release of nutrients from the sediments.
Reduction in water transparency due to increased
algal biomass has also been linked to recent losses of
submerged vegetation in the Chesapeake Bay.
Long-term changes in the relative dominance of
various phytoplankton groups affect higher trophic
levels and the direction (and efficiency) of transfers
within food webs. Very small cells, which predomi-
nate in the Chesapeake Bay, cannot be readily
ingested by many grazers (such as oysters or cope-
pods). It has been hypothesized that nutrient enrich-
ment may selectively enhance these small forms, and
that this increased dominance by nanoplankton in turn
favors gelatinous zooplankton (comb jellies and
medusae). Because growth efficiencies of these
"jellies" are so low, much of their ingested material is
released as dissolved nutrients. In this respect they
may represent a trophic "dead-end", with little
material being passed on to higher levels. As eutro-
phication increases algal biomass and production
while decreasing its usability by larger zooplankton or
benthic species, more organic material will be cycled
through the microbial loop. There is growing evi-
dence that this process may now be operating in the
Chesapeake Bay, but whether this represents a recent
change in trophic structure is still uncertain.
When food limitation is less significant, predation
may be the major process regulating population
abundance. Predators ranging in size from small
protozoans to large carnivorous fish have specific
prey preferences. Selection can influence the species
composition of prey communities, and thus the struc-
ture of pelagic food webs. Although the importance
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Executive Summary
of this process in marine ecosystems is uncertain, the
concept of cascading trophic interaction as a dominant
force regulating fresh water systems is well estab-
lished. For example, removal of piscivorous fish can
lead to increases in planktivorous species, declines in
zooplankton biomass, enhanced phytoplankton
densities, and decreased water clarity.
Relevance to Restoration and Protection Activities
There is a superficial similarity between this
scenario and that of the Chesapeake Bay. Reduction
in carnivorous species such as striped bass coupled
with apparent increases in plankton-feeders such as
menhaden is suggestive of top-down control. Esti-
mates based on filtering capacity of both menhaden
and benthic suspension feeders such as oysters
indicate that at high population densities these
planktivores can significantly alter phytoplankton
abundance and even species composition. For
example, oyster numbers present in the early 1900's
may have been capable of filtering the entire volume
of Chesapeake Bay in three days. During summer
months, gelatinous predators significantly reduce
zooplankton populations and may also affect larvae of
species such as oysters. Because nutrients released by
these forms can stimulate phytoplankton growth, their
role in structuring pelagic food webs may be consider-
able. Selective predation by both crustacean and
microzooplankton grazers also exerts control on
abundance and composition of lower trophic levels.
Thus each marine pelagic trophic level is impli-
cated in regulating the abundance and structure of one
or more adjacent levels. Whether long-term changes
observed (or postulated) in Bay pelagic communities
are due to natural causes such as climate or to anthro-
pogenic impacts is not readily determined, in part
because of lack of sufficient long-term quantitative
data. Human activities perturb natural controlling
mechanisms from below through nutrient loading or
from above by harvest of predatory or planktivorous
species. It might therefore be hypothesized that
eutrophication and a collapsed predator base are both
driving changes in pelagic trophic structure of the
Chesapeake Bay, and that these processes work in
concert with natural fluctuations in the estuarine
environment to produce observed trends.
Future Directions
These conclusions rely heavily on inference and
emerging new ideas rather than rigorous scientific
examination, and are thus best regarded as hypotheses
in need of further testing. Important questions
include:
- Do long-term changes in concentration and
ratios of nutrient supplies cause changes in phyto-
plankton species composition?
- Are cascading trophic interactions significant in
structuring estuarine communities?
- Does the relative importance of source vs. sink
control vary as a function of productivity?
- In particular, what are the roles of gelatinous
zooplankton, the microbial community, and fish
trophic relationships in structuring the food web?
How do these roles vary in response to nutrient supply
and predator abundance?
Several of these questions represent generic
research needs, but they are particularly applicable to
processes regulating food webs in the Chesapeake
Bay. The biota of the Bay will change as human
beings continue to modify the environment. As
management and restoration of the Bay proceeds, a
unique opportunity exists to study and document the
interactive effects of physical and biological factors
controlling pelagic trophic structure. This will require
a multifaceted approach incorporating data collection
and modeling. Previous studies have demonstrated
that results from one part of the estuary may not be
valid in other portions. Intensive field sampling and
in situ experimentation should be conducted on
pertinent temporal and spatial scales to determine
biological response to environmental variation. Bay-
wide monitoring programs should continue to provide
the long-term data necessary to identify the cause and
significance of trends. Finally, these efforts should be
coupled to multi-trophic level studies (e.g., meso-
cosms) and simulation models to test hypotheses in
detail. This approach, which will require significant
financial and managerial commitment, will quantify
the functional relationships between nutrient supply
and fish production. Thus we will increase our
understanding of the factors that regulate pelagic
trophic structure and improve our ability to predict
changes in that structure.
GAIL MACKlERNAN
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Executive Summary
Summary of Chapter 3:
Physical Processes that Control Circulation and Mixing
in Estuarine Systems
Summary of Major Findings
This chapter surveys the literature on circulation
and mixing as it applies to the Chesapeake Bay. The
approach is to consider processes in terms of their
time scale: seasonal, short-term, or short-period. The
seasonal processes are predominantly long-term
fluctuations (greater than one month) in heating and
fresh-water runoff. The short-term processes (from a
tidal cycle up to one month) include fronts, wind,
tides, and their variations, as well as the interactions
with the shelf circulation. The short-period processes
(less than a tidal period) are the dominant mixing
mechanisms that effect the vertical exchange of
properties in the estuary. There is a tendency for
long-period processes to have large spatial scales and
for short-period events to be smaller. Hence this
organization by time scales also tends to sort by size
as well. Often larger-scale, slower processes are
easier to study, so we have more knowledge about the
mean and seasonal circulation and some fairly good
insight into the short-term processes. The short-
period mixing has only recently been observed, and is
still not quantified.
All the different processes interact. The mean
circulation moves the fresh water down the estuary,
setting up the stratification, which in turn affects the
amount of vertical mixing. The vertical mixing is
driven by the energy that is available from the tides,
wind, internal waves, and mean circulation. The
amount of vertical mixing helps determine the vertical
density profile, which affects the downstream pressure
gradient and hence the intensity of the mean circula-
tion.
Relevance of Restoration and Protection Activities
Surprisingly little work has been performed in the
Chesapeake proper. Although much of the fundamen-
tal work on estuarine circulation is associated with the
Bay, it was predominantly performed in the tribu-
taries. The Bay is large enough to have modes of
motion that are not possible in the more restricted
rivers. Our picture is incomplete, and we must
understand the circulation and mixing in the Bay
better before we can model the system satisfactorily.
Even the classic two-layer circulation is now coming
into question as better instrumentation becomes
available to profile the currents.
Our needs for understanding of the circulation
and mixing of the Bay are rather serious. We have the
basic understanding of the circulation that was
fundamental for the initial planning of the Bay
cleanup. There is also a need to understand the causes
and intensity of the fluctuations in the circulation.
There is some insight into this problem, but we are a
long way from being able to anticipate climatic
variations and predict their effect in the Bay. Finally,
the most stringent demand will be the coupling of the
physical and biological systems to predict the fluctua-
tions in both water quality and biological productivity.
Future research on the physical oceanography in
the Chesapeake Bay must focus on identifying and
understanding the processes that drive the circulation
and mixing. Towards this end two types of work are
needed.
Future Directions
Long-term measurements with modern remote
sensing and profiling instruments are needed. New
techniques for remotely measuring the surface
currents with radar backscatter from shore-based
stations should be combined with new acoustic
profiling current meters mounted on the bottom of the
Bay. This type of information, in conjunction with
satellite remote sensing and data gathered on the many
research cruises in the Bay, could start to quantify the
physical processes over a wide range of scales.
Second, multi-disciplinary studies of specific
processes should be performed. Specific experiments
should focus on distinct problems such as the develop-
ment of anoxia, the interaction of physical processes
and larval recruitment, and the influence of wind on
primary productivity.
Although much past work in this region has
focused on estuarine circulation, there is at present
insufficient work, and insufficient support, for the
basic research into the physical processes in the Bay.
Our ignorance will limit our ability to plan wisely, to
monitor the situation, to interpret the observed
variations, and hence to maximize the return on our
investment in the Bay.
-THOMAS OSBORN
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Executive Summary
Summary of Chapter 4:
Genetics and the Conservation of Estuarine Species
Recent advances in scientific methodology,
especially in the area of genetics, are providing
researchers with the tools to analyze heritability of
traits in greater detail than ever before. New tech-
nologies are also bringing the ability to modify, or
"engineer", the genetic composition of organisms.
These new sources of information are greatly expand-
ing our understanding of the structures of populations,
isolation or mixing of genetic materials, species
complexes, the influence of environment upon genetic
expression, and many other aspects of relationships at
the organism and population level.
This chapter presents a comprehensive review
and synthesis of the emerging place for molecular
genetics as a fundamental tool in the evaluation and
management of biotic systems such as the Chesapeake
Bay. Of particular interest is the discussion of
approaches to conservation and resource management
through the application of genetics. The authors
recognize the potential for protection and enhance-
ment of stocks while citing the need for caution and
consideration of the resulting impacts of such manipu-
lations on the ecosystem. The use of genetic monitor-
ing in hatchery or breeding programs can be effective
in detecting deleterious changes or traits in organisms
being developed for release into natural systems.
The chapter also addresses various aspects of
methods for genetic assessment, including selective
breeding and molecular techniques such as DNA
sequencing and mitochondrial DNA analysis. Brief
descriptions are given of the approach and rationale
for using each of the methods presented.
Summary of Major Findings
The bulk of the chapter concerns the status of
genetic knowledge for selected species of recognized
importance to the Chesapeake Bay. Finfish examples
stress the striped bass; invertebrates selected include
the blue mussel, hard clam, and oyster. Notable by its
absence is the blue crab, for which very little genetic
analysis has been conducted.
The authors conclude that species in the Bay have
apparently undergone some differentiation in response
to varying conditions from the upper reaches to the
Bay mouth but none of the species studied to date
have produced highly localized populations. Thus it
appears that the various species may be treated as
single units for management purposes, and reliance on
captive broodstock for enhancement programs should
probably not introduce defective traits.
The authors suggest approaches for the enhance-
ment of striped bass and oysters while emphasizing
that less-recognized species provide food or other
important ecological linkages.
Relevance to restoration and protection activities
Ultimately, genetic information and the technol-
ogy for manipulating the genetic composition of
organisms will be considered as tools for modifying
the structure of aquatic communities, including the
Chesapeake Bay. In fact, such modification is under
way in the case of the striped bass and its hybrids,
which are being created with either white bass or
white perch.
The potential exists for resource managers to
have both positive and negative effects upon the
structure and functioning of a system such as the
Chesapeake Bay. Because molecular geneticists have
only recently begun to delve into genetic modifica-
tions on a large scale, there is much undiscovered
information concerning the biotic and abiotic interac-
tions of manipulated organisms, especially as they
might affect the stability of an ecosystem. With this
cautionary note in mind, we should nevertheless
continue to explore the potentials presented by this
emerging technology so that the ecosystem may be
better able to withstand perturbations. Questions
include:
- We are not able to determine how the Chesa-
peake Bay is responding to the decline in one of its
top predators, the striped bass. Would the release of
hybrids re-establish the necessary predator-prey
dynamics?
- Would the engineering of an oyster that could
withstand exposure to diseases or chemical pollution
permit us to maintain the necessary links between
primary producers and consumers?
- Would the restoration of oyster populations to
their former levels have a positive impact on removal
of primary producers that use excess nutrients and
apparently drive the spread of areas low in dissolved
oxygen?
The rapid expansion in genetic technology
suggests that within the next decade we will be able to
manipulate the genetic composition of most species
found in the Bay. The manner and degree to which
this capability is applied to managing the dynamics of
the stocks is a matter of growing concern, which
should be addressed in anticipation of the desire to use
the emerging technology. At the very least, fisheries
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Executive Summary
management plans should begin to take into account
the likelihood that genetically manipulated striped
bass and oysters will be available for release into the
Bay.
Necessary research and/or activities
The chapter authors point to the need for a variety
of research efforts as well as genetic monitoring of
hatchery-produced organisms. Also, readers of the
chapter will readily recognize that the body of
literature upon which the synthesis is based is not
extensive. Thus it is obvious that genetic technology
is opening many avenues of research and potential
applications in management. Perhaps some of this
chapter's readers will build upon its information to
enhance their use of genetics in the development of
Chesapeake Bay resources.
—WILLIAM RICKARDS
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Executive Summary
Summary of Chapter 5:
Chemical and Physical Processes Influencing Bioavailability
of Toxics in Estuaries
Summary of Major Findings
This chapter summarizes many of the processes
in estuaries affecting bioavailability of compounds
present in estuaries. Chemical factors discussed
include ionic strength, inorganic and organic specia-
tion and complexation, redox reactions, and produc-
tion of organo-metallic compounds. A second area of
discussion includes contaminant sources and proc-
esses affecting distribution including the microlayer,
adsorption to particles, flocculation, sedimentation,
and remobilization from sediments. A third area is
biological processes affecting bioavailability, includ-
ing uptake routes, food chain magnification, organism
reaction through metallothionen production, biologi-
cal transformation, degradation through biological
processes, and chemical and physical modification of
the environment by biota. The summary delineates
six areas for further research including partitioning,
the role of biota in transfer of pollutants, the effect of
seasonal anoxia, sediment flux and movement, the
role of communities, and descriptive studies of the
effects of toxics.
Relevance to restoration and protection activities.
The subject of toxics in estuaries is controversial
and complicated, and this chapter correctly states the
difficulties of knowing all the factors affecting trans-
port, availability, uptake and impact of pollutants. It
answers in part the question why the many toxic mate-
rials detectable in estuaries are not more effective.
Because the areas of ignorance greatly exceed
those of knowledge at present, restoration and
protection activities concerning toxics in estuaries
must be confined to the few specific cases where toxic
effects have been found. Unfortunately, although
chronic, sublethal, and synergistic effects of toxics on
estuarine biota are suspected to be important, little is
known about them. To be realistic and cost-effective
we must take a scientifically broad approach towards
deciphering the actual role of toxics in estuaries. This
will require basic, biologically-oriented research by
experts, which is not favored by managers but is the
best way to discover the principles permitting effec-
tive detection and control of toxics in estuaries.
Necessary research and/or activities
Questions to be considered in future research
activities include the definition of an estuarine toxic
material. Are estuarine toxic materials those chemi-
cals known to be toxic to humans who come into
contact with them through food or water activities?
Or are estuarine toxics defined through their effects on
estuarine biota? This distinction is critical for
determining the nature of the research and mediating
activities to be done. If the definition is the latter,
then the response of the biota (molluscs to organotin,
fish to creosote, plankton communities to copper, etc.)
must be examined in detail through sophisticated bio-
monitoring. Life cycles must be analyzed to detect
impacts on early life-history stages. Physiological
responses of estuarine animals must be studied to
determine chronic and sublethal effects attributable to
toxics. Monitoring of the health of estuarine biota
must include the incidence of susceptibility to
parasitism and other effects possibly synergistic with
the presence of toxics.
If an estuarine contaminant is defined as that
which is toxic to human beings, this initial research
should establish the routes from the estuary to human
populations and the probabilities of exposure as well
as lexicological effects.
Once a substance has been defined as toxic in
the estuarine system, the chemical and physical
mechanisms governing bioavailability can be re-
searched by using present or as-yet-undeveloped
methods of chemical analysis. Natural mechanisms of
reducing bioavailability and toxicity of the substance,
such as capping of contaminated sediments, must be
studied. Finally, laws should be drawn up and put
into place to control or mediate the toxic. This has
been the successful strategy in the past for handling of
toxics in estuaries.
The full cost and time needed in the past for
discovering and remedying problems due to estuarine
toxics such as power plant chemicals, Kepone, TBT,
and DDT must be examined and studied in detail if we
are to make any serious recommendations about the
handling of future and potential toxics.
With a careful combination of basic and applied
research, admitting our ignorance and working from
historical examples, perhaps we will achieve a
successful strategy for detection and control of toxics
in estuaries, our most impacted ecosystem.
—HARRIETTE PHELPS
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10
Executive Summary
CONCLUSION
In order for research activities in estuarine
processes to contribute to the development of effec-
tive Chesapeake Bay management strategies, the
scientific findings must be communicated to managers
and the public. It is equally important that the
scientific community understand management
priorities and concerns.
The Chesapeake Bay Program is the most ambi-
tious estuarine management program ever attempted.
The Chesapeake Bay system is hydrologically
complex and contains a wide diversity of habitats.
The land uses of the watershed cover the full range of
human activity. In addition, mangement is compli-
cated by different political jurisdictions with differing
philosophical approaches to resource management.
The scientific and management communities in
the Bay region are faced with a strong public and
political desire to "do something about the Bay." To
respond to this desire, the scientific community must
be prepared to address management questions with
state-of-the-knowledge scientific insight. This series
of papers, each in an area directly relevant to pro-
cesses critical to our understanding Cheapeake Bay,
should provide some of that required insight.
-MAURICE P. LYNCH
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Benthic-Pelagic Coupling in the Chesapeake Bay
Jonathan H. Garber
University of Maryland
Center for Estuarine and Environmental Studies
Chesapeake Biological Laboratory
Box 38
Solomons, Maryland 20688
INTRODUCTION
The Relevance of Benthic-Pelagic Processes to Water
Quality Management in the Chesapeake Bay
The purpose of this review is to summarize past
and ongoing research on the coupling of benthic and
pelagic processes in the Chesapeake Bay and to place
these studies in a context of work from other estuarine
and coastal regions. Nutrient management continues
to be one of the most pressing problems for agencies
managing the resources of the Chesapeake Bay [135].
Current conceptual and mathematical models of
estuarine nutrient dynamics, as well as recent monitor-
ing data and experimental studies in estuarine meso-
cosms, suggest that the responses of estuarine
ecosystems to eutrophication and nutrient control
measures are influenced by the exchanges of organic
matter and inorganic nutrients at the sediment-water
interface. The tight coupling of benthic and pelagic
nutrient flows in shallow bays sets the stage for a
positive feedback loop [12, 91] of enhanced primary
production, organic matter sedimentation, and deep
water anoxia that is ultimately driven by an overabun-
dance of inorganic nutrients.
Recent experiments at the Marine Ecosystems
Research Laboratory (MERL) have shown that
sediment-water exchanges of nutrients do indeed
increase with increases in inorganic loading and
primary production, but the increases are not directly
proportional to the loading rate [65]. The results of a
two-dimensional water quality model for the Chesa-
peake Bay [58] indicate that oxygen concentrations in
bottom water could be significantly improved (on the
order of 2 mg/1) if sediment oxygen demand and
benthic nutrient fluxes were reduced by about 30%.
However, the model's ability to simulate sediment-
water exchanges was limited, and the effect of
incremental changes in nutrient loading on water
quality in the Bay remains unclear. There is, in fact,
disagreement over the fundamental behavior of
nutrients in the Bay system [83]. Accurate modeling
and effective management of Bay water quality
therefore depends on understanding the nature of the
exchanges of organic matter, oxygen, and other
constituents across the sediment-water interface.
The recent agreement of the states bordering the
Chesapeake Bay to reduce nutrient loading of the Bay
by 40% by the year 2000 (Draft Chesapeake Bay
Agreement 1987) highlights the Bay managers' need
to know how benthic-pelagic exchanges of oxygen
and nutrients can assist or exacerbate clean-up efforts.
Questions of immediate concern that fall within the
scope of this review include:
(1) What is the quantitative relationship between
nutrient loading of the Bay and sediment-water fluxes
of oxygen and dissolved inorganic nutrients? For
example, does a 40% reduction in allochthonous
nutrient inputs produce an equivalent reduction in
sediment oxygen demand and inorganic nitrogen and
phosphorus release?
(2) What is the eutrophication memory of Bay
sediments? Are they a nearly inexhaustible nutrient
source and oxygen sink that will frustrate Bay clean-
up efforts? Will sediment oxygen demand and
nutrient fluxes respond rapidly (within a few years) or
slowly (within decades or more) to changes in Bay
nutrient loading and productivity?
More specific questions that need to be addressed
to achieve better understanding of the relationship
between sediment-water exchanges and Bay water
quality include:
(1) What is the role of sediment oxygen
demand in the onset and maintenance of deep water
anoxia?
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12
Chapter 1: Benthic-Pelagic Coupling
(2) What environmental factors control rates
of sediment oxygen demand and nutrient releases?
(3) How important are remineralized nutrients in
fueling the cycle of production, organic sedimenta-
tion, and oxygen depletion?
(4) How will increasing bottom water oxygen
concentrations affect rates of nitrogen and phosphorus
release from sediments?
(5) What controls the rates of specific nutrient
transformations, such as nitrification, denitrification,
and phosphorus sorption in Bay sediments? How are
the rates of these reactions linked to observed sedi-
ment-water fluxes?
Historical Overview of Benthic-Pelagic Coupling
The important ecological consequences of
shallowness in marine ecosystems were recognized
over 30 years ago [94]. However, the first explicit
references to the "coupling" of benthic and pelagic
processes appeared in the marine and estuarine
literature in the early and middle 1970's [49, 114].
Rowe et al. [114] claim to have made the "first
successful attempts to measure directly the nutrient
flux from nearshore sediments" in 1975. Hartwig [53]
and Hale [46], however, had completed studies of
sediment-water fluxes near Scripps Pier in California
and in Narrangansett Bay, respectively, the previous
year. Direct measurements of organic matter sedi-
mentation rates and sediment-water oxygen and
nutrient exchanges were also being carried out in
Narragansett Bay [88, 96] and the Patuxent River [16]
during the same period.
These early references to benthic-pelagic cou-
pling illustrate that quantitative work on sediments as
functional components of coastal marine systems
began relatively recently. The impact of these studies
on conceptual models of estuarine nutrient cycles has
nonetheless been dramatic. For example, Pomeroy's
thoughtful and wide-ranging 1970 review entitled
"The Strategy of Mineral Cycling" [107] contained
almost no references to the functional role of sediment
communities in nearshore nutrient dynamics. In
contrast, more recent reviews [66, 85, 87, 89] have
clearly identified nutrient transformations in sedi-
ments as an important feature of coastal ecosystems.
The recent publication of two symposium volumes,
one focused entirely on sediment oxygen demand [54]
and the other reporting European studies of nutrient
cycling and benthic-pelagic coupling [82], attest to the
high level of activity in this field.
Current investigations of sediment-water ex-
changes in coastal waters and estuaries represent the
drawing together of ideas and techniques that have
been developed in fresh water over the last century.
The first step in the progression toward the view of a
body of water as a system with interacting benthic and
pelagic components was the development of tech-
niques to estimate and then to measure directly rates
of oxygen consumption by river muds. Early efforts
to determine "sediment oxygen demand" (SOD) can
be found in the studies of polluted rivers carried out
the late 1800's. These early investigations of sedi-
ment-water interactions, recently reviewed by Davis
and Lathrop-Davis [32], were motivated by very
practical concerns about the effects of sewage on
municipal water supplies. The theme they sound—
that oxygen-consuming processes in sediments
influence the quality of the overlying water—is as
timely in the Chesapeake Bay today as it was in the
Illinois River a hundred years ago.
The first systematic study of the influence of
sediments on overlying waters took place in stratified
lakes [80] where the signal of sediment redox reac-
tions and nutrient exchanges could be clearly followed
in water column profiles. The early work in lakes
showed the relationships between oxygen concentra-
tion, redox potential, and the release of nutrients from
the sediment to the overlying water. Mortimer's
summary [79] of his findings in stratified lakes is
worth repeating here, because very similar events are
now observed in regions of Chesapeake Bay. Mor-
timer [79] concluded that in aquatic systems
... where biological production or organic
pollution is high, where the subthermo-
cline volume is relatively small, or both
... a progressive decline in oxygen
concentration from 2 mg/liter to analyti-
cal zero at the interface was... correlated
with mobilization of and transfer to the
water first of manganese and later of iron.
There is a concurrent transfer into the
water of substantial quantities of phos-
phate, previously held in complex form....
Other changes include liberation into the
water of ammonia and silicate. Further
reduction of the sediment-water system
permits microbial reduction of sulfate.
The interaction between the oxygen content of
overlying water and inorganic nutrient releases from
the sediment can therefore contribute to problems
associated with eutrophication by providing positive
feedback between enhanced organic matter produc-
tion, low oxygen conditions, and release of fertilizing
nutrients from the sediments to the overlying water.
The second stage in the evolution of ideas about
the coupling of benthic and pelagic processes took
place during the mid-1950's with the investigations of
ecosystem-level energy flows in shallow aquatic
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Garber
13
systems. Sediment chambers, or "bell jars", and their
laboratory analogs, initially used to determine rates of
sediment oxygen demand in river sediments [8, 38],
were employed to determine benthic community
production and respiration [23, 90, 103]. Although
clearly related in their need to understand the mecha-
nisms that control the exchanges of materials across
the sediment-water interface, the efforts of water
quality engineers and benthic ecologists proceeded
along parallel but essentially independent lines.
Engineers were concerned primarily with the role of
sediments as an oxygen and pollution sink. Ecologists
focused on the contribution of sediment communities
to the material and energy flow budgets of the
ecosystem. This distinction between the practical and
scientific approaches to sediment-water exchanges
persists today [cf. 35, 116].
The third important theme that weaves through
the development of ideas about the coupling of
benthic and pelagic systems is the powerful insight
that the photosynthetic formation and decomposition
of organic matter in marine systems follows more or
less fixed elemental ratios [29, 41, 108, 109]. This
stoichiometry permits interchangeable estimates of
energy (as oxygen and carbon) and nutrient fluxes
through pelagic and benthic systems; it stands
squarely at the base of sophisticated models of organic
matter diagenesis in sediments [4]. Analysis of
deviations from Redfield's stoichiometries during the
decomposition of organic matter in estuarine sedi-
ments has generated one of the very few explicit and
testable hypotheses to emerge from studies of benthic-
pelagic coupling [85]. Although more than half a
century old, Redfield's Ratio remains one of the
premier heuristic concepts in marine ecology.
Okuda's 1960 monograph [94] on nutrient
dynamics of Matsushima Bay appears to have been
the first quantitative work that clearly reflected the
drawing together of these research themes in the
context of coastal nutrient cycling. Early models of
nutrient cycling in the open sea [130] and estuarine
nutrient budgets [e.g., 52] emphasized the importance
of nutrient regeneration in the water column and
downplayed or ignored the functional role of the
sediments. Okuda recognized the value of using
Redfield-like stoichiometries to analyze the interac-
tions between benthic and pelagic nutrient cycles. He
clearly identified the important consequences of
shallowness in marine systems:
Since the organic matter produced in the
shallow water may reach the bottom in a
shorter time than that in the ocean, the
sediments in shallow water may contain a
greater fraction of undecomposed organic
matter... Accordingly, it may be con-
cluded that although the role of the ocean
sediments in the metabolic circulation of
organic matter is not important from the
quantitative stand point [sic], the sedi-
ments in shallow water may play an
important role qualitatively as well as
quantitatively.
Okuda then demonstrated the functional role of
sediments quantitatively by directly measuring
primary productivity, sediment nutrient profiles, and
rates of decomposition and sedimentation of organic
matter. His analysis of the importance of detrital
exports from seagrass beds and quantitative carbon
and nitrogen budgets clearly points to the importance
of sediment nutrient remineralization in the produc-
tion economy of Matsushima Bay.
Current Models of Benthic-Pelagic Coupling
Today's conceptual model of the interactions
between estuarine water sand sediments (Figure 1)
took shape in the early and middle 1970's, as tradi-
tional descriptions of the seasonal cycles in the
abundances of plankton and dissolved nutrients were
combined with measurements of the rates at which
organic matter and inorganic nutrients were added to,
removed from, and transformed within estuarine
systems. The shift from static to dynamic process-
oriented measurements was driven by a search for
the sources of the nutrients that supported phyto-
plankton production [e.g., 114] and the concurrent
efforts to assemble quantitative estuarine nutrient
budgets [e.g., 85].
Factors of importance in benthic-pelagic coupling
(Figure 1) include: (1) the supply of "new" inorganic
nutrients; (2) the rate at which organic matter is
synthesized in and imported to the water column, i.e.,
the rates of primary production and allochthonous
loading; (3) the rate at which organic matter sinks
through the water column and is deposited as "new"
sediment at the sediment surface; (4) the rates at
which organic matter is consumed by various groups
of organisms in the sediment community; (5) the rates
at which nutrients remineralized in the sediment
community are returned to the water column; and (6)
the rate at which nutrients are permanently buried in
sediment.
The conceptual model (Figure 1) also shows that
in plankton-based systems such as the Chesapeake
Bay, the coupling of pelagic and benthic processes
reflects the spatial separation of predominantly
autotrophic processes in the water column and
heterotrophic processes in the sediments. This does
not imply that heterotrophy does not occur in the
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14
Chapter 1: Benthic-Pelagic Coupling
I Sources of "New" Nutrients
1 ' / Rivers, Runoff, Rainfall, Sewage, Marshes
Remineralized Nutrients
Sediment
Oxygen
Demand
Water
Sediment
Figure 1. Conceptual model of nutrient element flows in plankton-based systems using Odum's energy flow symbols
(modified from Nixon [85]). Pathways relevant to the coupling of benthic and pelagic systems are identified, with pertinent
review papers, as follows: (1) inputs of nutrients from outside sources [33, 84, 87]; (2) uptake of nutrients by phytoplankton
(not considered in this review); (3) sinking and deposition of particulate matter [6, 7, 19, 20, 78]; (4) burial and long-term
diagenesis of organic matter [3, 55]; (5) oxygen consumption and other catabolic pathways [82, 102] (6) exchanges of
oxygen, dissolved nutrients and other constitutents across the sediment-water interface [85, 86, 89, 140]; (7) upward diffusion
and advection of remineralized nutrients (not considered in this review).
water column or autotrophy in the sediments. On the
contrary, there is growing evidence that microhetero-
trophic "microbial loops" are responsible for very
rapid cycling of nutrients in the water column [2,
136]. Autotrophy at the sediment surface by diatom
mats, seaweeds, and seagrasses may also flourish
where sufficient light reaches the bottom. However,
there is generally a net production of organic matter in
the euphotic zone of the water column and a net
consumption of organic matter in the sediments.
As illustrated in Figure 1, the cycling of materials
between Bay waters and sediments begins with the
formation of organic matter and its deposition on the
Bay bottom. Plankton production is the primary
source of autochthonous (formed in situ) material;
allochthonous (introduced from outside) sources of
sediment include terrigenous material carried to the
Bay by rivers, runoff, rain, wind, and shore erosion.
Thorough discussions of plankton production in the
Chesapeake Bay and particle sinking are beyond the
scope of this review, and the reader is directed to the
reviews of these topics cited in Figure 1.
The material deposited at the surface of estuarine
sediments comprises a complex mixture of organic
matter and inorganic minerals. Although inorganic
phases are unquestionably involved in early sediment
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Garber
15
diagenesis, the decomposition of organic matter is
primarily responsible for the fluxes of oxygen and
other constituents at the sediment-water interface.
Organic matter is consumed in sediments via a
sequence of energy-yielding biochemical reactions in
which oxygen, nitrate, manganese, iron, and sulfate
serve as progressively less efficient oxidizing agents.
Sediment macrofauna may be the most conspicuous
aerobic consumers of organic matter, but their
contribution to the total metabolism of benthic
communities is generally believed to be small [15].
More important to the sediment community metabo-
lism is a diverse assemblage of bacteria and other
microheterotrophs. Simplified equations for the
microbially-mediated reactions of particular impor-
tance in estuarine sediments [3] are:
Aerobic metabolism
CH2O + O2 -
Nitrate reduction (denitrification)
5CH2O + 4N03-
CO2 + H2O
Sulfate reduction
2CH2O + SO4=
2N2 + 4HCO3- + C02 + 3 H2O
Ammonium formation
CH2NH2COOH + 3(H+)
H2S + 2HCO3-
Methanogenesis
C02 + 8(H+)
CH3COOH + NH4+
CH4 + 2H20
The stoichiometry of nutrient remineralization
during decomposition of organic matter in anoxic
portions of estuarine sediments can also be described
by a Redfield-like equation [1 1 1]:
6(CH20)x (NH3)y (H3P04)z + 3xS04- — >
6XHCO3- + 3xHS- + 6 NH4+
These reactions result in the consumption of
oxygen, nitrate, and sulfate sequentially with depth in
the sediment column. The reaction products —
carbonate, di-nitrogen gas, hydrogen sulfide, ammo-
nium, phosphate, and methane — accumulate in
sediment pore waters, often in concentrations that far
exceed those of the overlying water. Such accumula-
tions are well documented in sediments from the
Chesapeake Bay [18, 77] and elsewhere. The deple-
tion of reactants and accumulation of reaction
products drives a flux of dissolved constituents across
the sediment-water interface. The rates at which
constituents cross the sediment-water interface is
influenced by many factors including the availability
of reactants, the steepness of concentration gradient
between the sediment and water, water currents over
the sediments, physical resuspension of the sediments,
sediment mixing by macrofauna (bioturbation),
advection of pore waters by benthic fauna (biopump-
ing), sediment compaction, and density-driven
flushing of pore waters [3, 126].
The linkage or "coupling" between Bay waters
and sediments can therefore be seen to consist of the
downward flux of organic matter from the water to the
sediment and the return flux of remineralized nutrients
from the sediments to the water. That these ex-
changes should be taking place has been known for a
very long time, and recent attention given them hardly
signifies a revolution in ecological thinking. What is
new, and what has become clearer over the past two
decades, is that Okuda [94] was right: these ex-
changes are both quantitatively and qualitatively
important in estuarine nutrient cycles. Nixon [85], for
example, demonstrated that bottom communities,
including two sites from the Patuxent River, consume
25-50% of the total organic matter produced in or
imported to the system. Studies in a variety of coastal
systems (reviewed in [85] and [140]) show that
nutrients released from coastal sediments can supply a
significant fraction, at times all, of the nutrients
required for annual phytoplankton production.
The cycling of nutrients through the estuarine
benthos may also profoundly affect estuarine produc-
tivity by altering the relative abundances of remineral-
ized nitrogen and phosphorus. In his analysis of the
Narragansett Bay nutrient budgets, Nixon [85] stated:
The remarkable conclusion seems to be
that much of the nitrogen limitation of
primary production in coastal waters
arises because of low inorganic N/P
ratios brought about because some 25-50
percent of the organic matter fixed in
these areas is remineralized on the
bottom....[where] rates of denitrifica-
tion....remove some 5-25 percent of the
nitrogen orginally incorporated into
organic matter.
Although couched as a conclusion, Nixon's
statement [see also 120, 121] has been taken as an
explicit hypothesis that denitrification in the sedi-
ments is responsible for the nitrogen limitation of
production in coastal marine systems.
TECHNIQUES USED TO MEASURE EXCHANGES
OF MATERIALS BETWEEN
ESTUARINE WATERS AND SEDIMENTS
The Downward Flux of Particulate Matter
The rates at which sediments are deposited on the
Bay bottom have been estimated in three ways:
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16
Chapter 1: Benthic-Pelagic Coupling
(1) sediment mass transports [5,10,113, 118, 129];
(2) the distribution of particle-active radionucb'des
such as 210Pb, 137Cs, 239Pu,240Pu, ^Th, and7Be with
depth in sediment cores [45, 56, 57,92] and (3)
sediment traps [12, 67]).
The time-resolutions of these techniques differ,
and each is subject to its own set of technical limita-
tions. Sediment mass balances integrate sediment
inputs from various sources over a year or several
years, geochemical tracers provide estimates of sedi-
ment deposition over periods of weeks (7Be, 234Th) to
decades (""Pb and 137Cs); and sediment traps intercept
the downward flux of materials over periods of hours
to days. Mass balance approaches require intensive
sampling of river-borne and aeolian fluxes, may
underestimate episodic events such as floods, and are
very sensitive to estimates of the area where active
deposition is believed to take place [84]. Geochemi-
cal tracers are better at integrating short-term episodic
depositional events into a longer-term sedimentation
rate, but interpretation of isotope distributions in cores
is complicated by bioturbation and non-uniform sedi-
ment deposition [92, 110]. Sediment traps, although
the best method for obtaining direct measurements of
short-term deposition, are subject to design and
deployment problems [for reviews see 6, 7,15, 20,
70], and collect a combination of resuspended and
newly deposited material [67,96].
Exchanges at the Sediment-Water Interface
Indirect techniques. The magnitude of fluxes
across the sediment-water interface can be estimated
indirectly, by modeling the gradient of pore water
constituents near the sediment-water interface [3, 4,
73], or directly, by determining the rate of change of
dissolved constituents in the water overlying the
sediments [53, 88, 114, 140]. The modeling approach,
essentially an application of Pick's first law of dif-
fusion, requires accurate measurement of the concen-
tration gradient and a good estimate of the constituent
"diffusion coefficient" near the sediment-water
interface [3]. This approach has therefore proved
most useful in totally anoxic sediments and in the
deep sea where sedimentation rates are low, physical
mixing of the sediment and bioturbation are minimal,
and deployment of in-situ chambers is impractical.
Comparisons of fluxes determined by pore water
modeling and intact sediment methods (described
next) show that the modeling technique works best
when fluxes at the sediment-water interface are
dominated by diffusion [39]. The technique under-
estimates fluxes in situations where waves, tides, and
benthic fauna stir the sediments and transport con-
stituents across the interface [1, 22, 39, and others].
Direct in-situ and intact core techniques. Benthic
fluxes can also be determined directly by various
adaptations of "bell jar" chamber procedures [140]
and the closely related "upstream-downstream"
techniques [40, 61]. These techniques involve
determining the change in concentration of dissolved
constituents in a known volume of water trapped
above or passing over a given area of the bottom. The
flux of a dissolved constituent across the sediment-
water interface in sediment chamber devices is then
calculated using some form of the basic equation:
Ji =
V (Co - CO
AT
where J is the flux of constituent i in mols m~2 time'1,
V is the volume of water contained in the chamber
(liters), Co and Ct are the initial and final concentra-
tions of constituent i (mols/liter) during an incubation
of time T, and A is the area of bottom enclosed by the
chamber (m2).
In the absence of a standard method for carrying
out direct flux determinations, a great variety of
coring devices, chambers, tunnels, and free vehicles
have been developed for in-situ, shipboard, and
laboratory applications of the technique [54, 101,
140]. The advantage of this approach, in theory at
least, is that it provides a better measure of fluxes in
an intact sediment community, because chemical and
biological contributions to the net fluxes are accu-
rately reflected in the changes in concentration
observed in the water. The simplicity and relative
ease of this technique give it great appeal, and its
application in all types of aquatic systems continues to
grow. It must be noted, however, that this approach is
not trouble-free. Fluxes determined in chambers are
affected by disturbance of sediment surface during
coring, transportation, and chamber deployments [64,
81], rates of water circulation [9, 15, 24, 76] and the
associated boundary-layer effects [62], and ground-
water flows [141]. Chemical and biological ex-
changes of dissolved nutrients generally increase as
current velocity over the sediment increases, rising
dramatically when current velocities are high enough
to resuspend the sediment [9, 15, 59]. The problem of
circulation in flux chambers is difficult to evaluate
against the background of considerable spatial and
temporal variability encountered in estuaries. It is, for
example, very difficult to simulate the "correct"
current speed during a 3-4 hour chamber experiment,
in regions where near-bottom currents vary from zero
to some tidal maximum.
Recently, Pamatmat [99] took issue with the
uncritical use of oxygen flux determinations as
indicators of "total sediment metabolism" and
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Garber
17
advocated the direct measurement of heat production
by sediment communities as a better measure of total
energy flow through the benthos [100,102,104].
Although conceptually appealing, the technique
appears fraught with technical difficulties, and with
few exceptions [72, 133], direct calorimetry has been
of limited use in benthic flux studies.
Microcosms and mesocosms. In addition to the
empirical appoaches to the study of benthic fluxes
described above, the exchanges of materials between
estuarine waters and sediments have been examined in
sediment-water microcosms ranging in size from
small cores [42,43, 63, 64] to the large mesocosm
facilities in Kiel Bight [126] and the MERL tanks at
the University of Rhode Island [65, 87,95,98,115].
In contrast with the in-situ and most intact-core
approaches described above, these experimental
systems are primarily designed to examine how
specific environmental factors or pollutants affect
sediment-water exchanges. They permit a degree of
environmental control and replicability not possible in
field studies. The MERL tanks are particularly
noteworthy because they were explicitly designed to
facilitate the study of benthic-pelagic coupling: the
sediment communities in the tanks can be isolated
from the water column for benthic flux studies.
STUDIES OF SEDIMENT DEPOSITION AND
BENTHIC NUTRIENT
FLUXES IN THE CHESAPEAKE BAY
The Downward Flux of Particulate Matter
Annual rates and composition. Nixon [83]
recently summarized estimates of the rates and
patterns of total sediment (i.e., inorganic plus organic
material) deposition in the Chesapeake Bay. These
geological and geochemical studies have indicated
that nearly all sediment entering the Bay from rivers
and offshore is deposited within the Bay system [5,
113, 118]. Rates of total sediment deposition decrease
from north to south along the main axis of the Bay,
then increase again near the bay mouth [92]. Highest
rates range from 1-5 kg dry wt nv2y' in the northern
reaches of the Bay to about an order of magnitude less
in the lower reaches. Estimates of annual sediment
accumulation (dry weight) averaged over the area of
the entire Bay range between 0.2 and 2 kg nr2 [5, 83,
92,118].
Bay-wide budgets for nitrogen and phosphorus
assembled during the first phase of the EPA-spon-
sored Chesapeake Bay Program [129] suggested that,
like total sediment, essentially all the nitrogen and
phosphorus entering the Bay from various sources is
deposited on the Bay bottom. If this is true (and for
the moment ignoring regional differences in deposi-
tion rates), ballpark estimates of the amounts of
nitrogen and phosphorus deposited on the Bay bottom
annually can be made by dividing the total amounts of
nitrogen (160 x 10* kg/year) and phosphorus (30.3 x
Wkg/ year) entering the bay [129] by the area of Bay
bottom (about 11 x 109 m2, including tributaries [30]).
These calculations suggest that Bay-wide annual
average deposition rates of nitrogen and phosphorus
should be on the order of 14 g/m2 for nitrogen (1.0 g-
at nr2 and 1.2 gnr2 for phosphorus (0.39 g-at nv2).
Similarly, the composition of paniculate material
deposited in the Bay should approximate the inputs
divided by the total sediment load (3010 x l^kg nv2
[129]). Material deposited on the Bay bottom, if it
were unaltered by any biological or chemical proc-
esses, would therefore be expected to contain 5.3%
nitrogen and 0.46% phosphorus by weight, and have
an atomic N:P ratio of about 25.
It is now possible to compare the estimates of
long-term sediment deposition with the short-term
direct measurements of sediment deposition rates
provided by sediment trap experiments. The only
sediment trap data for the mainstem Chesapeake Bay
available for this review were those provided by the
Ecosystem Processes Component (EPC) of the
Maryland Chesapeake Bay Monitoring Program ([12],
W.R. Boynton et al., unpublished data); for an
overview of the Maryland Monitoring Program see
[74]. Sediment trap studies have also been carried out
in some tributary estuaries including the Patuxent [17,
67], the Potomac (D. Shultz, pers. comm.), and
Choptank River estuaries [139].
Mainstem Bay studies. Annual rates of the
downward flux of sediment determined with sediment
trap arrays located in the mainstem Bay at buoy R64
(Figure 2) by Boynton et al. [12 and unpublished data]
were remarkably uniform over the period of July 1984
to June 1987. Time-weighted annual sedimentation
rates of total seston to the traps located at a depth of
4-5 m averaged 1.70 ± 0.32 kg nr2 (mean ± SD).
These rates are consistent with the estimate of Officer
et al. [92] for the mid-Bay region, of 1.6 kg nr2 (dry
weight). The shallowest traps were chosen for this
comparison because they best integrate the amounts of
material sinking through the euphotic zone [51, 122],
and are likely to be the least biased by resuspended
sediment. Traps located at mid-depths in the water
column generally collected two to three times more
material than the shallowest traps. The uncertainty in
the net sedimentation rate determined with sediment
traps is probably less than a factor of 2 or 3, well
within the ranges of sedimentation rates found with
-------
18
Chapter 1: Benthic-Pelagic Coupling
various transport and tracer approaches. The results
of the trap studies also indicate that potentially labile
components of the sedimenting particles (carbon,
nitrogen, and phosphorus) comprised only 5-15% of
the total mass of paniculate matter sinking through the
water column. Therefore, even if these labile con-
stituents were totally remineralized before the
particles reached the sediment surface, the sedimenta-
tion rates given by the traps would be high by only
about 15%.
The annual flux of sediment collected in the near-
surface traps of an array maintained in a 10-m water
column on the Bay's mid-western flank (Station DB,
Boynton et al., unpublished data) was four to five
times greater than that found in deeper water near the
central channel. The higher sedimentation rate found
on the western flank can be attributed in part to
sediment resuspension. In contrast with the material
collected in deeper water, the elemental composition
of the paniculate matter collected on the flanks was
fairly constant throughout the water column and more
closely resembled that of near-surface sediments in
the mid-Bay region (2-4% carbon, 0.2-0.5% nitrogen,
0.05-0.1% phosphorus) [12]. On the other hand,
enhanced deposition of organic matter on the flanks of
the Bay is consistent with recent findings of extraordi-
narily high summer chlorophyll levels and phyto-
plankton productivity on the Bay flanks [75, 124]. It
is particularly intriguing that the annual deposition of
carbon into the 3-m trap of some 400 g m2 is con-
sistent with daily net plankton production rates of
3-5 g m'2 found on the flanks in summer [75, 124].
The evidence from the mid-Bay sediment trap deploy-
ments shows that the rates of nitrogen and phosphorus
deposition determined with sediment traps in the mid-
Bay region are several times greater than would be
predicted if all sediments and nutrients entering the
Bay were deposited evenly over the Bay bottom. The
simple deposition calculation made above also
suggests that the nitrogen and phosphorus composi-
tion of sedimenting material should be about 5% and
0.5% by weight, respectively. The nitrogen and
phosphorus content of paniculate material collected
in the traps, as a percentage of total material collected,
is considerably less, averaging 1-2% nitrogen and
0.15-0.36% phosphorus. The N:P ratios averaged
about 12, or about half the ratio predicted by the
simple mass balance calculation.
Nixon [83] recently noted similar discrepancies
between the expected and observed composition of
Bay sediments and used these to challenge current
interpretations of nutrient budgets assembled during
the first seven-year phase of EPA's Chesapeake Bay
Program. These budgets were initially interpreted as
showing that sediments and nutrients behave simi-
larly, i.e., "nearly all of the materials that enter the
Bay remain there; nutrients trickle out of the Bay
mouth at a very slow rate" [129]. Nixon has carefully
constructed arguments using much of the same data to
arrive at exactly the opposite conclusion. According
to Nixon's calculations, only a small fraction of the
nitrogen and phosphorus introduced into the Bay
during non-flood years accumulates in the sediments.
A fundamental disagreement therefore remains about
the overall behavior of nutrients in the Bay. The
sediment trap data support Nixon's interpretations, at
least as far as indicating that the nitrogen and phos-
phorus composition of sedimenting paniculate matter
in the mid-mesohaline region of the Bay is considera-
bly less than would be expected if nitrogen and
phosphorus were being quantitatively removed from
Bay water by sediment formation. Whether similar
patterns of sediment and nutrient deposition hold for
other regions of the mainstem Bay and tributary
estuaries is not known.
Tributary estuaries. Boynton et al. [17] and
Kemp and Boynton [67] described a series of sedi-
ment trap deployments carried out between August
1979 and July 1980 at five locations along the salinity
gradient in the Patuxent River estuary. These studies
indicated that net sediment deposition was highest in
the low-salinity (0-5 ppt) regions of the estuary and
decreased toward the mouth of the estuary. These
authors suggest that differences in the timing of
maximum dissolved nutrient inputs (winter and
spring) and phytoplankton production (late spring and
summer) result in a cycle of sediment formation,
deposition, and nutrient remineralization that causes
spatial and temporal separation of "new" and "regen-
erated" production in the estuary [37]. Their results
suggest that nutrient deposition should be especially
high in the transition zone between well-mixed, low-
salinity water and the downstream region of two-layer
estuarine circulation.
A budget for total sediment transports and
deposition in the Choptank River estuary [139]
demonstrated that nearly 20% of the sediment carried
into the estuary was exported to the mainstem Bay.
This fraction seems surprisingly large in view of the
common notion that estuaries act as efficient sediment
and nutrient traps. These results from the Choptank
indicate that sediments derived from tributary basins,
and presumably the paniculate nutrients associated
with them, are not necessarily deposited quantitatively
in the Chesapeake's sub-estuaries.
Thus, although geological and geochemical
evidence suggests that the inorganic component of the
sediment load entering the Chesapeake Bay is
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Garber
19
deposited and buried in Bay sediments, it is far from
clear that nutrients behave similarly. Recent mass
balance and sediment trap studies suggest that only a
fraction of the nitrogen and phosphorus introduced
into the Bay is permanently buried in the sediment.
Seasonal patterns of sediment deposition. Sediment
resuspension unquestionably adds an as yet unresolv-
able dimension to sediment trap data from shallow
systems [51, 96]. Nevertheless, sediment fluxes
recorded by the trap arrays in the mainstem Bay [12]
followed seasonal cycles that are ecologically inter-
pretable and, except for the near-bottom traps, do not
appear to be so completely biased by sediment
resuspension that the signal from the deposition of
new material is lost The three-year record [12] of
paniculate carbon deposition to the near-surface traps
showed that the period from March to October could
be divided into three sedimentation regimes. The
spring period began in March, or perhaps earlier,
when sedimentation rates increased; they peaked in
late April or May at about 1 g C nr2 daily. This ex-
tended depositional event was followed by a steep
decline in May. The summer regime was character-
ized by an overall trend of increasing deposition rates
punctuated with strong depositional events of short
duration (1-2 weeks). A third period of high deposi-
tion occurred in mid- to late fall (October-November).
The deposition of total seston, paniculate nitrogen,
paniculate phosphorus, and total chlorophyll followed
patterns nearly identical to those of paniculate carbon
[12]. The seasonal pattern of paniculate carbon
deposition observed in shallower water on the western
flank followed a similar overall pattern, but peak
deposition rates were as much as two- to fivefold
greater than those found in deeper water.
Year-to-year differences occur in the magnitudes
and timing of major depositional events, particularly
during summer, but the overall pattern of sediment
deposition to traps in the mid-mainstem Bay is
consistent with seasonal patterns of integrated area-
based primary production for the Maryland Monitor-
ing Station MCB4.3C [123], which is located near the
trap array at station R64 of Boynton et al. The high
and sustained deposition rates observed in spring can
therefore be attributed to the "sink-out" of the spring
phytoplankton bloom, and the peaks of carbon and
other nutrient deposition observed in summer and fall
can be attributed to the periodic but irregularly timed
bloom-and-crash nature of summer phytoplankton
blooms. This seasonal pattern of sediment deposition
in the Chesapeake Bay differs somewhat from that
found in other coastal waters [31, 51, 105, 127] in
showing the effect of the spring diatom bloom and the
long period of high summer production rates on the
sedimentation of organic matter. Deposition in other
systems, such as Kiel Bight, is even more strongly
seasonally pulsed, with the major depositional event
of the year taking place in the spring. One of the
important implications of these findings is that a large
fraction of the spring bloom may not be grazed by
zooplankton and may sink to the bottom directly as
intact cells. The sinking of the bloom therefore
represents an important source of high-quality organic
matter to the benthos, and fuels high rates of sediment
oxygen demand in deep waters and the sediment as
the season progresses into summer.
A final point demonstrated by the results of
recent sediment trap deployments in the Bay is that
the composition of inorganic and organic paniculate
material suspended in the water column generally
shows little resemblance to the paniculate matter
collected in traps, and the concentration of suspended
seston does not provide a good indication of the
amount of material sinking through the water. The
theoretical basis for this difference is now well known
[78]: large, fast-sinking particles are relatively rare in
the water column and are therefore not adequately
sampled by water bottles and low-volume pumps. It
is therefore not surprising to find differences in the
elemental compositions of suspended paniculate
matter (sampled with bottles or pumps) and the
sinking particles collected in sediment traps. Func-
tionally, these represent two rather distinct popula-
tions of particles. Boynton et al. [12], for example,
found that the atomic C:N:P ratios for suspended
paniculate matter at station R64 averaged 250: 24:1
during two years of observations. During the same
period, the composition of material collected in the
sediment traps located within or just below the
seasonal oxycline (about 9 m) averaged 95: 12:1. In
terms of both the C:N ratio and N:P ratio, the material
collected in the traps represented significantly higher-
quality organic matter than that suspended in the
water.
Factors influencing the downward flux of particu-
late organic matter. Systematic analysis of physical
and biological factors that contribute to seasonal and
shorter-term rates of sediment formation, deposition,
and resuspension is not available for any region of the
mainstem Chesapeake Bay. Studies from other
coastal areas [48,96,105, 127] suggest that physical
factors including currents, wave action, and water
column stratification influence rates of sedimentation
and sediment resuspension. The relative importance
of sediment that originates from aeolian and other
atmospheric sources is also not known. There are no
reports of attempts to determine the origin of panicu-
late material collected in the recent sediment trap
-------
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Garber
21
deployments, or to quantify the relative abundances of
sedimenting particles that entered the water column
from either the air-water or sediment-water interface.
As in other coastal plankton-based systems, there
appears to be some coupling between plankton
production and the deposition of organic carbon in the
mainstem Bay. Carbon budgets assembled by Kemp
and Boynton [68] for the water column and benthos of
the western shore communities near Calvert Cliffs
have suggested that 20-40% of annual plankton
production reaches the Bay benthos. These estimates
are consistent with Nixon's [85] correlation between
sediment community metabolism and organic matter
input rates if, as shown by Kemp and Boynton's
budgets, nearly all the carbon deposited on the bottom
is respired by the bottom community.
Another factor that has not yet been adequately
evaluated in the Bay is whether the deposition of
paniculate matter could be significantly altered by the
abundance of filter-feeding benthic animals, as has
been suggested for San Francisco Bay [28] and other
systems [93].
Sediment Oxygen Uptake and Nutrient Fluxes
Identification and description of studies from the
Bay region. A search of the primary and grey litera-
ture turned up some two dozen reports that deal at
some level of detail with sediment-water exchanges of
oxygen or nutrients in the Chesapeake Bay and its
tributaries (Table 1). These reports are of three types:
(1) characterizations of spatial and temporal variabil-
ity in pore water profiles [e.g., 18,77, 129]; (2) direct
determinations of sediment-water fluxes using in-situ
or shipboard benthic chambers [10, 12, 22]; and (3)
laboratory studies of the effects of various environ-
mental conditions on rates of sediment oxygen
demand and dissolved nutrient exchanges [e.g., 25,
26,119,138].
NOTES for Table 1 (facing page): AEC = Atomic Energy
Commission; BC = benthic chamber; DIP = dissolved
inorganic phosphorus; DON = dissolved organic nitrogen;
DOP = dissolved organic phosphorus; EPA = Environ-
mental Protection Agency; LC = laboratory core; MdDNR =
Maryland Department of Natural Resources; MdOEP =
Maryland Office of Environmental Protection; MdPPSP =
Maryland Power Plant Siting Program; MdSGC = Maryland
Sea Grant College; MWCOG = Maryland-Washington
Council of Governments; NG = not given; N + N = nitrate =
nitrite; NSF = National Science Foundation; PW = pore-
water profiles; RRDPC = Richmond Regional District
Planning Committee; USGS = United States Geological
Service.
The early studies of sediment pore waters
indicated that the observed profiles could support
fluxes of ammonium and phosphate to the overlying
water but did not report such calculations per se
[18, 79]. Fluxes of ammonium and phosphate based
on sediment profiles were, however, combined with a
limited suite of in-situ chamber determinations of
ammonium and phosphate fluxes [131] and used by
Smullen et al. [129] to estimate the contribution of
sediment-water exchanges to the Bay's nutrient
budgets.
The compilation of reports presented in Table 1
revealed two important features of the benthic flux
work that has been carried out in the Chesapeake
Bay. First, in terms of both techniques and insights
into fundamental controlling mechanisms, research on
sediment-water exchanges in Chesapeake Bay began
relatively recently. Benthic oxygen fluxes in the
Chesapeake Bay were first directly measured a little
over 10 years ago [14] near the Calvert Cliffs nuclear
power plant, and the first in-situ determinations of
sediment-water nutrient fluxes (again in relation to
power plant operations), were done about a year later
near the Chalk Point power plant in the low-salinity
region of the Patuxent River estuary [16]. As men-
tioned in the Introduction, similar work had been
underway in other coastal and estuarine regions in the
early 1970's.
The second feature emerging from the summary
in Table 1 is that benthic flux work in the Chesapeake
Bay was initially supported almost entirely by
"mission-oriented" sponsors, particularly those
concerned with assessing the environmental inpacts of
power plant operations. Direct determinations of
nutrient fluxes across the sediment-water interface
was apparently not perceived as a major research
objective during the first phase of the Chesapeake Bay
Program. As a consequence, the nutrient-sediment
flux data reported in the Program Synthesis by Taft
[131] and Smullen et al. [129] were, by the authors'
own admission, very limited. The temporal and
spatial design of these studies often makes little
ecological sense, and much of the data generated are
available only in grey literature reports. Of the studies
listed in Table 1, fewer than half have appeared in
peer-reviewed journals or books. Only four areas (the
western flank of the mainstem Bay near Calvert Cliffs
[17], the Patuxent [67], a shoal near Gloucester Point
in the York estuary [112], and a site near Horn Point
in the Choptank [128]) were sampled often enough
over an annual cycle to permit some reasonable
estimate of annually-integrated sediment flux rates.
When the locations of study sites gleaned from
the works listed in Table 1 are brought together
-------
22
Chapter 1: Benthic-Pelagic Coupling
77-00-
76-30-
75°XX>/
Y
_ Chesapeake - Bay ^
Region
77-00'
76° 30-
76! 00*
75-30'
Ffgure 2. Map of the Chesapeake Bay showing locations of sediment-flux studies (letters) and sediment trap deployments
(asterisks). Letters correspond to studies listed in Table 1.
-------
Garber
23
(Figure 2), it becomes clear that work on sediment-
water exchanges in the Chesapeake system has been
concentrated in the Maryland portion of the mainstem
Bay and in the Patuxent, Potomac, and the upper
reaches of the James River tributaries. Studies along
the mainstem are almost entirely restricted to the
relatively deep mid-Bay channel. No data on sedi-
ment-water fluxes are available for the Rappahan-
nock, York, lower James, and eastern shore tributaries
(except the Choptank) or for most of the lower main-
stem Bay or shoal regions throughout the Bay. It is
also important to note that coverage of the sites shown
in Figure 3 is far from even. Most studies report
fewer than four measurements per year, and very few
report fluxes for oxygen and a complete suite of
inorganic nutrients.
Spatial and seasonal patterns of sediment-water
fluxes in the Patuxent River and mid-mainstem Bay.
Two regions of the Bay for which substantial amounts
of sediment flux data are available are the Patuxent
River estuary [17, 67, 71] and the mid-mainstem Bay
[10], especially the portion north of the Maryland-
Virginia state line [12].
Patuxent River Estuary. Lantrip et al. [71]
reported the results of benthic chamber determinations
carried out more or less seasonally for two years at 11
stations along the length of the tidal-fresh and
estuarine portions of the Patuxent River. In contrast
with earlier work [17, 67, 69] which indicated rather
clear temporal and spatial patterns in sediment fluxes
along the estuarine portion of the Patuxent, the
preliminary analysis of these data by Lantrip et al.
[71] failed to reveal any consistent patterns in the flux
data. Data from the two OEP monitoring stations in
the Patuxent [12] also exhibit considerably less
variability and fall squarely on the means of the fluxes
reported by Lantrip et al. [71]. In both of these
studies, however, sampling designs were heavily
weighted toward the spring and summer months. The
lack of measurements throughout the winter may
account for the overall lack of a temperature signal in
the flux data. Skewing the flux measurements toward
the summer months also obviously biases the overall
mean, as summer fluxes may well be higher and more
variable than other seasons (see, e.g., [88]).
Mid-Mainstem Bay. Longitudinal transects
summarizing spatial and seasonal patterns of benthic
oxygen and nutrient fluxes along the axis of the
mainstem Bay are shown in Figure 3 (overleaf).
These figures were assembled by comparing data for
May and August for stations located along the main
axis of the Bay [12, 67] with data for the shallow
mainstem stations near Calvert Cliffs [17]. Variability
of the mainstem flux data has been minimized by
showing standard deviations around each mean.
These data from the mainstem Bay point to significant
seasonal and spatial patterns in nutrient fluxes.
Benthic oxygen fluxes were relatively high near the
head of the Bay in the spring and summer. The spike
in oxygen fluxes in the mid-Bay region was attribut-
able to the shallow stations near Calvert Cliffs in May
and the deeper mainstem stations in August. Variabil-
ity in the mid-Bay oxygen fluxes increased in August
as bottom waters alternated between oxic and hypoxic
conditions. Almost no spatial pattern was evident in
nutrient fluxes along the main axis of the Bay in May.
In summer, however, fluxes of ammonium, phosphate,
and silicate increased dramatically in the mid-Bay
region, probably in response to the combination of
high temperatures and low oxygen in the bottom
water. Seasonal differences in nitrate + nitrite (N+N)
fluxes were also apparent, shifting from Bay-wide
sediment uptake in May to strong release of nitrate in
the upper Bay in August.
THE NATURE OF BENTHIC-PELAGIC COUPLING
IN THE CHESAPEAKE BAY
Factors That Influence Rates of Sediment-Water
Exchanges
The exchanges of oxygen and inorganic nutrients
between sediments and waters in the Chesapeake Bay
appear to vary with time and space in ways that defy
simple explanation. Understanding the sources of this
variability therefore remains an important research
goal. The approach most often taken has been to
search for correlations between benthic fluxes of
chemical constituents and various environmental
parameters, such as water temperature, depth, sedi-
ment characteristics, mixed layer depth, and rates of
organic loading [47, 50, 65, 86, 88]. Although this
approach may eventually lead to an empirical model
of sediment-water nutrient exchanges in the Bay
region, it sheds little light on the mechanisms respon-
sible for the observed correlations.
Water temperature and water column stratification.
An early summary of sediment oxygen fluxes from a
variety of aquatic systems [50] suggested that the flux
of oxygen across the sediment-water interface was
correlated with temperature. As more data from
marine systems became available, the model was
modified [49] to show that annual rates of sediment
oxygen uptake could be related to a parameter arrived
at by dividing annual primary production by the depth
of the mixed layer of the water column. More
recently, Rudnick and Oviatt [115] reported that
temperature and oxygen uptake by the sediment
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24
Chapter 1: Benthic-Pelagic Coupling
MAY
AUGUST
T3
CO
X
O
TJ
O
V)
1000
IL
"- $ 500
II
II °
-500
400
200
El E
to
0
-200
_l 1 1 1 L-
no data *=
ZSifC-.
no data
100 200 300 100
Distance from Bay Mouth, km
200
300
Figure 3 (above and next page). Seasonal and spatial patterns of oxygen and inorganic nutrient fluxes in the mainstem
Chesapeake Bay. The mean and standard deviations of data [10, 12, 14, Boynton et al., unpublished data] for the months of
May (left series of panels) and August (right series) are shown at stations located by the distance from the mouth of the Bay.
communities in the MERL mesocosms were highly
correlated (r=0.927) and could be described over a
temperature range of 0-25° C by the simple exponen-
tial equation O2flux = e0099u2-118. Similar relation-
ships between temperature and benthic nutrient fluxes
have also been described recently for the mesohaline
region of the mainstem Bay [44]. But in general,
Boynton et al. [12] were unable (with some excep-
tions discussed below) to demonstrate significant
single-variable correlation between sediment-water
fluxes determined at 10 stations in the Maryland
portion of the Bay (sampled four times per year) and
most environmental parameters, including bottom
water temperature.
Sediment characteristics. The conceptual model
of nutrient cycling presented in the INTRODUCTION
(Figure 1) shows that the amounts of paniculate
carbon, nitrogen, and phosphorus accumulating in
estuarine sediments represent the net effects of
organic matter deposition and nutrient remineraliza-
tion. Although it seems likely that some relationship
exists between organic deposition rates, benthic
nutrient fluxes, and the resulting carbon, nitrogen, and
phosphorus content of the sediment, most studies in
estuaries have reported no clear relationships between
sediment characteristics and sediment fluxes. This
lack of clear relationships was also reported by
Boynton et al. [12] with one important exception: a
significant positive correlation exists between the total
nitrogen content and the flux of total dissolved
inorganic nitrogen from the sediment. They attributed
this relationship to the effects of oxygen concentration
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Garber
25
1000
500
MAY
AUGUST
O)
eg
O)
X -C
3 N
C E
1?
CO O)
-500
50
25
0
-25
-50
no data
no data
600
400
200
no data
no data
100 200 300 100
Distance from Bay Mouth, km
200
300
on the products of microbial nitrogen transformations
in the sediment column or at the sediment-water
interface. For reasons that are not fully understood
but are probably linked to factors that regulate
microbially mediated-mediated nitrification and
denitrification reactions in Bay sediments [60, 120,
121, 134], nitrogen remineralization tends to be less
complete under anoxic conditions; thus nitrogen
accumulates in the regions of the Bay where the
sediment tends to be more reduced.
Dissolved oxygen. The influence of oxygen
concentration and water temperature on sediment-
water exchanges of oxygen and nutrients has been
examined in laboratory-maintained cores [25] and
empirically from field data [e.g., 12, 26]. The studies
carried out by Cerco [25,26] have shown that the
fluxes of ammonium, nitrate, and phosphate increase
with decreasing oxygen concentration. The rate of
sediment oxygen uptake also increases with tempera-
ture, but decreases as oxygen concentrations drop in
the overlying water.
Data collected by Boynton et al. [12 and unpub-
lished data] also show that rates of sediment oxygen
demand drop more or less linearly (O2 flux = 0.11
[O2] + 0.44, r=0.54) with the oxygen concentration of
the overlying water. Their field data exhibit consider-
able scatter, particularly at higher oxygen concentra-
tions, which indicates that additional factors are
involved in regulating rates of sediment oxygen
consumption.
In the passage quoted in the INTRODUCTION, Mor-
timer [79] predicted that the release of inorganic
phosphorus from sediments would increase when the
oxygen concentration of the overlying water dropped
below 2 mg/1. The studies of Boynton et al. [12 and
unpublished data] indicate that this prediction holds in
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26
Chapter 1: Benthic-Pelagic Coupling
the Maryland portion of the bay. Their data showed
that at oxygen concentrations above 2 mg/1 the
phosphate flux from the sediment was low and erratic.
When oxygen concentrations in the overlying water
dropped below 2 rng/1, the flux of phosphate from the
sediment increased dramatically. Webb and D'Elia
[137] had described a nearly identical relationship
between phosphate and oxygen concentrations during
an investigation of neap-tide stratification in the York
River estuary. They observed significant increases in
dissolved phosphate concentrations when the oxygen
content of the water dropped below 2 mg/1 and
attributed both the phosphate source and oxygen sink
to benthic fluxes of these constituents. The relation-
ship between sediment-water exchange of phosphate,
sediment redox conditions, and the formation of
phosphate-containing amorphous ferric oxyhydrox-
ides was also examined in the Potomac River estuary
by Callender and Hammond [22].
pH. Seitzinger [119] recently demonstrated that
the release of phosphate from sediments from the
upper Potomac estuary increased when the pH of the
overlying water rose above 9. Estuarine waters with
salinities above about 5 ppt are fairly well buffered
against increases in pH to these levels. The pH-
mediated release of sediment phosphate would
therefore be most likely to occur in the tidal-fresh and
low-salinity regions, such as the upper Potomac, that
are subject to intense algal blooms.
Rates of organic matter deposition. The concept of
benthic-pelagic coupling assumes a functional
relationship between the deposition of organic matter
at the sediment-water interface and benthic nutrient
remineralization. There is good reason to believe this
model is basically correct; the relationship between
annual rates of organic matter supply and benthic
carbon consumption described by Nixon [85] is very
compelling. Nevertheless, direct evidence for the
model requires simultaneous (or contemporaneous)
short-term determinations of organic deposition rates
and dissolved nutrient fluxes at the sediment water
interface. This has rarely been attempted in estuaries
(see Boynton's work on the Patuxent [16]), and only
recently have such data become available for the
mainstem Chesapeake Bay. The preliminary analysis
of the data from sediment trap deployments and
benthic flux determinations in the mid-mainstem Bay
[12] indicates that there may indeed be a linear
relationship between the amounts of paniculate
carbon sinking to the Bay bottom and the rates of
oxygen fluxes at the sediment water interface. As
predicted by the model of benthic-pelagic coupling
introduced earlier, the rate of oxygen consumption in
the sediment generally increased with higher rates of
organic carbon deposition. Boynton et al. [12] are
careful to point out that these relationships may not
hold as more data become available. Nevertheless,
they provide at least preliminary support for the view
that sediment-water exchanges respond relatively
rapidly (i.e., within a few months) to changing organic
loading rates.
The Stoichiometry of Nutrient Fluxes
Boynton and Kemp [10] have described seasonal
changes in Stoichiometry of nutrient fluxes along the
length of the mainstem Bay. In summer, the ratio of
the benthic fluxes of inorganic nitrogen and oxygen
was 18.5:1, similar to the Redfield ratio of 16:1. A
more complete stoichiometric model for the forma-
tion, deposition, and remineralization of organic
matter in the central Bay [12] included nutrient ratios
of suspended paniculate matter, material collected in
sediment traps, surficial sediments, and benthic
fluxes. The results of this analysis revealed that
particles suspended in the water column generally
have relatively high C:P and N:P ratios. The similar-
ity between the C:N:P ratios of suspended seston and
surficial sediment was striking and suggested that the
sediments suspended in the water column may contain
a considerable fraction of resuspended sediment.
Differences between nutrient ratios of suspended
seston and the paniculate matter collected in the top
and middle sediment traps were also striking. Unlike
the material suspended in the water column, material
collected in the traps followed Redfield's ratios
almost exactly in early spring and late summer, but
also exhibited significant phosphorus depletion
relative to nitrogen and carbon just before the major
depositional events in mid-spring and mid-summer.
This difference suggested that suspended and sinking
particles comprised two rather distinct populations of
particles, the former more similar to resuspended
sediment, the latter apparently reflecting the nutri-
tional state of the plankton, which undergo significant
seasonal changes. This analysis also suggested that
when sediment fluxes from all stations and all
sampling times were combined (mainstem plus
tributaries), the fluxes of inorganic nitrogen and
inorganic phosphorus across the sediment-water
interface averaged 15.7:1, almost exactly Redfield's
ratio for marine plankton. I have recently recalculated
this ratio using three years of flux data from only the
four mainstem stations and found the N:P ratio (5.7:1)
to be significantly less than the previous estimate.
This calculation for the mainstem stations is consis-
tent with reports (summarized by Nixon [85]) of
relatively low inorganic N:P ratios of estuarine
benthic fluxes. However, the benthic fluxes from the
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Garber
27
Bay's tributaries appear to be characterized by N:P
ratios that are close to Redfleld's ratio or higher.
Benthic-Pelagic Coupling in the Chesapeake Bay
Data available at present allow an examination of
the interactions of primary production, organic
deposition, and nutrient remineralization on a seasonal
(rather than annual) basis for regions of the mainstem
Bay and Patuxent River. For example, Kemp and
Boynton [68] reported daily rates of benthic respira-
tion (sediment oxygen uptake) for the shallow-water
communities near Calvert Cliffs ranging from 0.24 to
3.4 g O2 nr2. Converted to carbon equivalents, these
rates of benthic community respiration represented
30-40% of gross phytoplankton productivity in
summer, and 50-100% of plankton productivity in
winter. The annual carbon budgets assembled for this
community show that essentially all the carbon
available for deposition on the bottom could be
consumed in the benthos.
Boynton and Kemp [10] also determined spring
and summer benthic fluxes of oxygen and nutrients at
five locations along the axis of the mainstem Bay and
three locations in tributary systems. They estimated
that ammonium flux from the sediment could usually
meet 20-40% of the nitrogen requirement for phyto-
plankton production along the mainstem, and perhaps
10-25% of the plankton nitrogen requirement in
tributary estuaries. The more extensive data from the
Patuxent River estuary reported by Boynton et al. [16]
and Lantrip et al. [71] indicated that the fluxes of
ammonium and phosphate from the sediments of this
tributary are often more than sufficient (sometimes by
factors of two or three) to supply the daily production
requirments of the phytoplankton. Lantrip et al. [71]
showed that the flux of ammonium and phosphate
from the sediments represents a significant source of
inorganic nutrients in the estuary at all times of the
year except winter. During spring, summer, and fall,
the sediment release of ammonium was two to three
times greater than the inorganic nitrogen loading of
the estuary from runoff and other sources.
Returning to the mainstem Bay, Boynton et al.
[12] estimated the consumption of oxygen by mid-
Bay sediment communities could remove 2-16% of
the oxygen inventory of the water column per day.
Their estimates suggested that during periods of water
column stratification, sediment oxygen demand alone
could drive deep waters into anoxia in 10-13 days.
They also calculated that the amounts of carbon,
nitrogen, and phosphorus sinking out of the upper
waters of the mid-mainstem Bay represented a
significant fraction (3-63%) of plankton production.
The data also indicated that when oxygen levels in the
overlying water are > 1 mg/1, oxygen-consuming
processes in the sediments can account for the
consumption of 70-100% of the organic carbon
sinking to the bottom. Sediment oxygen demand
drops to zero when the overlying water becomes
anoxic, and the fraction of organic deposition con-
sumed by aerobic respiration in the sediment de-
creases during anoxic events. It is important to note
that the decomposition of organic matter can (and
undoubtedly does) continue in the sediment during
anoxic events via various metabolic pathways,
especially sulfate reduction. These reactions result in
the buildup of reduced products, such as sulfides, that
can contribute to sediment oxygen demand when the
deep waters of the Bay are reoxygenated. The rates of
these anaerobic processes in the Bay are poorly
known and need to be better characterized.
Finally, the seasonal budgets for the mid-Bay
region asssembled by Boynton et al. [12] indicated
once again that nutrients remineralized in Bay
sediments represent a significant fraction of the
amounts of inorganic nitrogen and phosphorus needed
to support primary production in the water column.
For nitrogen this fraction appears to fall in the range
of 15-40%. The fraction of phytoplankton phosphorus
demand met by the benthic flux of dissolved inorganic
phosphorus is more difficult to determine from these
seasonal measurements because the release of
phosphate from the sediments is highly sensitive to
the oxygen content of the overlying water. As
Mortimer [79] predicted, the fraction of phytoplank-
ton phosphorus demand met by benthic fluxes in oxic
regions of the Bay is probably measurable but small
(the data of Boynton et al. suggest between 0 and
40%). Phosphate flux from the sediments increases
during hypoxic and anoxic events in the mainstem
Bay and tributaries. During these events the daily rate
of phosphate release from the sediments may exceed
phytoplankton demand by as much as 200%.
SUMMARY AND RECOMMENDATIONS
The purpose of this review has been to examine
the interactions between the waters and sediment of
the Chesapeake Bay relevant to problems associated
with changes in nutrient loading of the estuary.
Studies to date have unequivocally demonstrated that
organic matter sinks through the water column, that
oxygen is consumed by Bay sediments, and that
inorganic nutrients are returned to the overlying water.
Two problems, however, complicate the task of
assembling a comprehensive quantitative model of
these processes in the Chesapeake Bay. First, no
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28
Chapter 1: Benthic-Pelagic Coupling
reasonably tightly-constrained estimates of annual
rates of sediment oxygen demand or benthic nutrient
fluxes are available for any region of the mainstem
Chesapeake Bay. This lack of annually-integrated
flux data makes it difficult to assess the contribution
of sediment fluxes to the annual Bay-wide budgets of
nutrient loadings and primary production. Second,
considerable amounts of pertinent data remain in
essentially raw form in grey-literature reports or have
been collected so recently that data reduction and
synthesis has not yet been possible.
The other findings of this review are: (1) A
survey of the white and grey scientific literature
uncovered only about two dozen reports (about half of
which are not available as peer-reviewed reports in
primary literature) that deal directly or indirectly with
the processes of sediment deposition or the exchanges
of oxygen and nutrients across the sediment-water
interface in the Chesapeake Bay. These reports
represent a small fraction of literature generated on
these subjects over the past two decades.
(2) Geographical distribution of the studies of
sediment deposition and benthic fluxes in the Chesap-
eake Bay is very uneven. While extensive data sets
on these processes are available (and continue to
grow) for the Patuxent River and the Maryland
portion of the mainstem Bay, virtually no data could
be found for the Rappahannock, York, lower James,
and eastern shore tributaries (except the Choptank
River). Very little information on these processes is
available for most of the Virginia portion of the
mainstem Bay. Studies of the mainstem Bay have
generally been conducted in relatively deep water
along the central axis, and almost no information is
available for vast shallow portions of the mainstem
Bay on the eastern and western flanks of the central
channel.
(3) Few of the studies of sediment-water ex-
changes undertaken in the Bay region have provided
data suitable for tightly-constrained estimates of
annual rates of sediment deposition or sediment
nutrient fluxes. Seasonal sampling strategies have
revealed a great deal about benthic-pelagic noise in
the Chesapeake Bay and very little about the signal.
The problem of periodic anoxia in regions of the Bay
exacerbates the problems associated with seasonal
sampling by introducing yet another important
variable in the suite of conditions that regulate
benthic-pelagic exchanges.
Recommendations for Future Work
(1) Field research aimed at determining annual
rates of sediment deposition, sediment oxygen
demand, and benthic nutrient fluxes, based on weekly
or bi-weekly sampling schedules, is needed to
adequately quantify these processes and assess their
importance in the budgets of nutrient loadings and
productivity in the Bay.
(2) Estimates of sediment deposition, sediment
oxygen demand, and benthic nutrient fluxes are
needed for more tributary estuaries on both the eastern
and western shores of the Bay. Similar work needs to
be extended into the Virginia portion of the mainstem
Bay and in shoal regions throughout the mainstem
Bay.
(3) Research needs to be undertaken to re-
examine whether, and how, various regions of the
estuary act to trap sediments and nutrients.
(4) Data from past and current monitoring
programs and modeling efforts need to be synthesized
and published in the peer-reviewed literature.
(5) Although this review was not aimed at
summarizing work on specific benthic processes that
generate the fluxes of constituents across the sedi-
ment-water interface, it is clear that further research
on reaction pathways such as nitrification, denitrifica-
tion, and sulfate reduction will lead to a better
quantitative understanding of oxygen and nutrient
dynamics in the Bay.
The Need For Experimental Work In Mesocosms
Investigations of sediment formation, deposition,
and benthic nutrient remineralization in the Chesa-
peake Bay have to date been limited to field-oriented,
essentially descriptive studies. These have provided,
and will continue to provide, important data on the
magnitude of these processes in various parts of
the Bay system. Progress in understanding these
processes will be slow, however, if this essentially
phenomenological approach is not complemented
with experimental work. This is because controlled
experiments cannot be carried out in the field.
However, an ecological experiment of mind-boggling
scale is already under way, as attempts are made to
regulate the amounts of inorganic nutrients that enter
the Chesapeake system of estuaries. Our ability to
predict the results of this experiment is extremely
limited.
I end this review by recommending that monitor-
ing and modeling efforts currently aimed at under-
standing nutrient dynamics in the Chesapeake Bay be
augmented with experimental studies in sediment-
water mesocosms. The power of mesocosm-based
research in addressing pollution-related problems has
been amply demonstrated by experiments in the
plankton-based systems at Benedict [36, 117], and the
MERL facilities in Rhode Island [63, 65, 95, 98, 115].
I close on this note because, as this review came
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29
together, it became apparent to me that the Chesap-
eake Bay yields its secrets very reluctantly. A
combination of monitoring, modeling, and experimen-
tal work is needed to make rapid strides in under-
standing its dynamics.
ACKNOWLEDGEMENTS
As always, a reviewer's first obligation is to
thank the scientists who actually did the work. I am
indebted to all those in the Bay region who have tried
to make this complex system understandable. I
especially want to thank Charles App, Mark Brinson,
Carl Cerco, Stuart Freudberg, Jack Kelly, Michael
Kemp, Edward Kuenzler, Scott Nixon, Candace
Oviatt, Sybil Seitzinger, Kevin Sellner, David Shultz,
and Bob Summers for responding generously to my
pleas for reports and data. Librarians Janice Meadows
and Kathy Heil helped with the literature search and
with location of obscure references. I thank Walter
Boynton for the opportunity to work on Bay problems
and for the unlimited access to his reprint collection
during his sabbatical leave. I thank Fran Younger for
drafting the figures. Deborah Grossman and Jeri
Phafis helped with word processing and printing the
text. I thank Michael Kemp and nine anonymous
reviewers for their comments on the first draft of this
review. I thank Elizabeth Krome, technical editor of
this series, for her patience with my revisions. This
review was commissioned by the Literature Synthesis
Committee of the Chesapeake Research Consortium.
I thank the Committee members for their support.
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Factors Driving Changes in the Pelagic Trophic Structure of Estuaries,
with Implications for the Chesapeake Bay
Peter G. Verity
Skidaway Institute of Oceanography
Post Office Box 13687
Savannah, Georgia 31416
INTRODUCTION
Physiological tolerances and food supply deter-
mine the potential success of an individual organism.
If growth is limited by the quantity or quality of food,
then population size, and therefore community
composition, is determined by factors regulating the
dynamics of prey organisms in the next "lower"
trophic level: "bottom-up" or source control. If,
however, adequate nutrition is available such that
growth is resource-independent, then population size
may be governed by predators comprising the next
"higher" trophic level: "top-down" or sink control.
Estuaries are characterized by environmental
stresses of greater amplitude and frequency than most
other aquatic habitats. Although environmental
variability occurs over broad temporal scales, the most
significant component is scaled to body size: the
smaller the organism, the faster the growth rate, the
shorter the relevant time scale. Thus, the importance
of source vs. sink control of trophic structure changes
as a function of processes determining the availability
of food and susceptibility to predation. Imbalances
between these processes cause population oscillations,
which are damped as balance is restored. However,
long-term environmental changes superimposed on
this short-term variability, whether climatic (e.g.,
temperature) or anthropogenic (e.g., eutrophication),
can significantly alter the structure of pelagic food
webs. This review considers factors regulating source
and sink control of pelagic food webs, and the role of
long-term environmental modifications in altering
these relationships. Potentially significant ramifica-
tions for trophic structure in the Chesapeake Bay are
highlighted in Figure 1.
SOURCE CONTROL
Early research in biological oceanography [88-
90] was fueled primarily by incentives to be derived
from understanding, predicting, and improving the
production of fish in the sea. Such attempts to predict
commercial yield from data on primary or secondary
production have met with little success. One reason is
that the quantity of production may not be nearly as
important as the number of steps in the food chain
[179]. Trophic dynamic theory proposes that major
energy losses, perhaps as high as 90% [200], accom-
pany each transfer between trophic levels. Such rapid
attenuation prohibits long food chains with predator
stacked upon predator. This scenario also implies that
food chains may increase in length in more productive
regions or under more eutrophic conditions [160].
Moreover, systems dominated by small primary
producers should exhibit reduced terminal production,
because of the increased length of the food chain
[179]. An alternative hypothesis states that top
predators subsisting on long food chains have lower
metabolic expenditures that balance the decrease in
available food, so that yield is constant [109].
Common to both propositions is the assumption that
changes in the size distribution or community compo-
sition at lower trophic levels have little impact on
succeeding trophic levels; that is, production is
production, and all prey are consumed with equal
efficiency and gusto.
More recent models recognize that size structure
is at least as important as biomass or production in
understanding the exchange of energy between trophic
levels [205]. Greve and Parsons' very provocative
hypothesis, which has received surprisingly little
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36
Chapter 2: Trophic Structure
Increased
Marine-Spawning
Piscivorous Fish
Reduced Fishing Elfort
Collapsed Anadromous
Piscivorous Fish
£
Figure 1. Flow chart illustrating some of the hypothetical trophic linkages which may be significant in Chesapeake Bay.
Clear arrows define pathways that appear to be strongly influenced by "bottom-up" or source control. Bold arrows show
trophic interactions in which "top-down" or sink control appears dominant. Pathways that appear to be less significant are not
illustrated for purposes of clarity. Note that most of the changes in stock sizes to the right of the anoxia compartment have
been documented, while those to the left await confirmation. SAV = submerged aquatic vegetation.
attention, speculates that dominance by phyto-
flagellates results in ctenophores and other gelatinous
organisms as top predators, compared with systems
dominated by diatoms, which support teleosts as top
predators [80]. Thus, a shift from fish to jellyfish is
predicted where eutrophication, pollutants, or other
conditions induce shifts in dominance from diatoms to
flagellates. The implications of such a shift are
underscored by the general notion that gelatinous
predators are "a dead end in the energy transfer"
[110]. Greve and Parsons' hypothesis contrasts
starkly with previous considerations of plankton
trophic interactions by implying that changes in
community composition within one trophic level, for
whatever reasons, reverberate through the food web
and result in significant alterations in the structure of
the entire food web. Evaluation of this general
concept is a focal point of this review.
Phytoplankton
Phytoplankton are the dominant primary produc-
ers in most marine environments, including estuaries,
and are the principal food supply for both the bacteria-
based microbial food web and the more traditional
zooplanktonic food web. Thus changes at this level
may have significant effects on pelagic trophic
structure. Growth rate (i.e., primary production) and
community composition may be altered independently
or concurrently. There is little evidence that long-
term changes in the magnitude of production have a
direct effect on trophic structure; the indirect effects,
however, may be considerable. For example, ample
support from mesocosm studies [91, 149] and from
the Chesapeake Bay [15, 222] indicates that increased
primary production, resulting in elevated organic
loading of the water column and benthos, stimulates
microbial utilization of most or all of the available
oxygen. The development of hypoxia and anoxia
restricts habitat space of phytoplankton and zooplank-
ton [195, 224] reducing access to nutrients and food,
and perhaps increasing exposure to predation. Anoxia
also inhibits the germination of resting spores of
phytoplankton [44] and resting eggs of zooplankton
[105, 231, 232], potentially altering the composition
of both groups. Anoxia reduces the habitat space of
fish and shellfish [39, 165, 186] and has been hy-
pothesized to induce mortality of larvae, juveniles,
and adults unable to locate oxygenated waters. In
addition to anoxia, chronic increased primary produc-
tion is responsible for decreased water transparency,
which has been associated with major declines in
submerged aquatic vegetation in the Chesapeake Bay
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37
[13, 147]. These macrophytes provide important
nursery areas for juvenile fish [86, 165], and their
disappearance may have significant implications for
predator-prey relationships among pelagic organisms
[69, 146].
In contrast to changes in the rate of primary
production, long-term alterations in the relative
dominance of phytoplankton species may directly
affect the structure of higher trophic levels and the
efficiency of transfer within food webs. The literature
is replete with data describing the nutritional value of
diatoms and their preferred status in the diets of
herbivores; however, two taxonomic groups of
phytoplankton, dinoflagellates and cyanobacteria, are
highly variable in their susceptibility to predation and
their contribution to diets of grazers. Species of
dinoflagellates differ considerably in their quality as
food for ciliates [207, 240], crustacean zooplankton
[93], and fish larvae [185]. Some dinoflagellates are
acceptable at low concentrations but are less attractive
at bloom densities [210]. It has been suggested that
certain species' lethal or inhibitory effects on grazers
are partially responsible for development and mainte-
nance of their significant blooms [61, 93].
Dinoflagellates are apparently a poor food for oyster
larvae [45], a characteristic attributed to interference
with larval feeding by dinoflagellate trichocysts [227].
At least one dinoflagellate passes undigested from the
guts of adult oysters [70]. Dinoflagellate blooms have
also been implicated in mortality of larval fish [163],
perhaps because of exudates that clog their gills [100].
The effect of cyanobacteria on pelagic food webs
appears to be due more to small size or lack of
nutritional quality than to direct toxicity. Cyano-
bacteria are ingested by both heterotrophic nanoplank-
ton and ciliates [25, 98, 157], but their nutritional
value remains in doubt. They do not satisfy the
dietary requirements of numerous planktonic ciliates
[172, 240, 246], although one species was maintained
in culture for an extended period on a pure diet of
cyanobacteria [102]. Crustacean zooplankton clearly
do not utilize cyanobacteria, which pass undigested
through their guts [102, 198]. Cyanobacteria are also
inadequate food for oyster larvae and significantly
decrease their feeding on other, more nutritious food
[46,227].
Thus changes in both the magnitude of primary
production and the dominant algae can significantly
affect pelagic trophic structure. From the perspective
of source control, the major factors driving changes in
phytoplankton growth and community composition
include temperature, the quantity and quality of
subsurface irradiance, the supply rates and ratios of
nutrients, and the influence of physical processes on
their distributions in time and space. Temperature
influences the growth rates and seasonal occurrence
patterns of phytoplankton, although primary produc-
tion and species succession do not appear to be
regulated by temperature [57, 104]. The role of
temperature in seasonal patterns is beyond the scope
of this review, but the effects of long-term changes in
temperature on biological variability are considered in
a later section. In addition to the physical and
chemical factors outlined above, a considerable body
of evidence indicates that phytoplankton growth and
community composition are affected by the availa-
bility, toxicity, and interactions between various trace
metals and other chemical compounds that originate
from anthropogenic sources [17, 62, 136]. Acute and
sublethal effects have also been documented in higher
trophic levels [170, 219]. The influence of toxic
compounds on biological communities of estuaries is
reviewed in detail in Chapter 5 of this publication
[182].
The intensity and spectral distribution of in situ
irradiance affects the growth rates of phytoplankton.
Irradiance is attenuated in estuarine waters primarily
by phytoplankton pigments and mineral suspensates
[159]. Light transmittance in the Chesapeake Bay has
decreased historically as eutrophication has proceeded
[243], to the point that primary production in portions
of the Bay and the adjacent Delaware estuary appears
to be regulated by light limitation [83, 155, 156].
Reductions in the proportion of light in the photosyn-
thetically important regions of the spectrum, which
have been documented in the Bay [32, 243], also
contribute to light regulation of phytoplankton
growth. In laboratory studies, such changes in the
light regime can affect nutrient requirements and
relationships among phytoplankton species by altering
their optimal ratios of cellular constituents [248].
Whether the temporal trends in light availability in the
Chesapeake Bay are responsible for driving changes
in the composition of phytoplankton communities has
not been demonstrated. Theoretical considerations
imply that reductions in irradiance might favor
enhanced growth of cyanobacteria and equally small
eucaryotic phytoplankton [64, 74].
Increased attenuation of irradiance may also favor
motile taxa that can readily adjust their vertical
position in the water column. The migratory ability of
bloom-forming dinoflagellates, which permits them to
concentrate in upper sunlit layers during the day and
to utilize the higher nutrient concentrations found in
deeper layers during the night, was proposed to
account for their numerical superiority over diatoms
and non-motile species in the Southern California
Bight [58]. A similar scenario may explain the
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38
Chapter 2: Trophic Structure
continued occurrence of dinoflagellate blooms in the
nutrient-poor surface waters of the Chesapeake Bay
during the summer [226]. Blooms of at least one
dinoflagellate, Prorocentrum mariae-lebouriae,
apparently develop in the northern Bay from the
introduction of "seed" populations derived from the
southern Bay via subsurface transport during high
spring streamflow [225, 226]. As a result, interannual
variability in blooms of this species [223] and others
[224] may be correlated with fluctuations in
streamflow [184], locations of associated frontal
regions [188], and pycnocline-tilting events [123].
Thus seasonal, interannual, and longer-term variabil-
ity in physical transport mechanisms, described in
detail by Itsweire and Phillips [97], may change the
composition of plankton communities. The effects on
higher trophic levels may be direct [101] or magnified
through trophic interactions [247].
Extensive field evidence has demonstrated that
annual nutrient inputs from the surrounding watershed
into the Chesapeake Bay have increased and have
resulted in enhanced primary production [15,22,142].
Mesocosm studies have confirmed the positive
correlation between eutrophication and primary
production [51, 141, 149]. Similar enclosure experi-
ments have also demonstrated that nutrient enrichment
may alter phytoplankton community composition over
periods of days to weeks [77, 181], with the limiting
nutrient and the extent of limitation varying season-
ally. Changes in phytoplankton composition would be
expected as nutrient fluxes and ratios change, because
each species has different requirements [111,217].
However, evidence demonstrating significant long-
term changes in community structure in response to
eutrophication remains equivocal. The effects of
progressive eutrophication on trophic structure in
Moriches Bay and Great South Bay (New York) are
examples often cited. Extremely dense populations of
a chlorophyte and a cyanobacterium developed in
waters heavily fertilized by effluent from duck farms,
and these blooms coincided with collapse of an
extensive oyster fishery [178]. These, and other
similar occurrences of dense blooms of phytoflagel-
lates in hypereutrophic waters [122], imply that
elevated nutrient supplies alter phytoplankton commu-
nity composition toward prevalence of small, less
desirable species [201]. Dominance by flagellates and
other nanoplankton, however, is not predicated upon
an excessive supply of nutrients [59]. Rather, the
ratio of silicon to other limiting nutrients appears
critical in determining species dominance in phyto-
plankton communities [143]. Contemporary calcula-
tions based on the supply rate of silicon suggest that
diatom production may represent one half of the total
annual primary production [50]. Much of this
apparently occurs during the spring diatom bloom.
Silicate limitation of diatom growth has been pro-
posed as a contributing factor in the collapse of then-
blooms in some years [192], although this hypothesis
remains to be experimentally verified.
Summer communities in the Bay are dominated
by flagellates, dinoflagellates, and cyanobacteria
[125]. These taxa have the "least desirable" charac-
teristics according to Ryther and Officer [180],
prompting some investigators to suggest that eutrophi-
cation has resulted in changes from diatom- to
flagellate-dominated communities [142]. Analysis of
historical data indicates that the contemporary Bay
has a longer growth period, more diversity, and a
greater total abundance of phytoplankton than in the
past [125]. Although the importance of photo-
synthetic nanoplankton has received increased
recognition [128, 189, 235], significant changes in
techniques of sampling, preservation, and enumera-
tion preclude quantitative interpretation of long-term
trends in abundance of nanoplankton, including
cyanobacteria. To the extent that historical changes
towards dominance by various nanoplankton taxa
have occurred, the structure of higher trophic levels
may have been influenced as previously outlined.
Bacteria and Heterotrophic Nanoplankton
Models of freshwater trophic relationships predict
that eutrophication increases the proportion of
inedible phytoplankton, and that "inedible algae serve
as a buffer against variations in nutrient supply" [19].
Primary production, however, provides the major
substrate for bacterial growth, and therefore enhanced
primary production can be expected to stimulate
microbial production. Such a cause-effect relation-
ship was observed along a eutrophication gradient in
mesocosms [91]. Historical data are not available
from the Chesapeake Bay, but present bacterial
communities are extremely abundant and metaboli-
cally active [123, 222, H. Ducklow, pers. comm.].
Bacterial density is positively correlated with chloro-
phyll concentration in the Chesapeake Bay [222] and
elsewhere [9]. Data are not available to determine if
phytoplankton composition directly influences
bacterial metabolism, but indirect effects are poten-
tially significant, as that proportion of primary
production not consumed or assimilated by grazers is
available to support bacterial production. The
magnitude of the spring bloom is sufficient to account
for development of summer anoxia in the Chesapeake
Bay, where bacterial abundance is a good predictor of
oxygen concentration in the water column [222].
Thus, shifts in pelagic trophic structure toward
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39
dominance by inedible or less desirable phytoplankton
do not necessarily serve as a nutrient buffer, but rather
may divert the flow of energy from "higher" trophic
levels to the microbial food web.
In addition to utilization of dissolved oxygen, the
production of new biomass by bacteria has significant
implications for the structure of pelagic communities.
Observations of relatively constant bacterial biomass
despite rapid growth rates [47,162, 244], which
suggests that bacterial production is approximately
balanced by losses to microzooplankton, form the
conceptual basis for a microbial food web [5,161].
Experimental studies indicate that bacterial mortality
is mostly due to predation by heterotrophic nano-
plankton, primarily colorless flagellates [2, 48, 82].
Other components of the heterotrophic nanoplankton,
such as heterotrophic dinoflagellates, feed primarily
on eucaryotic cells. Field investigations showed that
the dynamics of nanoplankton were coupled to those
of bacteria, albeit lagging by a few days [3], suggest-
ing that the heterotrophic nanoplankton were initially
resource-limited. The abundance and production of
bacteria in the Chesapeake Bay imply that hetero-
trophic nanoplankton may not be food-limited, at least
during the late spring and summer, indicating that this
trophic link may be significant. This conclusion may
require modification to the extent that small phyto-
plankton are an important component in the diets of
heterotrophic nanoplankton [76, 152, 214]. Phototro-
phic nanoplankton are abundant in the Chesapeake
Bay; however, the previously discussed uncertain
food quality of many of these cells may limit their
contribution to the diets of heterotrophic nano-
plankton.
The role of nanoplankton in the development of
anoxia may be especially important. As noted above,
field studies have demonstrated a strong relationship
between bacterial abundance and oxygen consump-
tion. The latter measurement includes respiration by
all plankton, most of which are assumed to be bac-
terial. Laboratory experiments have demonstrated
that the contribution of colorless flagellates to oxygen
consumption, carbon mineralization, and nutrient re-
generation may equal or exceed that of bacteria [1, 29,
75]. Assuming that an average nanoplankton,
approximated by a 4-pm sphere, respires at a mean
rate of 2.7 x lO'5 nl O2 urn'3 h'1 [29], a community of
nanoplankton present at 5x10* cells/1 [47,241] would
consume 0.1 ml Ojl"1 d"1. The closeness of this value
to measured utilization rates in the Chesapeake Bay
[142, 215] suggests that an operational microbial
food web, rather than simply bacterial metabolism,
may be responsible for the onset of anoxia. This
hypothesis, however, has not been experimentally
tested in the field.
dilates
Planktonic ciliates also prey on bacteria [175,
220], but their primary role is thought to be as grazers
of photosynthetic [73, 236] and heterotrophic [3, 196]
nanoplankton. Ciliates are a food source for suspen-
sion and filter feeders ranging from other ciliates to
adult menhaden (see summary in [239]), and this
trophic connection was hypothesized to be a signifi-
cant pathway influencing the recruitment of larval fish
[154]. In this regard, network analysis suggests that
food passing through ciliates is an important compo-
nent in the diet of striped bass in the Chesapeake Bay
[6]. There is a paucity of information on the abun-
dance and temporal distribution of both protozoan and
metazoan microzooplankton in the Bay [21], although
limited data suggest abundant ciliate populations in
spring and summer [190]. The arguments concerning
the effects of food quantity and quality on hetero-
trophic nanoplankton production in the Bay are
equally applicable to ciliated protozoans. The growth
and production rates of ciliates in Narragansett Bay
over an annual cycle were strong functions of the
quantity and size distribution of phytoplankton [237,
239]. In comparison, chlorophyll concentrations and
production of nanoplankton appear to be higher in the
Chesapeake Bay [128, 235], so that limitation of
ciliate production by food quantity may be less
significant. In Narragansett Bay, however, food
quality was a major factor regulating the structure and
abundance of ciliate populations, and it is likely to be
important elsewhere.
When abundant, ciliates can be a significant
source of nutrition for copepods [176, 221] and fish
larvae [92], especially when edible or preferred food
is scarce. They are also a critical food source during
the development of larval ctenophores, and represent
an important component in the diets of adult cteno-
phores and coelenterate medusae when stocks of
macrozooplankton are low [209,211]. This trophic
relationship is consistent with the coincident impor-
tance of small phytoplankton and gelatinous predators
in the Chesapeake Bay, and provides a documented
mechanism in support of the hypothesized dominance
of ctenophores and coelenterate medusae under such
conditions.
Crustacean Zooplankton
Copepods generally dominate crustacean zoo-
plankton communities in estuaries [21, 24], although
cladocerans may be seasonally important. Both
groups, particularly copepods, depend on phytoplank-
ton for their primary nutrition; bacteria are too small
for copepods to collect efficiently [11]. Thus changes
in phytoplankton production and community composi-
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40
Chapter 2: Trophic Structure
don represent the major source of potential food
limitation. Field studies in Narragansett Bay [55],
Puget Sound [115], and the Southern California Bight
[33] indicate that copepods are often food-limited.
Because the quantity of daily primary production in
these waters far exceeds the copepod's requirements,
the poor quality of food is generally implicated.
Specifically, ingestion of nutrient-limited phyto-
plankton [34], interference by or low nutritional value
of non-living suspended particulates [177], and
occurrence of toxic or inedible species [94] are
thought to contribute to food limitation .
In contrast, elevated nutrient levels in the
Chesapeake Bay imply that phytoplankton should be
in excellent physiological state for much of the year,
except perhaps during the collapse of the spring
diatom bloom. Contemporary levels of suspended
sediment hi the Bay do not significantly interfere with
copepod feeding [191]. Comparison of present
chlorophyll concentrations (which may be similar to
those in previous years [123, 128,192,235]) with egg
production rates ofAcartia tonsa [55] suggests that
the quantity of phytoplankton may not limit macro-
zooplankton production. These considerations imply
a nutritionally satisfactory environment for crustacean
zooplankton in Chesapeake Bay. Comparisons with
20 years ago indicate a possible increase in zooplank-
ton standing stocks in the Patuxent River [24], but it is
unknown whether these data are representative of the
entire Bay. Moreover, the different techniques of data
collection preclude definitive conclusions of historical
trends in zooplankton abundance. Indeed, copepod
production in the Chesapeake Bay may be food-
limited because of the poor nutritional value, small
size, toxicity, or low preference for the dominant
phytoplankton species. The most abundant copepods
in the bay, Acartia tonsa and Eurytemora affinis, do
not actively feed on the dominant co-occurring
dinoflagellates and avoid vertically migrating popula-
tions [193]. Blooms of inedible dinoflagellates have
been implicated in regulating copepod populations in
Swedish waters [118], and may be similarly influen-
tial in the Chesapeake Bay.
Gelatinous Zooplankton
Ctenophores and coelenterate medusae are
generally the most abundant gelatinous zooplankton in
estuaries. Characterized by voracious appetites and
rapid growth rates [166,171], these organisms are
capable of decimating prey populations. This aspect
is discussed in detail in the next section. These
organisms, however, have high metabolic rates that
require a constant food supply, and they lack storage
reserves; thus they are very susceptible to food
limitation. Elsewhere, the seasonal disappearance or
collapse of blooms of both medusae and Ctenophores
has been attributed to starvation [94,135]. In the
Chesapeake Bay, Ctenophores (Mnemiopsis) and
scyphomedusae (Aurelia, Chrysaora) co-occur and
dominate gelatinous zooplankton communities in
spring and summer [21,23, 60]. All three species
prey on crustacean zooplankton as adults, but
Chrysaora also eats Mnemiopsis and other gelati-
nous organisms [27, 36, 133], indicating that food
limitation, competition, and predation may interact to
regulate their population dynamics. In one study [60],
the mid-summer decline in Ctenophores coincided
with low levels of crustacean zooplankton and
increasing concentrations of Chrysaora, implying
regulation from above and below. Chrysaora de-
clined after the collapse of its primary food sources,
Ctenophores and copepods. This seasonal appearance
and disappearance of gelatinous zooplankton was
likely influenced by several factors in addition to food
limitation, but the availability of food may be a strong
determinant of standing stocks. In Narragansett Bay,
interannual variations in the seasonal biomass of
Ctenophores was significantly correlated with the
standing stocks of crustacean zooplankton: years with
large zooplankton populations were also years with
large ctenophore populations [49]. If this correlation
is applicable to the Chesapeake Bay, the possible
increase in copepod biomass during the past 20 years
could support larger populations of gelatinous
zooplankton. Quantitative data are not available to
evaluate this hypothesis.
Teleosts
The considerable motility of fish compared to
plankton complicates evaluation of the potential role
of resource limitation in structuring fish commu-
nities. Moreover, abundance data are generally
derived from commercial statistics, which must be
adjusted for fishing effort to resolve historical trends.
Finally, there is the uncertain relationship of parental
stock size to larval survival and recruitment. Gushing
[40] and Lasker [116], among others, give detailed
consideration to these processes. A few relevant
points are offered here on the role of environmental
and nutritional influences in the success of three
important fishes in the Chesapeake Bay: menhaden,
anchovy, and striped bass.
First-feeding menhaden larvae prey visually on
individual dinoflagellates andciliates [208]. They
select larger prey as they grow, concentrating on
crustacean zooplankton [103]. Development of gill
rakers upon metamorphosis permits switching by
juveniles to a filter-feeding mode that retains phyto-
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41
plankton cells as small as 7-8 urn [67]. Cells of 2-8
lint are also collected, but with very low efficiency
[35,67]. Larger adults remove particles ranging in
size from 16 pun to 1200 nm with increasing effi-
ciency; cells much smaller than 16 (im are not
retained [54]. Their feeding is linked to zooplankton
and larger species of phytoplankton, i.e., to particle
size rather than concentration. In contrast, the
survival of first-feeding larvae depends upon encoun-
tering sufficient numbers of dinoflagellates and
ciliates during the critical first few days. Copepod
nauplii may also be important in larval nutrition by
providing a rarer but considerably larger ration (E.
Houde, pers. comm.), as they do for growth of
ctenophore larvae [211]. Spatial variability in prey
composition and concentrations is likely important, as
adults spawn and larvae develop predominantly
outside the Chesapeake Bay. The apparent historical
increase in menhaden stocks within the Bay, sup-
ported by both catch statistics and juvenile indices
[63], suggests that good conditions exist for all life
history stages, perhaps due to the increased standing
stocks of both phytoplankton and zooplankton, and to
the ability of juveniles to utilize detritus [117]. In the
tributaries of the Chesapeake Bay and North Carolina,
the abundance of juvenile menhaden was highly
correlated with that of diatoms and microflagellates,
but poorly correlated with large dinoflagellates and
cyanobacteria [66]. These data suggest that menhaden
have specific prey preferences and use chemical cues
associated with phytoplankton to alter their distribu-
tion patterns.
Bay anchovies are zooplanktivores throughout
their lives [52] and are therefore dependent on factors
regulating the abundance of crustacean zooplankton.
Zooplankton stocks, if they have increased with
eutrophication, could support larger populations of
anchovies. Anchovies have declined in certain
portions of the Bay [63], but long-term trends have
not been documented Bay-wide. The major period of
spawning, May through August [145], coincides with
the annual maximum in copepod biomass [24, 87], but
also with the maximum abundance of gelatinous
zooplankton. Independent of possible predation on
larvae by gelatinous zooplankton, the similarity in
their food preferences suggests that interannual
variations in anchovy stock size might be inversely
related to the biomass of gelatinous zooplankton
because of competition for food. This hypothesis is
speculative and awaits experimental investigation.
Juvenile and adult striped bass are piscivorous,
while larvae feed on crustacean zooplankton. Field
surveys have implicated food density in influencing
the survival of larvae [108,194], and significant
positive correlations have been found between bio-
chemical indices of nutritional state and the abun-
dance of copepods and cladocerans [126]. These
relationships may be coincidental, but greater survival
of larvae would be expected if zooplankton densities
have increased accompanying eutrophication. Recent
laboratory experiments have suggested that striped
bass are more prey-insensitive than many fish larvae,
and contemporary plankton concentrations in Chesa-
peake Bay tributaries appear to be sufficient to sup-
port good survival of larvae (E. Houde, pers. comm.).
Adult populations have decreased in recent years,
apparently because of diminished spawning stocks or
poor survival of juveniles [78]. This declining trend
may partially reflect loss of deep-water habitat space
for adults, caused by anoxia, and loss of shallow-
water habitats for juveniles, due to declines in sub-
merged aquatic vegetation [39,165]. Thus, para-
doxically, conditions that may promote survival of
larvae (enhanced primary and secondary production)
may indirectly contribute to declines in parent stocks.
SINK CONTROL
The previous section considered the role of
resource limitation in structuring pelagic food webs.
Where limitation of growth rate is less significant,
predation may be a major process regulating popula-
tion abundance. Predators ranging in size from small
protozoan grazers to large vertebrate carnivores have
specific preferences for prey, implying that selective
predation influences the species composition of prey
communities. Thus changes in the magnitude and
distribution of predation pressure can alter the
structure of pelagic food webs. "Feeding is such a
universal and commonplace business that we are
inclined to forget its importance. The primary driving
force of all animals is the necessity of finding the right
type of food and enough of it" [56].
The nature of predation in marine ecosystems has
generally been investigated as limited trophic interac-
tions, e.g., jellyfish preying on fish larvae, copepods
preying on diatoms. Armed with such data, investiga-
tors have used theoretical and deterministic models to
examine the roles of various trophic levels in structur-
ing marine food webs [112, 203], often concluding
that "observed changes in ecosystem structure may be
caused as much by changes at higher trophic levels as
by environmental factors" [205]. The extent to which
pelagic trophic structure in marine ecosystems is
regulated by primary production, herbivory, or
predation remains uncertain [114,204].
In contrast, cascading trophic interactions induced
by predation have been proposed as a dominant force
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42
Chapter 2: Trophic Structure
regulating freshwater ecosystems [31]. Case studies
in artificial pools [96], lake enclosures [121], reser-
voirs [8], and the Great Lakes [20] have consistently
shown that addition of piscivorous fish (or removal of
planktivorous fish) results in increased zooplankton
populations, reduced phytoplankton biomass, and
increased optical transmission in the water column.
Conversely, removal of piscivores leads to an increase
of planktivores, which results in decreased zooplank-
ton biomass, enhanced phyto-plankton, and reduced
water clarity.
The superficial similarity between the latter
scenario and the present status of the Chesapeake Bay
implies that cascading trophic interactions induced by
predation may be operating in the Bay. Evaluation of
this hypothesis is complicated, however, by apparent
discrepancies and inadequate data bases. For ex-
ample, zooplankton populations do not appear to have
declined, and may actually have increased in response
to the higher phytoplankton biomass accompanying
eutrophication. This apparent trend conflicts with
cascading trophic theory; but it agrees with observa-
tions that top-down regulation of lower trophic levels
weakens as productivity increases, and is supplanted
by bottom-up control [130]. The effects of predator
control and nutrient addition on trophic structure are
independent, but may be synergistic [4, 20, 121].
Acartia tonsa, which remains the dominant copepod
in Chesapeake Bay, is particularly well-adapted to
utilizing productive estuaries [55, 87]. The enhance-
ment of Acartia dominance by elevated nutrients
[68] suggests that losses to predators may become less
significant because of increased growth under more
eutrophic conditions. The potential role of predation
in structuring other components of pelagic food webs
in estuarine systems is considered below.
Teleosts
The relative importance of source and sink
controls in driving changes in marine piscivorous and
planktivorous fish stocks is uncertain. For example,
stocks of striped bass have decreased in the Chesa-
peake Bay and along the entire Atlantic seaboard [12].
The Bay is a major spawning and nursery area, and
the decline has been attributed to reduced spawning
stocks or first-year survival [78]. The relative
contributions of environmental stress [165] and
fishing mortality [78] have not been resolved,
although previous collapses in pelagic fisheries were
attributed to environmental and climatic changes [40].
Interestingly, all major anadromous fish stocks have
declined in recent years, while recruitments of marine-
spawning stocks have increased or fluctuated with
little apparent long-term trend [63]. These data
suggest that aspects of water quality or resource
control within the Bay may be more important in
determining stock sizes than cascading trophic
interactions. The theory of top-down control predicts
that larger menhaden stocks should accompany the
collapse in striped bass. Menhaden abundance has
increased, but much of the increase has been attrib-
uted to reduced fishing mortality. Provocatively,
other predators on menhaden, such as bluefish and
weakfish, have increased in abundance concurrently
with their prey. This relationship implies that
resource control may be dominant even at the top of
the food web. Irrespective of the cause, the increase
in forage fish has significant implications for the
Chesapeake Bay. Additional evidence suggests that
predation may substantially affect pelagic trophic
structure, and these data are summarized below.
Adult menhaden preferentially collect large
phytoplankton cells and zooplankton [54] and filter
nanoplankton inefficiently. If the elevated primary
production in the Chesapeake Bay is chiefly due to
small cells, then apparent increases in menhaden
stocks [63] may reflect processes such as reduced
predation on adults or increased survival of recruits.
The increase in menhaden stocks along the entire
Eastern seaboard is at least partially due to reduced
fishing mortality. The increase might be even greater
if present fishing effort were not focused on fish aged
0-1 year (E. Houde, pers. comm.). In Narragansett
Bay, the occurrence of menhaden schools is associ-
ated with reductions in phytoplankton biomass (and
oxygen concentration) and increases in NH4 [148].
Thus, adult menhaden theoretically contribute to the
dominance of nanoplankton by removing large cells
and herbivores, while releasing nutrients that support
the continued growth of small cells. The effects of
juveniles [67] may partially ameliorate size-selective
predation pressure. The potential importance of
menhaden predation in the Chesapeake Bay was
underscored by McHugh [129], who calculated that, if
all the menhaden landed annually were present at one
time, they could filter a volume of water equivalent to
the Virginia portion of the Bay in 12 hours.
In a simulation model of Narragansett Bay [112],
menhaden exerted a controlling influence on food web
stability, particularly in damping herbivore-phyto-
plankton interactions. Increases in menhaden stocks
caused reductions in zooplankton populations, leading
to larger phytoplankton standing stocks. Menhaden
schools allowed to migrate freely around the Bay,
seeking the highest food concentrations, depleted the
zooplankton stocks by the end of each day. Mixing
from adjacent areas was almost sufficient to replenish
the zooplankton stocks. A provocative observation
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43
was that the menhaden in the model avoided areas
with abundant ctenophore populations. Although
similar models have not been constructed for the
Chesapeake Bay, these observations suggest that
interannual variations and long-term changes in
menhaden stocks may significantly alter the abun-
dance, size distribution, and spatial patterns of
phytoplankton and zooplankton communities.
Benthos
From the perspective of top-down control,
benthic suspension feeders are potentially capable of
translocating significant quantities of organic matter
from the water column to the sediment, and may be a
major factor regulating plankton populations in certain
regions [43]. Garber [72] and others [14, 109, 140]
have discussed the seminal role of benthic-pelagic
coupling in influencing water column processes. The
regulatory nature of benthic grazing has been demon-
strated in mesocosms [53] and implicated in North
and South San Francisco Bay and in Pamlico, Core,
and Bogue Sounds [37, 138, 144]. Oyster reefs [42]
and mudflats dominated by blue mussels and soft-
shell clams [28] removed significant quantities of
phytoplankton during tidal excursions.
Environmental conditions over much of the
shallow flanks and in the tributaries of the Chesapeake
Bay meet the criteria required for benthic control of
plankton populations [144]. For example, suspension-
feeding by recently-settled meroplanktonic larvae was
estimated to remove 70-100% of particles in a 7-m
water column in June during 1978-1980 [85]. In the
Potomac River, a mid-stream depression in phyto-
plankton biomass was correlated with high densities
of the Asiatic clam, Corbicula [38]. Experimental and
field abundance data indicated that Corbicula could
filter the volume of water contained in the phyto-
plankton "sag" in three to four days, suggesting that
benthic grazing was responsible for the removal of
phytoplankton.
In the Bay proper, Malone et al. [123] observed
distinct differences in primary production and
phytoplankton biomass along lateral transects in 1984.
High productivity and low standing stocks of phyto-
plankton, coupled with an apparent absence of
planktonic suspension feeders, suggested that benthic
grazing may have controlled phytoplankton biomass
along the eastern shore, where dense aggregations of
oysters occurred. Much of the mainstream of the Bay
appears to be too deep for benthic grazing to exert
substantial influence on water column structure.
Moreover, the shallow flanks apparently experience
frequent exposures to hypoxic and anoxic waters via
lateral displacements of the pycnocline [123,192],
which could reduce their grazing impact. Thus
benthic regulation of planktonic abundance and
composition is likely to exhibit considerable spatial
and temporal variability.
In the past such effects may have occurred over
larger spatial areas, as oyster populations in the early
1900's may have been capable of filtering a volume
equivalent to the entire bay in three days [106].
Current estimates suggest that the oyster population
may be <4 million bushels, compared to 80 million
bushels as recently as 1985 [183, cited in 222].
Oysters prey primarily on phytoplankton of 1-10 |im
in size [84, 245], indicating that their collapse might
have contributed substantially to the current domi-
nance of nanoplankton in the Chesapeake Bay. While
speculative, such an effect would be exacerbated by
the selective removal of large phytoplankton and zoo-
plankton by menhaden.
Gelatinous Zooplankton
Gelatinous zooplankton are notorious predators
that can rapidly crop prey populations. In estuaries,
the dominant ctenophores and medusae both ingest
fish larvae [36, 167, 233], although this is particularly
well-documented for medusae such as Aurelia [134,
234]. In Mississippi Sound, ctenophores were
reported to ingest larval menhaden and anchovies
[158]. The most frequently documented predation,
however, is on crustacean zooplankton, which has
significant implications for other trophic levels.
Predation by gelatinous zooplankton may drastically
reduce stocks of crustacean zooplankton, resulting in
diminished survival of fish larvae and reduced
recruitment of herring [133], plaice [81], haddock
[65], and salmon [153]. Other investigations have
demonstrated that predatory removal of zooplankton
by leptomedusae [94], scyphomedusae [118, 132], and
ctenophores [49] reduces grazing pressure on phyto-
plankton and results in their blooms. Thus, in other
estuarine systems [49, 113], interannual variability in
stock sizes of gelatinous zooplankton regulates the
dynamics of organisms one or more trophic levels
removed. Changes in phytoplankton and zooplankton
composition may also accompany variations in the
abundance of gelatinous predators [71, 134].
In the Chesapeake Bay, the ctenophore Mnemiop-
sis and the medusae Chrysaora and Aurelia are the
dominant gelatinous zooplankton. Calculations based
on laboratory feeding rates and field abundances
suggest that ctenophores may remove a substantial
fraction of copepod production in the Bay [10, 23]. In
other estuaries, where they have few natural predators
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44
Chapter 2: Trophic Structure
or predation pressure on them exhibits significant
interannual variability [49, 135], there are extensive
oscillations in the abundance of herbivores and
phytoplankton. In the Chesapeake Bay, Chrysaora
predation on ctenophores and other, presumably
young, medusae such as Aurelia (cited in [36])
induces seasonal fluctuations in community structure
suggestive of alternate control by food limitation and
predation [60]. In that study, increasing densities of
Mnemiopsis and Aurelia were associated with
decreasing crustacean zooplankton and higher
phytoplankton stocks, presumably reflecting predation
and release from grazing, respectively. Declines in
Mnemiopsis and Aurelia coincided with pulses of
Chrysaora, resulting in rebounds of herbivore
populations, which then reduced phytoplankton
biomass. Chrysaora collapsed after the decline of its
prey populations, suggesting starvation. Predation by
butterfish and harvestfish on Chrysaora and cteno-
phores, although not quantified in the Chesapeake
Bay, may be significant there and elsewhere [124,
150]. Predation by the atentaculate ctenophore Beroe
on Mnemiopsis may also be important in the lower
Bay [23].
There is another mechanism, in addition to
regulation of prey abundance, by which gelatinous
zooplankton may affect pelagic trophic structure.
Growth efficiencies of ctenophores are <10% [171],
and those of medusae are likely similar in view of
their low carbon content [197]. The bulk of their
assimilated food is required for energy metabolism,
and therefore much of their paniculate ration is
returned to the water column in the form of regener-
ated nutrients. Phytoplankton production is enhanced
directly by this remineralization and indirectly by
removal of herbivores. Mass mortality of medusae
apparently occurs after spawning; it may also be
related to starvation or declining temperatures in
autumn [60, 124]. This input of biomass, plus that
derived from enhanced primary production, was
implicated as the cause of repeated anoxia in Swedish
waters [118], and may possibly contribute to hypoxia
or anoxia in the Chesapeake Bay. Quantitative, long-
term data on standing stocks of gelatinous zooplank-
ton are required to address this hypothesis.
These studies suggest that predation by gela-
tinous zooplankton may have significant implications
for the structure of pelagic communities in the
Chesapeake Bay during the summer months, a
possibility supported by network analysis [6]. In
addition to cascading effects caused by changes in
abundance, predation may substantially affect recruit-
ment of oyster larvae. Strong inverse correlations
were reported between interannual variations in oyster
spat set and ctenophore abundance in Barnegat Bay
[137]. The guts of ctenophores there contained an
average of 14 larvae, which is equivalent to daily
ingestion of 168-336 larvae/ctenophore, using gut
residence times of 1-2 hr [169,212]. This calculation
assumes constant feeding, which is probably valid
since wild animals consume more food in situ than is
predicted from estimates of clearance rate and
apparent food concentration [212]. They do not
necessarily feed constantly, but apparently find
adequate food. The mean summer concentration of
larvae in the Choptank River over a two-year period
was 12 larvae/liter (range 2-55 [187]), or 12xl03 nr3
The average concentration of ctenophores during their
blooms is 10-100/m3 [113, 131]. At lower abun-
dances, ctenophores might ingest 14-28% of the
standing stock of larvae each day. Higher but still
ecologically relevant densities could rapidly eliminate
oyster larvae and thus dramatically affect oyster spat
set. These data, if representative of the Chesapeake
Bay, suggest that the coincident occurrence of large
populations of gelatinous zooplankton and reduced
oyster stocks [63] may not be accidental.
Interannual variability in the density of oyster
spat is positively correlated with salinity [230],
resulting in the prediction that oyster spat sets would
be poor in years of low salinity. Low salinity also
induces cyst formation and mortality in the polyp
stage of the sea nettle Chrysaora [26, 27], whose
abundance as adults appears negatively correlated
with streamflow (D. Cargo, pers. comm.). Since sea
nettles are hypothesized to be major predators on
ctenophores, and ctenophores may be significant
predators on oyster larvae, there is the provocative
implication that the positive relationship between
salinity and oyster set may be stimulated by predator-
prey interactions. Moreover, the decline of oysters
may contribute to the dominance of nanoplankton,
which is postulated to favor gelatinous zooplankton
through an enhanced microzooplankton food web.
These relationships remain to be documented, but they
suggest that, once initiated, biological feedback may
serve to reinforce changes in pelagic trophic structure.
Crustacean Zooplankton
Understanding of the potential role of copepods
and other crustacean zooplankton in structuring
pelagic food webs has recently undergone a concep-
tual revolution. An enlarging body of evidence
indicates that herbivorous zooplankton do not use
simple mechanical means to collect food particles, but
that they use complex mechanosensory and chemo-
sensory behavior to identify, select, capture, ingest,
and reject individual prey items [151, 164]. Such
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45
selective grazing may regulate the size structure of
plankton communities [205]; it also has significant
implications for, and is dramatically influenced by,
species composition of the phytoplankton. Acartia
tonsa is the major crustacean grazer in the meso- and
polyhaline portions of the Chesapeake Bay during the
summer [87]. It apparently prefers large particles,
gradually switching to small ones as those are
removed [173]. However, the smallest particle that it
can collect is 3-5 (im [139]. Thus the grazing activi-
ties of Acartia remove large cells and regenerate
nutrients to enhance the growth of competing nano-
plankton, promoting their dominance of phytoplank-
ton communities during the summer [128]. In
addition to modifying the size distribution of phyto-
plankton, selective grazing by Acartia and other
zooplankton may alter phytoplankton species compo-
sition. Lack of feeding on, or direct avoidance of,
certain dinoflagellates [193] and microflagellates
[218], combined with selective predation on other
edible species, may contribute to development and
maintenance of monospecific phytoplankton blooms
[95]. These accumulations of algae, if ungrazed, may
eventually fuel continued anoxia. In addition, the
documented regulation of crustacean zooplankton by
gelatinous zooplankton suggests that interannual
variability or long-term changes in the abundance of
ctenophores and jellyfish may influence the structure
of phytoplankton communities.
Protozoan Zooplankton
Ciliates and heterotrophic nanoplankton are both
capable of growth rates faster than those of their prey
[7, 30], implying that their in situ population dynam-
ics are controlled by predation. In agreement with
these theoretical considerations, laboratory and field
studies have demonstrated that predation by ciliates
regulates nanoplankton abundance [3, 168], while
grazing by nanoplankton controls bacterial numbers
[48, 119]. The applicability of these conclusions to
the Chesapeake Bay is uncertain because of the
paucity of data on these protozoans [21]. Preliminary
calculations based on radiolabel experiments and
abundance data suggest that microzooplankton
(excluding heterotrophic nanoplankton) may remove
13-55% of phytoplankton biomass during the spring
and summer [190]. Similar data are not available for
heterotrophic nanoplankton, but grazing rates (200
bacteria nanoplankton'1 hr1 at ecologically relevant
densities (2xl03 nanoplankton ml'1) suggest potential
removal of 107 bacteria ml'1 d"1- which approximates
the standing stock [123]. Clearly, predation by
ciliates and heterotrophic nanoplankton may signifi-
cantly affect the abundance of their prey, and model-
ing efforts support this hypothesis in the Chesapeake
Bay [6].
Elsewhere, grazing by microzooplankton also
influences seasonal variations in the size structure of
phytoplankton [238] and bacterial [220] communities.
Moreover, ciliates can regulate the timing and
magnitude of dinoflagellate blooms in coastal New
England embayments [242]. Although similar data do
not exist for the Chesapeake Bay, the apparent
availability of underutilized dinoflagellate populations
in the Chesapeake Bay provides a substantial food
source for microherbivores, whose grazing may
reduce the contribution of such blooms to anoxia. The
growth efficiencies of ciliates and heterotrophic
nanoplankton, however, indicate that a large propor-
tion of ingested food is remineralized [75, 236]. Since
bacteria are efficient competitors with phytoplankton
for nutrients [18], an abundant and active microbial
food web may function to sequester and recycle
nutrients.
Thus, each marine pelagic trophic level is
implicated in regulating the abundance and structure
of one or more adjacent levels. The theory of cascad-
ing trophic interactions predicts that changes at higher
trophic levels should be transferred sequentially to
lower levels. This "domino effect" has recently been
demonstrated, in freshwater enclosures, to extend
from fish to bacteria [174]. The addition of plankti-
vorous fish reduced crustacean zooplankton, resulting
in higher phytoplankton biomass. Release of dis-
solved organic carbon by algae stimulated bacterial
productivity and biomass, which supported elevated
populations of heterotrophic nanoplankton. Provoca-
tively, nutrient additions, which had no direct effect
on fish and crustacean zooplankton, further stimulated
primary production and hence bacterial production,
resulting in even higher biomass of bacteria and
heterotrophic nanoplankton. These results, if appli-
cable to more complex marine ecosystems, suggest
that eutrophication and a collapsed predator base,
whether visualized as diminished piscivore stocks or
reduced fishing effort on planktivores, may both be
driving changes in the pelagic trophic structure in
Chesapeake Bay.
LONG-TERM CHANGES AND PATTERNS
Changes are discerned by comparing differences
over time, with more data or longer periods conferring
greater significance or certainty in describing a trend.
The time period over which changes are effected is
proportional to the generation time of the pertinent
organism, with larger organisms integrating over
longer periods. Thus, changes from the bottom up
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46
Chapter 2: Trophic Structure
and from the top down will interact on different time
scales to determine pelagic trophic structure. The
potential effects of eutrophication have been consid-
ered throughout this review, because of documented
effects on food web interactions and ample data
demonstrating nullification in the Chesapeake Bay;
however, other long-term changes, principally those in
climatic patterns, significantly affect trophic structure
and likely are integrated with changes induced by
eutrophication.
Climatic changes over large spatial and temporal
scales have been recognized for some time [216]. Of
these, changes in temperature and solar radiation
appear to have the greatest impact on pelagic commu-
nities. For example, steady declines in incident
irradiance due to increased cloud cover over the
Northeast Atlantic Ocean from 1948 to 1965 were
responsible for progressive delays in the inception of
the spring phytoplankton bloom [41]. Similar, more
recent meteorological events (e.g., the "dark decade"
of the 1970's [202]) may also have influenced the
timing and magnitude of phytoplankton blooms.
Long-term data sets show a consistent pattern of
declining temperatures during the late 1800's,
followed by increasingly warmer years from ca. 1905
to 1950, and decreases again from 1950 to the late
1960's. These trends occurred in the waters of the
Chesapeake Bay [16], Narragansett Bay [99,202], and
the Gulf of Maine [213], and appear to be repre-
sentative of the eastern seaboard [216] and perhaps
the entire northern hemisphere [41].
Such climatic variability is associated with
substantial changes in the timing, magnitude, and
duration of phytoplankton blooms and zooplankton
responses [41, 202]. Perhaps more important,
significant alterations in community structure of both
groups accompany these climatic patterns. Fluctua-
tions in recruitment and year-class strength of
planktivorous and piscivorous fish, particularly
clupeids, may be related to climatically-driven
changes in their food supply [41]. Interannual
variations in the yield of commercial fisheries have
been linked to long-term fluctuations in the physical
environment [213]. Climatic changes may also
influence commercial stocks in the Chesapeake Bay
[228], although potential relationships are compli-
cated by long-term trends in fishing effort. The
mechanisms proposed to account for these biological
responses include the match or mismatch in timing of
predator and prey populations [41], and changes in
dominance between subordinate and dominant
organisms due to intraspecific and interspecific
competition [199]. Evidence suggests that both
mechanisms are likely operational. Interannual
variability in sea surface temperatures may have
altered the structure of predator communities on the
Scotian Shelf, where unusually warm years were
associated with increased overwintering survival of
ctenophores, whose depredation of zooplankton stocks
was hypothesized as the agent of extensive mortality
of haddock larvae [65]. Increased occurrence of dino-
flagellate blooms and shifts in dominance from fish to
gelatinous predators have also been attributed to
climatic fluctuations [118]. Thus long-term changes
in the physical environment may be responsible for
driving complex changes in pelagic trophic structure,
and must be considered when evaluating biological
trends and their significance.
SUMMARY AND IMPLICATIONS
Clearly, there is no simple answer to the question
implicit in the title of this article. The nutrition of one
organism requires the death of another. In the absence
of predation, a population is eventually limited by
resource availability, and the decline is seldom
gradual. Predation acts to minimize such oscillations.
In the century since Darwin identified the mechanisms
behind the origin of species, research has resolved a
number of details concerning relationships among
organisms. We still do not know, however, how food
webs are structured [127]. The available data on
marine ecosystems indicate that pelagic trophic
structure in estuaries is a synergistic product of source
and sink control. Attempts to model the structure and
function of pelagic trophic levels in the Chesapeake
Bay [6,229] and elsewhere [112,205] have identified
potentially important linkages and controlling
mechanisms. Consideration of the relationship
between the temporal scales of change and the size-
dependent generation times of organisms suggests that
the relative importance of predation and resource
limitation shifts continually in response to perturba-
tions in an unstable environment, seasonality in
physical forcing functions, and system productivity.
In the Chesapeake Bay, the crash of the spring diatom
bloom may indicate a lack of predation control at that
time. In contrast, the resulting high production but
constant biomass of bacteria implies significant
control from above. Over longer time scales, the
apparent increase in menhaden stocks suggests
diminished effects of both resource limitation and
predation.
Human activities perturb both mechanisms, and
these perturbations apparently act in concert and may
be autocatalytic (see Figure 1). From below, nullifi-
cation enhances primary production. From above,
predation, disease, and anoxia decimate oyster popu-
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47
lations, while decreased fishing effort, reduced
anadromous piscivore stocks, and increased food
availability promote menhaden stocks. These changes
may alter trophic structure by size-selective stimula-
tion of nanoplankton growth in excess of sufficient
regulation. This shift to small plankton, combined
with enhanced bacterial production, apparently
supports an active microbial food web that retains and
shunts nutrients away from the "traditional" predators.
Such an ecosystem off balance may select for
gelatinous zooplankton as important top carnivores.
Paradoxically, their growth efficiencies are so low that
they would seem to be trophic cul-de-sacs. It remains
uncertain whether their hypothesized relationship with
nanoplankton food chains is supported experimentally
and, if so, whether it reflects cause or effect. Once
the dominance is established, however, the food pref-
erences and metabolic rates of ctenophores and coe-
lenterate medusae appear to reinforce their predomi-
nance. As significant predators and inconsequential
prey, perhaps they represent a cybernetic response by
an ecosystem attempting to restore stability [120].
Much more information is needed to fully understand
their role in system energetics.
These general conclusions rely heavily on
inference and emerging new ideas rather than rigorous
scientific examination, and are best viewed as
hypotheses. Three questions are particularly salient
and appear amenable to field examination in a testing
ground such as the Chesapeake Bay: (1) Do long-
term changes in the concentrations and ratios of
nutrient supplies cause changes in phytoplankton
species composition? (2) Are cascading trophic
interactions significant in structuring estuarine
communities? (3) Does the relative importance of
source vs. sink control vary as a function of produc-
tivity? Clearly, these questions are at the core of
understanding the processes responsible for structur-
ing pelagic food webs and thus the steps necessary to
restructure and properly manage the Chesapeake Bay.
Addressing these questions requires substantially
more information on several specific trophic links. In
addition to the absence of information on gelatinous
predators, two other food web components require
further investigation.
The potential importance of bacteria and the
associated microbial food web is now well estab-
lished, but relatively little is known about its function
(i.e., as a link to higher trophic levels or as a nutrient
sink), its importance in processing cyanobacteria and
phytoflagellate production, its relationship to rneta-
zoan food webs, and its role in maintaining or
promoting the establishment of ctenophores and
medusae as significant carnivores. Does the role of
the microbial food web vary seasonally and spatially,
and are its effects amplified or minimized by changes
in nutrient supply and predator structure?
Fisheries is another component of pelagic food
webs for which there is inadequate information.
Studies in fresh water suggest that cascading trophic
interactions can structure pelagic food webs through
predation, and that adequate stocks of top piscivores
can ameliorate the effects of eutrophication. To
evaluate these relationships in estuarine environ-
ments, we will need data on the forage base and food
habits of predators, age and growth studies of prey,
predator-based mortality data, and fishery-independ-
ent stock assessments.
Several of these represent generic research needs,
but they are particularly applicable to understanding
processes regulating food webs in the Chesapeake
Bay. The biota of the Chesapeake Bay represent a
balance, however tenuous, between those that would
prosper under "pristine'"or natural environmental
conditions, and those whose tolerances encompass
effects due to anthropogenic inputs. This suite of
organisms and their interrelationships will change
with continuing human modification of the environ-
ment. This review indicates the difficulties in
ascribing cause and effect in the absence of long-term,
quantitative data. As management of the Bay pro-
ceeds, a unique opportunity exists to study and
document the interactive effects of the physical and
physiological factors that drive changes in pelagic
trophic structure. This will require a multi-faceted
approach incorporating data collection and modeling.
The Bay is very large and chemically, physically, and
biologically diverse. Previous studies have demon-
strated that results from one portion of the Bay may
not be valid in other portions. Part of the challenge
for the future is to continue documentation of spatial
and temporal variability, as exemplified by Malone et
al. [123]. Intensive field sampling and in situ experi-
mentation should be conducted on pertinent temporal
and spatial scales to determine biological responses to
environmental variation. Bay-wide monitoring
programs should be continued to provide the long
time series of data necessary to identify the cause and
significance of trends. Finally, these efforts should be
coupled to multitrophic level studies (e.g., meso-
cosms) and simulation models to test hypotheses in
detail. This approach will require a substantial
funding and management commitment; however, it
will quantify the functional relationships between
nutrient supply and fish production, and increase
our understanding of, and therefore our ability to
predict changes in, the factors that regulate pelagic
trophic structure.
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Chapter 2: Trophic Structure
ACKNOWLEDGEMENTS
Numerous friends and colleagues made
significant contributions to the evolution of concepts
incorporated in this paper. In particular, I acknowl-
edge the generous sharing of unproven ideas and
unpublished data by Steve Brandt, David Brownlee,
Hugh Ducklow, Ed Houde, Mike Kemp, Evelyn
Lessard, Tom Malone, Jenny Purcell, Mike Roman,
Kevin Sellner, Jon Tuttle, and Bob Ulanowicz. Grey
literature and various reports on the status of Chesa-
peake Bay were kindly provided by Gail MacKieman
and Karen McDonald. A literature search was run by
Janice Meadows, and Tom Turner tracked down
numerous, often obscure papers. I also gratefully
acknowledge the considerable efforts of Debbie
Craven, Dee Peterson, and Susan Salyer in word-
processing and literature citation. Comments by Ed
Houde, David Menzel, Kevin Sellner, and several
anonymous referees provided valuable refinement to
an earlier draft. The preparation of this review was
supported by a grant from the Environmental Protec-
tion Agency to the Chesapeake Research Consortium,
and by the Skidaway Institute of Oceanography.
REFERENCES
1. Andersen, O. K.; Goldman, J.C.; Caron, D. A.; 12.
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Physical Processes that Control Circulation and Mixing
in Estuarine Systems
E. C. Itsweire and O. M. Phillips'
Chesapeake Bay Institute
The Johns Hopkins University
Baltimore, Maryland 21211
INTRODUCTION
The Chesapeake Bay and its tributary rivers form
the largest and most productive estuarine system in
the continental United States. The Bay stretches 315
km from the Susquehanna River south to Norfolk,
VA. At a depth of 20 m the average width is 3.5 km,
while at 10 m the width varies from 5 to 20 km. The
Bay's general circulation has been studied over the
past 35 years by means of short- to long-term current
meter moorings, tide gauges, wind observations,
surveys of water column properties, and dye tracing
studies [15, 20, 59,61,85,86].
The physical processes driving the circulation and
stratification in a partially stratified estuary, such as
the Chesapeake Bay, can be classified according to
three time scales [10]: (1) seasonal processes (time
scale larger than a month); (2) short-term processes
(time scale of a tidal period to a month); and (3) short-
period, small-scale mixing processes (time scale less
than a tidal period). Seasonal and short-term processes
control horizontal distribution, transport, and diffusion
of water properties, while small-scale mixing proc-
esses dominate the vertical exchanges of properties
across the pycnocline (density step between two
layers of different density).
Solar heating and fresh water inflow drive the
seasonal processes that generate the general, two-layer
estuarine circulation and the mean stratification [60,
61]. In turn, this circulation controls the subsurface,
longitudinal transport of nutrients and non-motile
organisms [81, 82]. The strength of the mean stratifi-
cation created by solar heating and freshwater inflow
affects the amount of vertical mixing between the two
layers. A sharp pycnocline effectively isolates the
deeper layer from downward mixing of dissolved
oxygen and therefore contributes to the creation of
* Also Department of Earth and Planetary Sciences,
The Johns Hopkins University.
anoxic basins [50,72]. On the other hand, wind
storms in the fall and winter, when the stratification is
at a minimum, can produce top-to-bottom mixing over
large portions of the Chesapeake Bay [31,34].
The short-term processes, which include wind
forcing, tidal variations, long-period internal waves,
cross-bay seiching, fronts and plumes, ocean cou-
pling, and shelf interaction, and diurnal variation in
heat flux affect the sub-tidal circulation. Only the
overall effect of these processes on the circulation and
stratification has been studied [51, 84-86, 88]. The
spatial and temporal resolution of the past studies was
rather coarse. Typical current meter moorings had two
meters, with one meter presumably in each layer of
the water column, and the current data were averaged
over periods of 15 minutes to an hour. Higher-
resolution current profiles have been obtained only in
the tributary rivers, either with profiling current
meters (e.g., the James River [58], the Potomac River
[20], and Baltimore Harbor [15]) or with a bottom-
mounted Acoustic Doppler Current Profiler (in the
Patuxent River [66]).
Finally, short-period internal waves and turbulent
mixing control the vertical distribution and exchange
of water column properties, nutrients and organisms.
Very little is known about turbulent mixing mecha-
nisms: what are the predominant mechanisms, how
often do mixing events occur during the tidal cycle,
how long do they last, what is their intensity, and what
are the resulting vertical fluxes? Only one attempt has
been made at measuring the turbulent kinetic energy
and turbulent fluxes in the Patuxent River estuary
[70]. The identification of possible mixing mecha-
nisms in the Chesapeake Bay has resulted from field
work in other partially stratified estuaries [27,48,49,
54, 81] and the open ocean [36, 76], laboratory experi-
ments [11, 57, 76], and observations of small-scale
internal waves in the upper Chesapeake Bay [10].
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58
Chapter 3: Circulation and Mixing
SEASONAL PROCESSES: GRAVITATIONAL
CIRCULATION
Pritchard [59-61] characterized the Chesapeake
Bay and its tributary rivers as moderately or partially
stratified estuaries. At the peak of the freshwater
runoff, the major tributaries can be classified as salt-
wedge estuaries (the river flow is much larger than the
tidal oscillations). In partially mixed estuaries, a mean
two-layer flow system exists (after removal of the
tidal components) in which the freshwater upper layer
flows towards the sea, while the bottom saltier layer
travels from the ocean to the head of the estuary. This
mean estuarine circulation is commonly termed
gravitational or non-tidal circulation.
The tidal current amplitude in the mid-channel of
the Chesapeake Bay ranges from 0.5 to 2 knots (0.25
to 1 m/sec), with the strongest currents at the mouth
and the head of the Bay. The mean horizontal velocity
in each layer is about 20% of the maximum tidal
velocity [63]. In order to conserve mass, the saltier
water that flows up the Bay in the bottom layer returns
to the sea in the upper layer, thus creating a net
upward flow from the bottom layer to the upper layer.
The large ratio between the area of the Bay and its
cross-section implies that the velocity of this upward
flow is small compared with the horizontal currents
[65]. Finally, in fairly straight estuaries like the
Chesapeake Bay proper, the Coriolis force is balanced
by lateral pressure gradients resulting from lateral
variations in mass and acceleration [61]. In the
tributaries, like the James River, where curvature
effects are important, the centrifugal force contributes
significantly to the lateral balance [17,73].
The National Ocean Service of the National
Oceanographic and Atmospheric Administration
(NOAA) conducted a large-scale current meter study
of the Chesapeake Bay involving over 130 current
meters deployed for periods of a few weeks to a year.
The study covered the entire Bay in three phases,
with four long-term stations occupied year-round
throughout the entire survey. In addition to current
meter data, conductivity, temperature, and depth
(CTD) data, tidal data, and meteorological data were
recorded. Some of the data have been analyzed by
Goodrich and his co-workers [31-34], but global
scientific results from the survey report [12] have yet
to be assessed.
Mean Equations of Motion
Pritchard [61] and Cameron and Pritchard [14]
wrote out the mean equations of motion (equations
averaged over at least one tidal cycle) for a geo-
metrically simple estuary. They considered a long,
comparatively narrow estuary with a single source of
fresh water at its head. This is a good representation
of the Chesapeake Bay, except for the multiple
freshwater sources in the Lower Chesapeake Bay. The
choice of a rectilinear, right-hand coordinate system is
as follows: the x-axis is along the estuary, starting at
its head and going towards the ocean; the y-axis is in
the transverse direction, along the width of the
estuary, and the z-axis is directed vertically, with the
positive direction upward. The corresponding velocity
components are denoted as U, V, and W, respectively.
The classical Reynolds decomposition of the flow
field into a mean flow (over the averaging period) and
a fluctuating velocity can be modified to include
periodic flow oscillations such as waves and tides.
The three components of the instantaneous velocity
field can be expressed as:
[/= U + U + u
V = V + V+ v
W= W + W+ w
(1)
where the (—) represents a time average over one or
more tidal period. (Tj, V, W) denotes the tidal flow
(averaging to zero over the averaging period) and (M,
v, w) the fluctuating flow field (with time scales less
than a tidal period).
Following Cameron and Pritchard [14] and
Bowden [7], the evolution equations for the mean
longitudinal and transverse velocity components in an
elongated, partially mixed estuary are:
^+U^+W^L + -^UU+-i-UW= (2)
and
where a is the specific volume of water and/the
Coriolis parameter. The molecular terms can be
neglected compared with the turbulent stress terms
(except very near the boundaries), and the mean
transverse velocities V and V are assumed to be
small compared with their longitudinal counterparts,
so that the acceleration terms involving them can be
neglected. Finally, the Coriolis term involving the
transverse mean velocity is neglected in the
longitudinal velocity equation. Similar assumptions
apply to the vertical velocity component so that the
mean vertical equation reduces to the hydrostatic
equation involving a balance between gravity and
vertical pressure gradient.
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Itsweire and Phillips
59
In the case of the Chesapeake Bay, Cameron and
Pritchard [14] expressed the tidal velocity in terms of
a single harmonic function representing the dominant
semi-diurnal tide M2 (period 12.4 hours). For better
accuracy two other semi-diurnal tidal components (S2
and N2) and three diurnal tidal components (Kl, Ol
and PI) should be considered. The tidal amplitudes
should also be modulated to include the spring-neap
variations when the averaging period is greater than a
week. In the case of a single harmonic and a short
averaging period, the tidal flow can be written as:
U = l/0cos^u
V = V0cosw
(4)
In elongated estuaries, the longitudinal tidal
component U is the most important. The transverse
component V reflects changes in width of the estuary
and the vertical component W results from the tidal
changes in water depth. The 01 are the tidal phases for
the three velocity components. They include the time-
varying part and a constant-phase reference.
Cameron and Pritchard [14] made several
assumptions based on observations from the James
River estuary study [57] to simplify the mean equa-
tions. Pritchard [60] argued that the horizontal
Reynolds stress terms (involving a and v) were small
compared to d(uw)/dz and d(vw)/dz. Furthermore, in
the Chesapeake Bay and its tributaries, the tidal wave
is a progressive wave, so that U and W are 90° out of
phase.) For a steady state, equations 2 and 3 reduce to:
-uw
and
(5)
(6)
Pritchard [61] found that, in the James River estuary,
the main balance along the estuary was between the
pressure gradient and the vertical gradient of the
turbulent stress duw/dz but that the non-linear tidal
term UdUldx could be important. The acceleration
terms due to the non-tidal, mean motion were
negligible. In the Chesapeake Bay, the tidal term
could be significant near the mouth and the head of
the estuary where tidal amplitudes are large (about 0.6
m) and at spring tides when the tidal amplitudes are
maximal. The transverse velocity component was
found to be in geostrophic balance, with the lateral
pressure gradient nearly balanced by the Coriolis
force. The turbulent shear stress was of secondary
importance. Pritchard [61] speculated that this term
might be more significant in highly curved estuaries.
Stewart [73] used the James River data of Pritchard
[61] and Pritchard and Kent [64] to show that the
imbalance between the Coriolis force and the lateral
pressure gradient could be due to the centrifugal force
(mostly in the tidal fluctuations - V2IR) created by a
curvature of radius R [17]. The depth-varying nature
of the flow curvature suggests that a lateral shear
stress could exist and play some role in the lateral
balance [73].
The James River estuary study [61] showed that
the mean, non-tidal vertical velocity was of the order
of 10"5 m/sec, compared with a mean horizontal
velocity of order 1 m/sec. It also determined that the
level-pressure surface and the surface of net no-
motion were both near mid-depth. The water surface
and upper isopycnals (surfaces of constant density)
were tilted towards the Bay proper, and the isopycnal
surfaces near the bottom were tilted in the opposite
direction.
Mean Conservation Equation for Salt and Other
Conserved Properties
The local rate of change of salt (or other constitu-
ent) after neglecting molecular diffusion can be
expressed as:
(7)
fa™ -
where the instantaneous salinity S is decomposed into
a non-tidal mean 5 , a tidal fluctuation S and a
turbulent component s. Pritchard [59] assumed that
the tidal fluctuation for the longitudinal velocity and
salt were 90° out of phase and that the terms
involving the transverse and vertical tidal oscillations
were negligible. Then equation 7 reduces to:
dS
""
at
"1"
dus Ovs dws
dx
dz
dx
_
dy dz
(8)
In most cases, the time rate of change of salt is small,
so that the main balance in equation 8 is between the
mean salt advection (left hand-side) and turbulent
diffusion (right hand-side).
Equation 8 can be reduced further by integration
in the transverse direction [7, 59]:
OS TjOS jy#5 __ 1 dbus _ 1 Dbws
i)t Ox dz ~5 Ox 75 c'Jy
where b is the breadth of the estuary (width at a given
depth). Pritchard [59] was able to compute all the
terms in equation 9 in the James River estuary and
(9)
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60
Chapter 3: Circulation and Mixing
14.KOnl
Figure 1. Longitudinal component of the non-tidal velocity
averaged over 13 days (17-29 October 1977) showing a
two-layer circulation (from Pritchard and Rives [65]). All
velocities are in cm/sec. Top and bottom measurements
were taken at 38° 29' N lat. and 38° 23' N lat., respectively.
showed that the salt balance was between the mean
horizontal advection UdS/dx and the vertical
turbulent diffusion (l/b)d(b~ws)/dz. The mean vertical
advection WdS /dz can be important near the surface
of no-motion, where the mean horizontal velocity
vanishes.
Observations of Gravitational, Two-layer Circulation
The early current measurements for examining
the non-tidal circulation and its associated salt balance
were made in the tributary rivers rather than the Bay
proper. Pritchard [58] measured vertical profiles of
horizontal currents and estimated the vertical velocity
in the James River during the summer of 1950. These
extensive measurements of salinity and current
profiles were made across three sections during three
five-day periods. The tidally averaged velocity
profiles were consistent with a two-layer circulation
and a diffusive vertical salt flux [59, 61].
Another study [65, 84] consisted of five moorings
at 38° 29' N lat and three moorings at 38° 23' N lat,
north of the mouth of the Patuxent River. All moor-
ings, containing a total of 20 current meters, were
deployed for about three weeks starting in mid-
November 1977. Three additional moorings were
deployed between these two sections. The tidally
averaged records showed that maximum flood
currents occurred within an hour of each other in the
two cross-sections, in good agreement with the tide
tables from NOAA. Daily averages of the non-tidal
velocity showed substantial variability in current
distribution and strength. Long-term averages (over
13 days) show a classical two-layer circulation with
maximum longitudinal currents varying from -15 to
+15 cm/sec and a depth of no motion around 8 meters
(see Figure 1). The difference observed in flow rates
between the two cross-sections by Vieira [84] implies
that the vertical advection from the lower to the upper
layer contributes significantly to the balance of
equation 5. Vieira also reports large fluctuations in
the laterally integrated subtidal flow, reaching values
of 20,000 m3/sec over a day, which appear to result
from wind forcing.
Departures from the Two-layer Circulation
It is often assumed that the two-layer, gravita-
tional (averaged over several tidal cycles) circulation
observed by Pritchard [58] in the James River also
exists in the Chesapeake Bay proper and the other
tributaries. More recent current meter measurements
[20] in the Potomac River showed two-layer circula-
tion only 43% of the time. A more complex three-
layer circulation occurred during the other part of the
tidal cycle and persisted when the current measure-
ments were averaged over a tidal cycle. The measure-
ments, made with vertically profiling current meters
and a CTD taking hourly samples, covered two five-
day periods at two cross-sections.
High-resolution (1m vertically) current measure-
ments made in the upper Chesapeake Bay in the fall of
1986 with a ship-mounted Acoustic Doppler Current
Profiler [42] showed a three-layer circulation during
part of the tidal cycle (Figure 2). It is unclear whether
a three-layer circulation would appear in an average
over several tidal cycles since no long-term, high-
resolution current measurements exist. As discussed
in SHORT-PERIOD PROCESSES: TURBULENT MIXING, a
good knowledge of the velocity shear profile is critical
to the identification of turbulent mixing mechanisms.
Three-layer circulations have also been observed
in other tributaries where the freshwater input is small
compared with the tidal oscillations, e.g., the entrance
of Baltimore Harbor [15] and the Patuxent River [66].
In this type of circulation, waters near the surface and
the bottom move towards the head of the subestuary
(harbor) and water at intermediate depths moves away
from it, with important consequences on the transport
of pollutants. The reasons for this type of circulation
appear to involve enhanced mixing near the head of
-------
WEST
3 2
Distance (km)
EAST
.c
-*-•
Q.
Q
10
20
30
40
0
Distance (km)
Figure 2. Evidence of three-layer circulation in the Chesapeake Bay: cross-sections of along-bay currents (from Itsweire and
Osbora [42]). Measurements were taken in November 1986 with an Acoustic Doppler Current Profiler. Vertical resolution
was 1 m, horizontal resolution 100 m. Negative velocities are northward (up-estuary) and positive velocities are southward
(down-estuary). Top: cross-section at 38° 58' N lat.; end of flood tide. Bottom: cross-section at 38° 56' N lat.; end of ebb tide.
the harbor, with consequent reduction of the stratifi-
cation there [5, 52]. Surface water becomes denser
and deep water less dense than at corresponding
depths in the Bay outside. The water column in the
harbor tends to collapse and move outwards at an
intermediate depth, drawing in Bay water at the top
and bottom. These tributaries of the Chesapeake Bay
have a very small freshwater flow compared to the
tidal oscillations and cannot be classified as partially
mixed estuaries.
Turbidity Maximum
Regions of maximum turbidity have been
observed in the upper reaches of estuaries throughout
the world. Schubel [68] showed that the turbidity
maximum near the head of Chesapeake Bay is caused
by the continual resuspension of bottom sediments in
combination with the sediment trap produced by the
non-tidal, gravitational circulation. Throughout the
year, sediment is resuspended by tidal pumping
(maximum near the head and mouth of the Bay) and
wind waves (important near the head of the Bay where
the mean depth is less than 5 m). In the bottom layer,
Schubel [68] observed maximum concentrations (60-
280 mg/1 at depths of 8-9 m) of suspended sediments
near maximum ebb and flood tides and minimum
concentrations (20 mg/1) shortly after slack water. In
contrast, the upper layer had a nearly constant, lower
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62
Chapter 3: Circulation and Mixing
concentration (14 mg/1) of suspended sediments. The
sediment trap produced by the gravitational, two-layer
circulation can be explained as follows: during intense
tidal pumping, resuspended sediments are mixed into
the upper layer. Then, as they are transported down-
stream (seaward) in the upper layer, sediments settle
into the lower layer and are carried back upstream by
the net upstream flow of the lower layer. The location
of the turbidity maximum depends on the freshwater
discharge from the Susquehanna River. Observations
in the Tamar estuary [83] suggest than the turbidity
maximum reaches an extremum during spring tides
when tidal currents are the largest.
Development of Anoxic Basins
Anoxic basins have been observed for many years
in the middle portion of the Chesapeake Bay [50].
These basins develop yearly from March to Septem-
ber. From April to September 1980, all the deep
water (below 15 m) in the mid-portion of the Bay
(between Annapolis and the mouth of the Potomac
River) was completely anoxic. In 1984, meteorologi-
cal conditions combined to produce catastrophic
anoxia in the upper and middle Bay [72]. Although
anoxia is usually limited to the deep central channel of
the Bay, a shallow pycnocline (5 to 8 m) caused it to
extend to the shallow reaches of the eastern shores.
Officer et al. [50] attribute the development of the
anoxia to a combination of benthic respiration and
high summer stratification. The rate of change of
dissolved oxygen concentration would be mainly a
balance between the reoxygenation rate of the lower
layer through advective and diffusive turbulent
processes (source of oxygen) and the benthic respira-
tion rate (sink of oxygen) due to the decay of organic
material. Bacterial metabolism of organisms in the
water column could also be a substantial sink of
dissolved oxygen [79].
As stratification increases in the spring with high
freshwater runoff, vertical mixing across the pycno-
cline is reduced, minimizing the reoxygenation rate of
the lower layer. Concurrently, the benthic respiration
rate increases in the spring due to the oxidation of
organic detritus. Plankton blooms, which have a slow
decay rate, have accumulated from the previous
summer and fall. In September, large wind mixing
events break the stratification [34], through processes
not clearly understood. One (less likely) possibility is
that enhanced turbulence in the upper layer associated
with the wind event, combined with the reduced
stratification, erodes into the anoxic region and
enables the mixed layer to reach the bottom. Reoxy-
genation rates would then be larger than the benthic
respiration rates and deep-water dissolved oxygen
levels would return to normal values. More likely is
the sudden instability described by Bell and Thomp-
son [3]. The anoxic layer is essentially stagnant
throughout the summer, with the estuarine circulation
above it, the shear at the top of the anoxic layer being
stabilized by the strong stratification there. During fall
or early winter, the stratification is reduced and during
a storm, particularly one moving up or down the Bay,
the storm surge augments the shear until the system
becomes unstable. The whole anoxic basin is flushed
out and then ventilated by more normally oxygenated
water.
Officer et al. [50] point out that other parameters,
such as water exchanges between the mid-Bay and the
upper and lower portions of the Bay, internal recy-
cling of organic material on the bottom, variations in
the freshwater runoff flow, and spring bloom of
diatoms should also be included in the analysis.
Finally, it should be pointed out that the physical
processes governing gravitational circulation, density
stratification, and vertical mixing cannot account for
the long-term aggravation of the development of
anoxia in the Chesapeake Bay. Rather, this long-term
trend is correlated with increased yearly plankton
production and nutrient inputs in the Bay [25].
Officer et al. [50] postulated that the increase in
nutrients (particularly nitrogen and phosphorus)
stimulates growth of algae, which ultimately die and
decay, consuming large amounts of dissolved oxygen.
SHORT-TERM PROCESSES: SUBTIDAL
CIRCULATION
To the simple schematic pattern of gravitational
circulation, one must add the effects of topography,
atmospheric forcing, and oceanic forcing. They
produce an important subtidal circulation that can
completely obscure the gravitational, two-layer
circulation [20]. The effects of both longitudinal and
transverse winds have to be considered.
Atmospheric and Oceanic Forcing: Baroclinic and
Barotropic Modes
The two-month study (mid-July to mid-Septem-
ber 1974) of coastal sea level and surface winds of
Wang and Elliot [88] showed a coupling of Bay and
coastal sea levels consistent with driving via the
coastal Ekman flux. A one-year study (October 1974
to September 1975) of subtidal sea level variations
[85] showed the existence of a large barotropic
fluctuation that dominates the estuarine circulation on
time scales of several days. Wang [85] also found that
on time scales of two to three days the barotropic
fluctuations (i.e., those influencing the whole water
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Itsweire and Phillips
63
column) were driven by the local longitudinal wind
(seiche oscillation), whereas on scales of three to five
days they were driven by the local lateral wind. The
local winds set up the horizontal velocity shears
required to create the barotropic instability. In the
three to five day time scale the transport was also
found to be much larger than the river runoff.
Non-local forcing was important for time scales
>10 days, and coupling of Bay and coastal sea levels
was observed [85, 88]. A Wang and Elliot study [88]
suggested that, at these large time scales, a coastal
Ekman flux can cause a rapid rise of sea level. Winds
blowing parallel to the coastline create an Ekman flux
in the shelf waters, in turn causing sea level fluctua-
tions that propagate up the Bay.
Later data on currents, sea level, and wind
observations [86] in the upper Bay (38° 54' N lat.) and
lower Bay (37° 31' N lat.) implied that, in the deeper
upper mid-Bay, the response to wind forcing is mostly
baroclinic (i.e., one in which the upper layer responds
differently from the lower layer). Only the near-
surface current in the upper layer was frictionally
driven by the wind, resulting in large vertical velocity
shears, a necessary condition for baroclinic instability.
More recent studies of subtidal circulation
conducted in the upper Chesapeake Bay [35, 39] and
the middle part of the Bay [84] also show the impor-
tance of external forcing on the non-tidal circulation
in the Chesapeake Bay. Olson [51] used the volume
transport and wind stress data of Grano [44] and
Pritchard and Rives [65] to develop a two-layer model
explaining the variability in the subtidal circulation in
terms of a linear response to irregular, time-varying
meteorological forcing. Olson's work [51], extending
the model of Wang [85] and Vieira [84], showed that
coastal sea level tends to dominate the low-frequency
part of the transport spectrum, while wind stress tends
to dominate the higher frequencies of the spectrum.
Olson [51] also concluded that 90% of the transport
variance at the mouth of the Bay is due to coastal sea
level variations, in agreement with the Wang and
Elliot [88] observations. In the upper Bay, the
opposite is true: 90% of the transport fluctuations
come from longitudinal wind stress and freshwater
runoff from the tributaries as predicted by Vieira [84].
This conclusion agrees with that of Elliot et al. [21],
whose three-month study of the circulation near the
head of the Chesapeake Bay showed a baroclinic
response at all time scales. In the mid-portion of the
Bay, wind stress and sea level fluctuations contribute
more evenly to the transport fluctuations, with the
frequency marking the transition from sea level to
wind stress dominance between 0.15 and 0.30 cycle/
day (three- to six-day period).
Response to Tropical Storms and Cyclones
The response of the Chesapeake Bay to severe
tropical storms and extratropical storms (cyclones)
can be a combination of three phenomena: response to
wind setup, sea level variations, or intense flooding of
the tributaries.
In June 1972, tropical storm Agnes brought a
very heavy rainfall, resulting in record flooding of the
major tributaries [69]. In contrast there was little
wind associated with Agnes when it reached the
Chesapeake Bay, so that the intense freshwater runoff,
primarily from the Susquehanna River, remained
relatively unmixed as it moved down the surface of
the Bay. Salinity levels as low as 4 parts per thousand
were measured near the Chesapeake Bay Bridge [69],
and significant depression of the normal salinity levels
continued for six weeks.
In September 1975 the effects of another tropical
storm (Eloise) on the circulation near the head of the
Bay were observed by Elliot et al. [22]. Like Agnes,
Eloise brought no severe winds to the Bay, so that the
freshwater discharge from the Susquehanna River
controlled the circulation and stratification for more
than 10 days after the passage of the storm. Near the
head of the Bay, the flow of the Susquehanna River
determines the local stratification and the location of
the salinity front between landward low-salinity,
homogeneous water and seaward saltier, stratifed
water. Therefore, large river discharges associated
with tropical storm rainfalls reduce water salinity near
the head of the Bay and move the salinity front
seaward. Consequently, in the upper 50 km of the
Bay, the two-layer gravitational circulation is replaced
by a net seaward flow for several weeks.
Wang's observations [86] of the response of the
Chesapeake Bay to wind forcing by extratropical
storms have documented that large sea level fluctua-
tions (0.5 and 1 m) are frequently produced, superim-
posed on the astronomical tides. The transports
associated with these storm surges are, at maximum,
much larger than the river runoff. The local seiching
effect is reasonably well understood, but the coupling
between the estuary and the coastal ocean beyond has
only been partially explored.
Coastal Plume and Estuary-shelf Interaction
The dynamics of river and estuarine plumes
discharging fresher water in the adjacent shelf waters
have been studied intensively in New York Bight [9],
the Connecticut River estuary [28-30] and the
Chesapeake Bay [4, 6, 32, 33]. Observations of the
Chesapeake Bay coastal plume [6] have shown that
the fresher estuarine water takes a broad anticyclonic
turn at the mouth of the bay and then forms a narrow,
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64
Chapter 3: Circulation and Mixing
high-velocity jet propagating southward along the
coast. The plume spreading and the intensity of the
jet are directly affected by the direction of the winds.
The water exchange between the Chesapeake Bay
and the adjacent shelf'is highly dependent on wind
forcing and coastal sea levels [4, 32, 33]. Goodrich's
data [32, 33] showed that the water exchange between
the Bay and the ocean consists of a steady, seaward
surface flow on which a larger, oscillatory, depth-
independent, meteorologically driven flow is superim-
posed. Goodrich [33] also estimated that only 3% of
the outflowing water was of mean estuarine salinity.
Wind-induced Destratification
Between early September and late February the
mean stratification in the Chesapeake Bay reaches a
minimum value for two reasons: surface cooling
decreases the vertical temperature gradient (from
September to January), and freshwater runoff remains
small after reaching a minimum in August, so that the
vertical salinity gradient can be reduced by vertical
diffusion. A redviced stratification means a lower
available potential energy of the system, so that a
lower amount of kinetic energy is required in order to
break the stratification and convert more potential
energy into kinetic energy and viscous dissipation.
Goodrich et al. [34] showed that large-scale destratifi-
cation in the fall resulted from strong wind mixing
events. During the fall of 1981, complete vertical
mixing of the water column occurred after the wind
induced a strong top-to-bottom velocity shear, i.e, the
value of the Bulk Richardson <1. These results
suggest that the flow might be dynamically unstable
during those periods (Richardson number <0.25).
Spring-neap Overturn
Complete vertical mixing of small, partially
mixed estuaries can also occur after strong spring
tides. Haas [38] observed that the York and Rap-
pahannock Rivers (both partially mixed tributaries of
the lower Chesapeake Bay) oscillated during late
summer between states of vertically stratified density
profiles and homogeneous density profiles. The
vertical velocity shear induced by the high spring tides
was probably large enough for the flow to become
unstable. Haas [38] and Hayward et al. [40] observed
that complete vertical mixing of the water column was
achieved four days after the high tides. The velocity
shear produced by the following weaker tidal oscilla-
tions was insufficient to sustain vertical turbulent
mixing, and the water column completely restratified
within a week. This spring-neap overturn has been
observed only in late summer when the right combina-
tion of large-amplitude spring tides and low river
Breaking waves
Figure 3. Mixing processes in the upper ocean (from
Thorpe [75, 76]).
runoff is achieved. It is likely that this type of vertical
mixing is confined to the tributaries rivers rather than
the Bay proper. Goodrich [31] showed that spring-
neap overturn is significant in the lower Bay, but is
not the dominant mechanism for vertical mixing of the
water column.
SHORT-PERIOD PROCESSES: TURBULENT
MIXING
Much less is known about mixing processes in
partially stratified estuaries than about those in the
open ocean and laboratory flows. At best, past
estuarine studies have suggested potential mecha-
nisms that could produce vertical mixing in the
Chesapeake Bay. It is therefore necessary to briefly
review the fundamental results on small-scale mixing
in oceanic and laboratory flows before examining the
problem of turbulent mixing in the Chesapeake Bay
and other partially mixed estuaries.
Small-scale Mixing Processes in the Ocean and
Laboratory Flows
In their reviews on small-scale turbulent pro-
cesses in stably stratified flows, Turner [78], Browand
andHopfinger [11], Gregg [36], and Thorpe [76] have
presented complete summaries of the various mecha-
nisms producing vertical mixing in the open ocean
and laboratory flows. Only those processes most
likely to occur in the Chesapeake Bay and its tributar-
ies are reviewed in this section. Depictions of oceanic
mixing mechanisms compiled from Thorpe [74-75]
are presented in Figure 3.
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Itsweire and Phillips
65
Turner [78] pointed out the importance of
identifying the sources of energy for turbulent mixing
in order to classify mixing processes. Turbulence can
have its source in kinetic energy (tides, mean shear,
surface and internal waves) or potential energy
(unstable density profile leading to intrusions and
convection). Turbulence can also be thought of as
being internally generated (internal wave breaking,
Kelvin-Helmholtz and Holmboe instabilities) or
externally generated (energy input at the boundaries
due to wind stress and topography).
Turbulent kinetic and potential energies equations.
It is convenient to introduce the equations for the evo-
lution of the fluctuating kinetic and potential energies
to illustrate the differences between the various insta-
bility mechanisms. Depending on the turbulence gen-
erating mechanism considered, the balance in these
equations will be between different terms. The instan-
taneous velocity and scalar (temperature, salinity etc.)
can be decomposed into a mean field (over an averag-
ing period of 5 to 15 minutes) and a fluctuating field
(with no mean value over the averaging period). This
Reynolds decomposition is similar to the decomposi-
tion used in MEAN EQUATIONS OF MOTION and MEAN
CONSERVATION EQUATION FOR SALT AND OTHER CON-
SERVED PROPERTIES with two exceptions: the tidal oscil-
lations are not specified (instead they are assumed
constant over the averaging period); and the averaging
period is much shorter than a tidal period. The coordi-
nates system is identical to the one used in MEAN
CONSERVATION EQUATION FOR SALT AND OTHER CON-
SERVED PROPERTIES (the mean flow is along the x-axis).
Following Phillips [56] and Turner [77], who
assumed that the fluctuating field is horizontally
homogeneous, one can approximate the turbulent
kinetic energy equation as:
dz
wp
(10)
dz
where 2 is the total fluctuating kinetic
energy, p is the fluctuating pressure and e the
dissipation of q2-. Eq. (10) is a good approximation
for the upper ocean. The physical meaning of the
various terms in equation 10 is as follows.
Temporal evolution of the turbulent
kinetic energy, usually considered to
be small.
Horizontal advection of the turbu-
lent kinetic energy by the mean flow.
It could be important in tidal estuaries.
Y-wq*
uw
dU
dz
Redistribution of the turbulent
kinetic energy by the pressure field.
Initially, the energy is put into the
longitudinal component u, but is
redistributed among all three compo-
nents by the pressure. If the redistri-
bution is slow, the turbulence will be
anisotropic (e.g., in stably stratified
flows, the vertical component w2 is
smaller than the horizontal compo-
nents due to the loss to buoyancy
effects). The relative importance of
this term compared with other terms
in the right-hand side is not known,
as pressure-velocity correlations have
never been measured in water.
Redistribution of the turbulent
kinetic energy by the fluctuating field.
It includes turbulent diffusion (spread-
ing of turbulent energy by breaking
internal waves).
Production of turbulent kinetic
energy from the mean shear (usually a
source term). Small-scale eddies in
the presence of a mean shear extract
energy from the large scales of the
flow. This term is the primary source
of energy in internal wave breaking,
Kelvin-Helmholtz and Holmboe
instabilities, forced convection, and
mixing along a solid boundary.
Buoyancy flux, could be a source
(free convection, restratification) or
sink of kinetic energy (any shear-
induced mechanism). Turbulent mix-
ing in a stably stratified fluid converts
kinetic energy into potential energy,
raising the center of gravity of the
system (positive buoyancy flux). This
term is usually small compared with
the shear production and the dissipa-
tion rate. On the other hand, an unsta-
bly stratified flow (produced by day-
time evaporation and nighttime sur-
face cooling) will convert potential
energy into kinetic energy (negative
buoyancy flux). Buoyancy flux can
then be the dominant source of energy.
Dissipation rate of turbulent kinetic
energy into heat due to viscous effects.
-------
66
Chapter 3: Circulation and Mixing
This term is the main sink of turbulent
kinetic energy.
Similarly, the equation for the evolution of scalar
(C=C+c) variance in a mean vertical gradient can be
expressed as:
DC2
(11)
CJX
dwc2
dz
The terms in equation 11 can be interpreted as
follows:
Time rate of change of scalar
variance.
Horizontal advection of scalar
variance by the mean flow.
Redistribution of scalar variance by
the turbulent field (includes turbulent
diffusion).
Divergence of scalar variance trans-
port through molecular diffusion
(usually neglected)
Rates of production of scalar
variance from the mean scalar
gradient. This term is the dominant
source of scalar variance due to
vertical turbulent mixing. It is mostly
balanced by the rate of destruction of
scalar variance.
Rate of destruction of scalar
variance, always positive.
Xc
Fundamental parameters in turbulent stratified
shear flows. Several parameters defined from the
mean velocity and scalar field are useful for character-
izing the state of stratified flows and predicting
whether turbulent mixing is likely. They are defined
as follows:
The Brunt-Vaisala or buoyancy fre-
quency, which is the highest frequen-
cy at which internal waves can exist.
Ri = M2/S2
Richardson number, which is a
measure of the dynamical stability of
the flow. A sufficient condition for
stability is Ri > 0.25. In hydraulics,
the Froude number Fr=Ri "2 is usu-
Ra = -2^
ally used. These two numbers are the
most important parameters in strati-
fied shear flows. When there is no
mean shear (or no measurements of it)
a Bulk or overall Richardson number
can be defined as Ri = (gApdl p)u2
where « and d are representative
turbulent velocity and length scales
and Aplp the fractional density change
over d.
The Flux Richardson number, which
is a measure of the mixing efficiency.
It represents the fraction of kinetic
energy (extracted from the mean
shear) going into mixing the fluid
(buoyancy flux).
The Rayleigh number. Ap/p is the
destabilizing density fraction over the
vertical distance d and K is the thermal
diffusivity. This number is important
when convection is the dominant
turbulent mechanism. Temperature
and salinity Rayleigh numbers can be
defined specially when convection is
driven by surface cooling and
evaporation, respectively.
The turbulent Reynolds number
where u is a turbulent velocity scale
and d a turbulent length as defined
previously. It is a measure of the
strength of the turbulence.
The density ratio, which is a
measure of the contribution of
temperature and salinity to density
across an interface. This ratio is the
most important parameter for doubly-
diffusive processes (not likely to occur
in the Bay). In the Chesapeake Bay,
salt contributes mostly to the density
gradient, yielding a density ratio
|R |«1. The temperature contribution
to the density gradient can be either
positive (up to 20% in the summer) or
negative (-10% in the fall).
Vertical mixing in the upper ocean boundary layer.
In the upper ocean most of the energy-producing
vertical mixing is introduced externally through the
air-sea interface. Surface stress, surface waves, and
internal waves are thought to be responsible for
Re =
ud
n _
a AT
/HAS
-------
Itsweire and Phillips
67
creating most of the turbulence. Laboratory and
oceanic observations lead to the following conclu-
sions: The temperature and salinity (and hence
density) profiles in the surface layer are well mixed.
The mean velocity is nearly constant with depth,
implying a rapid flux of momentum through the layer.
The depth of the mixed layer is limited by buoyancy
and the turbulent fluxes at the bottom of the mixed
layer depend on the deepening rate of the layer. The
influence of surface waves on the small-scale proc-
esses in the mixed layer is not well understood. The
best-known phenomenon created by the interaction of
surface waves and the wind-driven current is Lang-
muir circulation [74] shown in Figure 3. Langmuir
circulation could be a very important mechanism for
transfering gases to the upper ocean. Finally, convec-
tion due to nighttime cooling can induce motion
through the depth of the mixed layer and contribute to
the entrainment across the thermocline. Recent
measurements of turbulent dissipation rates at the
equator [46] show a strong diurnal cycle due to
surface cooling.
Mixing in the ocean interior and in stratified shear
flows. Turbulence in the thermocline [36,74] results
from internal wave breaking and interactions, shear
instabilities (small, short-lived turbulent puffs), and
near inertial shear (creating persistent mixing patches
extending several meters vertically). In all cases,
turbulent mixing is created and maintained by extract-
ing energy from the local mean shear. The strength
and duration of mixing events appear to be highly
correlated with the mean shear distribution [41].
Figure 3 shows the role played by internal waves
in generating turbulent mixing. Internal waves can
break or interact with each other or solid boundaries
to form critical layers. The most important role
played by internal waves is probably more indirect:
propagating waves can increase the shear at interfaces,
thereby reducing the Richardson number and leading
to instabilities such as Kelvin-Helmholtz and Holm-
boe instabilities. Lee waves generated over topogra-
phy can break and produce rotors (see Figure 3).
Bottom and boundary mixing. As mentioned in
the previous section, internal wave interaction with a
sloping boundary or reflection off the bottom can
create mixing in the ocean interior. The most impor-
tant mechanism for generating turbulence near the
bottom is flow (generated by large-scale and gravity
currents) around topography. These currents create
mixed layers that are advected into the ocean interior.
Bottom mixing itself is localized in the bottom
boundary layer, whose thickness is limited by
buoyancy effects. In the deep ocean, it is not clear
whether the boundary layer is more likely to be well-
mixed or stratified. In shallow water, mixing driven
by the bottom produces turbulence that can extend to
the surface. This weak stratification is governed by
the balance between the kinetic energy put in by the
bottom-generated turbulence and the potential energy
input due to solar heating.
Small-scale Mixing Processes in Partially Mixed
Estuaries
In his review of estuarine circulations, Pritchard
[63] speculated that the dominant mixing agent of
moderately stratified estuaries is turbulence caused by
tidal action (bottom, surface, and boundary mixing)
rather than the velocity shear at the interface between
the two layers (dynamical instabilities). Turbulent
eddies mix the salty water upward and the fresh water
downward. In addition to the salt flux, there is a
momentum flux with turbulent shear stresses extend-
ing through the water column. These conclusions
come from indirect inferences, since turbulent mixing
processes have not been directly measured. The
enhanced mixing (compared with highly stratified
estuaries) produces a horizontal axial pressure
gradient in both layers.
Some early turbulence measurements made in the
Patuxent River [70, 89] with a three-component
Doppler shift meter showed that the turbulence was
very anisotropic (the vertical turbulent velocity was
much smaller than the horizontal turbulent velocities).
That work was never pursued in the Bay proper with
modern instrumentation.
Time dependence of mixing events in tidal estuaries.
Nearly all studies of mixing processes in estuaries
(partially mixed, salt-wedge or homogeneous estuar-
ies) have implied that turbulence is very patchy in
time and space. Mixing events across the pycnocline
(density interface) seem to occur for short periods
(several minutes) during part of the tidal cycle (ebb
tide).
Partch and Smith [54] noted that turbulent mixing
in the Duwamish River (a shallow, salt-wedge
estuary) was highly time-dependent. They observed
that the kinetic energy and the vertical salt flux are
maximum when the tidal currents are the largest
(maximum ebb). Their most exciting finding was the
high intermittency of turbulent mixing. Nearly half of
the vertical salt flux occurred in 16% of the tidal
cycle. Partch and Smith [54] also identified the
dominant mixing mechanism in the Duwamish River
as an internal hydraulic jump generated by the tidal
flow over bottom topography. In that case, entrain-
ment driven by the turbulent bottom boundary layer
and Kelvin-Helmholtz shear instabilities were less
important.
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68
Chapter 3: Circulation and Mixing
Bottom mixing and topographical effects. In well-
mixed and weakly stratified estuaries, most of the
turbulence is produced near the boundaries, and
especially the bottom. Large velocity shears and
turbulent shear stresses are created in the bottom
boundary layer by tidal action. Three-dimensional
variations in the topography, curvature effects, and the
time-dependent nature of the tidal forcing will affect
the magnitude and spatial and temporal distribution of
near-bottom velocities and turbulent stresses, and
hence vertical mixing. Stratification due to suspended
sediments can also affect the turbulent stress near the
bottom [37].
In partially mixed estuaries the situation is even
more complex because the gravitational circulation is
superimposed on the tidal oscillations, and because
the bottom boundary layer and the pycnocline may
interact to add buoyancy effects. Erosion of the
pycnocline by entrainment from the bottom boundary
layer can be an important factor in the ventilation of
the anoxic bottom water during the summer [43].
This erosion of the pycnocline observed in Lake Erie
[43] could be significant in the shallow reaches of the
Bay and needs to be investigated.
Shear instability and internal wave breaking.
Brandt et al. [10], using thermistor chains and an
acoustic echo sounder, observed both trains of mono-
chromatic internal waves and random internal wave
fields in upper Chesapeake Bay. Some of these
internal waves eventually break, producing turbulence
around the pycnocline. The generating mechanism for
the internal waves is thought to be similar to the
internal hydraulics associated with tidal flows in
Knight Inlet [24]. In Knight Inlet, during ebb tide, the
flow separates behind the sill, setting up a mode 1 lee
wave above the sill. Then, at slack water, the lee
wave moves over the sill, leaving a train of internal
waves propagating up-inlet at a speed of 0.5 m/sec
[23]. Gargettetal. [27] observed that the small-scale
structures shown on the acoustic images of the echo
sounder were indicative of turbulence. Since there are
important differences in depth, width, and stratifica-
tion between Knight Inlet and the Bay Bridge area of
Chesapeake Bay, smaller amplitude waves, and hence
less turbulent mixing, are expected in the Chesapeake
Bay. The thermistor chain observations of Brandt et
al. [10] are encouraging enough to justify pursuit of
their study with actual turbulence measurements from
a towed body, which would be the best suited instru-
ment for shallow water turbulence measurements.
New et al. [49] observed intense mixing periods
in the Tees River estuary. The intense mixing periods
were thought to be produced at high water by a series
of bridges, with mixing subsequently advected down-
stream by internal waves traveling downstream during
ebb tide. As a result the surface layer increased in
thickness and salinity, and the center of mass of the
water column was raised. New et al. [49] also ob-
served some internal waves breaking into billows over
topography. In the Chesapeake Bay, similar intense
mixing events could be generated by the pilings of the
Bay Bridge and could play a significant role in the
exchange of nutrients and plankton between the upper
shallow Bay and the deeper mid-Bay.
Boundary mixing and lateral dispersion. Tyler's
current meter study [79] was done in conjunction with
dye tracing of a 22-meter subsurface chlorophyll
maximum of a red-tide dinoflagellate near the Bay
Bridge. Three strings of current meters were moored
at 38° 58' N laL south of the Bay Bridge, and three
strings were moored north of the bridge at 39° 2' N
lat Each string had three or four meters (depending
on the water depth), and current data were recorded
from 1 May to 12 May, 1980. Again the 12-day
average of the non-tidal velocity showed the two-layer
circulation pattern predicted by Pritchard [59, 61] with
maximum velocities of 10-16 cm/sec. In addition to
the longitudinal non-tidal currents, the isotach
contours of the cross-channel currents at 38° 58' N lat.
averaged over the 12-day period showed a mean
convergence at the interface of the two layers.
Phillips et al. [57] showed in the laboratory that a
similar flow pattern results from boundary mixing
along a sloping wall and produces a third layer in the
middle. Their experiment suggested that, in a tidal
estuary like the Chesapeake Bay, a mean lateral
circulation develops with an intrusion from the
boundary layer into the pycnocline. This lateral
convergence in the pycnocline is compensated for by
a divergence in the homogeneous layers below and
above the pycnocline. The Phillips et al. conclusions
are supported by Tyler's observations [80] of weak
lateral dispersion and some vertical spreading of a
Rhodamine dye patch injected in the lower part of the
pycnocline below the Bay Bridge.
Estuarine and plume fronts. Small-scale fronts
occurring along the shallow side of the Bay and near
the mouth of the tributaries on time scales of a tidal
period can be important factors in vertical and lateral
mixing of the water column and the development of
an unsteady lateral circulation. Bowman and Iverson
[9] described the physical mechanism leading to
frontogenesis for both estuarine and river plume
fronts.
Estuarine fronts can extend several kilometers
along the side of the estuary [9,44] and have a life-
time on the order of a tidal period. In the Chesapeake
Bay, estuarine fronts are most likely to occur along
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Itsweire and Phillips
69
the eastern side of the Bay where the pycnocline is
tilted upward and the river runoff is smaller than on
the western side. There, tidal bottom mixing can
homogenize the shoaling water when the tidal currents
are large, causing the pycnocline to break the surface
and creating a front [2]. Beardsley and Boicourt [2]
attributed the occurrence of lateral fronts in the lower
Chesapeake Bay to a combination of factors: an
increase of the gravitational circulation (in geostro-
phic balance), the addition of fresh water from the
western side and the widening of the bay south of the
Patuxent River. Aerial photography [44] of several
estuarine fronts inside Delaware Bay showed strong
convergence velocities (-10-20 cm/sec) associated
with the fronts. This unsteady lateral circulation is
very efficient at accumulating organic and toxic
material. Seliger et al. [71] observed increased
phytoplankton production in interfrontal regions of the
Chester River. Garvine [29] suggested that both
vertical and lateral mixing could be important
components of lateral fronts.
Plume fronts often form where the freshwater
runoff from the tributaries discharges into the estuary
[9]. Plume fronts have also been observed inside the
tributaries, at flood tide [13]. These plume fronts are
similar to the larger estuarine plume discharging on
the continental shelf [28,30] described in COASTAL
PLUME AND ESTUARY-SHELF INTERACTION. Byrne et al.
[13] showed that a plume front is a persistent feature
of the flow at Hampton Roads in the lower James
River, which plays an important role in retaining
oyster larvae within the James River estuary. Al-
though the driving mechanism for a plume front
(buoyancy spreading) is different from that of the
lateral estuarine front (tidally generated bottom
mixing), the two types of front appear to have similar
properties (cross-front convergence and intense
vertical and lateral mixing).
Lateral seiche. Dyer [18] and New and Dyer [48]
showed that lateral first- and second-order surface
seiches interacting with the longitudinal currents and
topography could generate lateral internal waves
several times during the tidal cycles. Their observa-
tions were made in Southampton Water, a partially
mixed estuary, 10 km long, 2.5 km wide, and less than
13 m deep. Dyer [18] proposed that these lateral
internal waves would move the pycnocline closer to
the high-shear bottom layer when resonant conditions
were satisfied. Near the bottom, the velocity is
stronger and the local Richardson number (a measure
of the flow dynamical stability) could fall below the
critical value necessary to produce turbulence.
Although lateral seiching has been frequently ob-
served in the Chesapeake Bay, it is not clear whether
it could generate large enough velocity shears to
produce turbulent mixing.
SUMMARY
Our present knowledge of the physical processes
that control circulation and mixing in the Chesapeake
Bay is very incomplete. The external energy sources
driving the circulation (tides, wind forcing, solar
heating, freshwater discharge, and oceanic coupling)
have been identified, but their effects on the circula-
tion are only qualitatively known. In contrast,
fundamental questions remain about turbulent mixing
processes in the Bay: what are the generating mecha-
nisms, where and when do mixing events occur, and
how do they affect biological productivity.
Gravitational, tidal and sub-tidal circulations
combine to set the local currents, stratification, and
vertical shear. The vertical shear is important in
creating and sustaining mixing events. It gives a
qualitative measure of the kinetic energy available for
vertical mixing. In turn, vertical mixing is an impor-
tant process in partially mixed estuaries. It determines
the vertical distribution of salinity, temperature,
dissolved oxygen and nutrients, but it has not been
localized or quantified in time and space. Studies by
Partch and Smith [54] and Brandt et al. [10] have
underlined the highly time-dependent nature of
mixing events in partially mixed and salt-wedge
estuaries, although much more remains to be done to
uncover the long-term effects of the various transient
events.
The mean estuarine circulation consists of a two-
or three-layer flow, but the vertical and horizontal
resolutions of past current meter moorings have been
insufficient for a definite conclusion. High-resolution
(1 m vertically and 100 m horizontally) current
profiling in the mid-Chesapeake Bay with an Acoustic
Doppler Current Profiler revealed a three-layer
circulation during part of the tidal cycle. This study
emphasizes how poorly the shape of the current
profile is known in the Bay and how it varies during a
tidal cycle. We need to accurately determine the
shape of the current profile (two, three, or more
layers) and the relationship between the shear and the
density profiles to quantify the dynamical stability of
the flow. Once the dynamical stability of the flow is
known, it will be possible to estimate when and where
mixing will occur. As can be seen in the data of
Goodrich et al. [34], the existence of a strong vertical
shear is a necessary condition for complete vertical
mixing.
Past studies were based on the premise that the
two-layer non-tidal circulation observed in the smaller
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70
Chapter 3: Circulation and Mixing
tributaries also applied to the Bay proper. The scale
difference between the Bay and its tributaries contrib-
utes to making the circulation of the tributaries
simpler than the Bay's. Since the tributaries are
narrow, very little transverse variation in the flow is
possible except near river bends where curvature
effects are important, and wind forcing tends to be
simpler. Also, the effects of the earth's rotation are
small, as the internal Rossby radius of deformation (a
measure of the width required for propagating long
period internal waves in a channel) is larger, at 100
km, than the width of most tributaries (4 to 7 km,
depending on the stratification). In the Bay proper,
Kelvin waves can freely propagate and a longitudinal
Ekman transport can be set up. Although few direct
observations have been made [13], large-scale eddies
generated by the fresh water inflow, topography and
coastline variations are likely to exist These eddies
could play an important role in the horizontal turbu-
lent mixing and dispersion of water properties and
momentum.
PROSPECTS FOR FUTURE RESEARCH IN THE
CHESAPEAKE BAY
Future research on the oceanography of the
Chesapeake Bay should emphasize identifying and
assessing the relative importance of all the physical
processes driving the circulation and turbulent mixing.
Such knowledge is a minimum requirement for the
development of three-dimensional numerical models
of the circulation of the Chesapeake Bay with a
predictive ability. Two complementary sets of
observations using new technologies developed for
open ocean research should be conducted.
Long-term Monitoring
First, surface features such as eddies, fronts, and
plumes should be observed synoptically with remote-
sensing (satellite and aircraft) and land-based (CO-
DAR) instrumentation. Satellite instrumentation can
now measure surface wind, ocean surface topography,
sea surface temperature, and chlorophyll concentra-
tion on 1-km scales. These observations should be
complemented with higher-resolution (10 m scales)
and more accurate aircraft measurements, as well as
surface current measurements using a CODAR system
such as the one deployed in Delaware Bay [1, 55]. A
similar system could be deployed in the Chesapeake
Bay and combined with the remote sensing observa-
tions of the surface temperature and phytoplankton
concentration to define the general circulation of the
Chesapeake Bay and validate the numerical models.
Ultimately, the surface observations should be
complemented with high-resolution in-situ vertical
measurements from ships and moorings.
Second, long-term stations with high vertical
resolution comprising upward-looking acoustic
Doppler profilers (currents), CTD, dissolved oxygen
and microstructure (kinetic and potential energy
dissipation) mooring should be used to quantify the
seasonal and meteorologically induced changes in the
vertical distribution of properties, examine the
development and erosion of the pycnocline, and
estimate the intermittency of mixing events. These
stations should be located in the Bay proper along the
deep channel, and should be complemented with
meteorological land-based stations. The long-term
observations should also be used to calculate budgets
for mass, heat, salt, momentum, and energy.
Studies of Specific Processes: A Multldisciplinary
Approach
First, the development of anoxia and its conse-
quences on the environment should be studied.
Detailed field studies aimed at capturing the forma-
tion, break-up, and ventilation of anoxic regions
should clarify the processes involved and the precise
conditions necessary for break-up. Measurements
must be frequent and of high resolution, particularly
in the vertical dimension, if they are to catch the
transient but important mixing events. These meas-
urements should be combined with measurements of
bacterial and benthic oxygen consumption measure-
ments in order to model the oxygen budget.
Second, relationships between physical processes
and larval recruitment at the mouth of the Bay should
be clarified. This type of study should be aimed at
determining the variations in recruitment of larvae
(e.g., blue crab, croaker, and menhaden) from the
ocean into the Bay due to natural physical processes
(gravitational, tidal, and meteorologically induced
flows) at the mouth of the Bay. Acoustic sensors with
multiple frequencies may be useful for measuring
larval concentration. These measurements should be
conducted during or shortly after wind events
(Goodrich, pers. comm., 1987).
Third, the influence of wind-induced mixing on
primary productivity should be studied. Wind-
induced mixing events produce temporary break-
downs of the pycnocline resulting in discrete bursts of
nutrient input to the surface. This intermittent change
in environmental conditions affects plankton produc-
tivity, and therefore affects the whole ecosystem.
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Itsweire and Phillips
71
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Atlantic bight In: Warren, B.A.; Wunsch, C.,
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3. Bell, R.C.; Thompson, R. Valley ventilation by cross
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4. Boicourt, W.C. Circulation in the Chesapeake Bay
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-------
Genetics and the Conservation of Estuarine Species
Laura Adamkewicz
Department of Biology
George Mason University
Fairfax, Virginia 22030
Robert W. Chapman
The Chesapeake Bay Institute
The Johns Hopkins University
Shady Side, Maryland 20764
Dennis A. Powers
Department of Biology
The Johns Hopkins University
Baltimore, Maryland 21218
INTRODUCTION
The physical and chemical structure of an estuary
presents organisms with a mosaic of habitats to which
they must adapt. The two most obvious features of
this environment are the saline and thermal gradients
produced by the interaction of fresh water from rivers,
land runoff, and rain with water of oceanic salinity.
As discussed elsewhere in this volume, the interaction
of these two water masses, coupled with physical
features such as topography, wind forcing, and tides,
produces a complex and variable environment This
environmental mosaic presents many challenges to
species that depend upon estuaries during all or part of
their life cycles. A species may respond with a single,
broadly adapted form, or it may subdivide into many
local populations, each adapted to a narrow range of
environmental conditions. Understanding which of
these adaptive responses has occurred in a particular
species is critical for both conservation and
management.
Evidence of adaptive response by a species can be
derived from studies of morphology, physiology,
demography, ecology, or genetics. In general, such
studies compare the characteristics of two or more
groups of individuals separated in time or space and
seek to infer similarity or difference. These studies
take on added significance when the characteristics
used have a known pattern of origin and transmission,
as is the case with genes. Differences of morphology
or physiology can be misleading because they may
represent transient responses to the environment by
individuals that are genetically alike. Conversely,
because not all genes have morphological effects,
individuals that appear to be identical may be
genetically unlike. Studies of genetic composition
will look past surface similarities and differences to
the permanent characteristics of an individual. Any
differences in genetic characteristics must represent
differences in parental background, and such
differences arise only from mutation followed by
differentiation, either through drift or through natural
selection.
Except for a small number of domesticated
species where pedigrees are well known, data on
genetic relationships come from analyses of nuclear
genes or their protein products. The total genetic
variation obtained from these data can then be divided
into variation within populations and variation
between populations [35]. At the extremes, a species
may form a single unit which contains much variation,
uniformly distributed, or it may form many quite
different units, each one uniform within itself. This
range of possibilities is referred to as genetic
structure. Where along the continuum of possibilities
a species falls depends in part on the patchiness of
the environment and in part on the biological charac-
teristics of the species—its dispersal capabilities, its
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76
Chapter 4: Genetics
reproductive schedule, and its typical population size.
Because the genes with all their variants represent the
permanent heritage of a species, efforts to preserve or
protect an organism must rest on an understanding of
its genetics.
GENETICS AND CONSERVATION
Conservationists and resource managers must
make a fundamental decision on the intensity of
intervention for any species that is either threatened or
in need of management for greater productivity.
While in real instances their strategy will often be
mixed, the choice is among four levels of increasing
intervention: (1) to protect existing natural popula-
tions through control of harvesting; (2) to enhance
reproduction and survival through the creation of
suitable habitat; (3) to supplement natural recruitment
with captive-bred and released individuals; or (4) to
attempt full domestication so that natural populations
are not harvested at all. An understanding of the
genetic structure of the species is essential at this
stage. No matter what a program's goal may be, the
manager must know whether the species involved has
many highly differentiated populations or whether it is
genetically similar throughout its range. A high
degree of differentiation implies isolated gene pools
and genetic adaptation to highly localized conditions.
Interference with such a system risks destroying that
adaptation, with irreversible loss of genetic potential.
When a species has a more uniform distribution of
genetic variation, managers and conservationists can
act more freely.
If the manager's choice is protection or enhance-
ment, the effort must extend over a much wider area,
must be tailored to each site and must include detailed
monitoring of many local populations when the
species is highly differentiated [35]. When the choice
includes captive propagation, the manager must be
aware of the genetic background of animals chosen as
breeding stock and must consider what genetic impact
captive-bred animals may have on natural populations
in the areas where they are released. When a species
is highly differentiated, release of genetically incom-
patible animals may in itself cause disruption or
extinction of the native population. Even when the
species involved is genetically uniform throughout its
range, release of hatchery-bred individuals should be
done cautiously. The manager must guard against the
unintended genetic drift and loss of useful variation
that can occur when a small number of parents
produces all the animals released [1, 21, 35]. Another
problem against which the hatchery manager must
guard is inadvertent selection in favor of traits that
enhance survival in the hatchery but are deleterious in
nature. A system of genetic monitoring can detect
such changes before they become a serious problem.
On occasion, the manager may wish to produce
animals different from those in wild populations and
here captive breeding offers a special opportunity. If
the situation warrants, managers may attempt to
influence the genetic composition of a natural popula-
tion through the release of genetically improved
individuals. Such a situation may arise when a newly
introduced disease threatens unprepared natural
populations, as has happened with the disease MSX
infecting oysters in the Chesapeake. If full domestica-
tion is the choice, then careful selection of broodstock
from the wild followed by selection of genetically
improved animals will be the normal course. For this
review, we focus on genetic information useful to
managers and conservationists as they decide on
measures to protect the living resources of the Chesa-
peake Bay. In particular, we assess the evidence on
genetic subdivision and local adaptation for selected
species.
METHODS OF GENETIC ASSESSMENT
The most direct method of genetic assessment is
analysis of the breeding results of controlled matings.
In some cultured species, genetic relationships among
individuals can be calculated directly from pedigrees
that trace the parentage of individuals back to a few
wild-caught ancestors. Such data are available for
catfish, salmonids, and a few other freshwater species.
For invertebrates, similar records are kept on certain
hatchery-bred stocks. These stocks are mass-mated,
however, with both identities and numbers of parents
known only approximately, so that analysis is
complex. The pedigree method also has the substan-
tial disadvantage that the species must be breedable in
captivity. Molecular techniques developed over the
past few years avoid this problem and allow one to
infer actual genotypes to individuals for which no
breeding data are available.
The molecular method longest in use is gel (either
starch or acrylamide) electrophoresis of enzymes and
other proteins. It is based on the well-verified
principle that genes are DNA sequences which
directly encode protein sequences. If a change in
DNA sequence produces a change in a protein, it often
also changes the electric charge or the shape of that
protein. Electrophoresis separates proteins according
to their movements in an electric field, the rate of
which depends on a combination of molecular charge
and size. The investigator then employs a specific
stain to identify a particular enzyme known to be the
-------
Adamkewicz, Chapman, and Powers
77
product of a single gene. Each mobility variant is
believed (and sometimes verified) to represent a
different allelic form of the gene produced by a
mutation. The method is indirect and cannot detect all
variants, but electrophoretic surveys have the advan-
tage that many individuals can be examined for a
large number of genes relatively quickly, often
without killing the organism. Most judgments about
genetic structure, genetic relatedness, and speciation
are based on these data.
Sometimes, useful genetic information can be
developed from karyotyping, that is, the microscopic
examination of chromosome numbers, sizes, and
shapes. This technique has been very fruitful with
mammals. Fish, however, are remarkably uniform in
their chromosome numbers, and the techniques that
work so well in mammals have not proved transfer-
able to these animals. Menzel [28] has pioneered the
application of the technique to bivalves, but again the
transfer of a technique developed for mammals has
been too difficult to permit general use.
In the newest and most direct molecular method,
the DNA sequence of a gene is determined directly for
each individual. Comparison among individuals and
populations, however, requires that the gene be
recloned and resequenced separately for each individ-
ual. Because this technique is extremely laborious, an
approximation of it using restriction enzymes and
mitochondria! DNA, rather than nuclear DNA, is
growing in popularity [3]. Mitochondrial DNA
(mtDNA) is a small, circular DNA molecule that
codes for only a few genes and is found in the mito-
chondria within the cells of ah1 eukaryotic organisms.
It can be purified separately from the nuclear DNA
that encodes most of a cell's genes. After isolation,
the mtDNA is treated with restriction endonuclease
enzymes that cut the DNA only where a specific
sequence is present. The resulting fragments are
separated by gel electrophoresis (much as in the
enzyme technique) and their lengths estimated from
their mobilities. If mutation causes a new site to
appear, a smaller fragment will be found; and if a
mutation causes a restriction site to disappear, a new
and larger fragment will be seen. The method permits
one to construct maps of these sites and compare them
among individuals. Unlike nuclear genes, which
follow the rules of Mendelian inheritance, these
mtDNA variants are maternally transmitted. Only the
mitochondria present in the egg appear in the progeny,
while the mitochondria present in the sperm are lost.
This fact makes mtDNA particularly useful for the
assessment of geographic limits of lineages, of
colonization patterns, and of contribution by individ-
ual lineages to recruitment in a given year. Protein
electrophoresis is the more useful technique in
determining the genetic structure and genetic identity
of a species.
Seldom have all of these techniques been used in
the study of a single organism. Data from enzyme
electrophoresis are available for many species in
diverse phyla, while mtDNA data exist primarily for
vertebrates. Studies based on chromosomal analysis
or pedigrees are very rare. For this review, we focus
our attention on a few well-studied organisms that are
also of biological, recreational, or economic impor-
tance to the Chesapeake Bay. We have selected the
mummichog (Fundulus heteroditus), the striped bass
(Morone saxatilis), and the white perch (Morone
americana), from the fishes, because these species
represent sedentary, anadromous, and semi-anadro-
mous life histories, respectively, and because their
biology and genetics are better known than those of
most species. Some fish species not considered here
have been the subjects of interesting studies that
specialists will wish to consult: the American eel [4,
41], menhaden [19], sea trout [34], Atlantic croaker
[40], and blue fish [40]. From the bivalve mollusks,
we have chosen to review the blue mussel (Mytilus
edulis), the hard clam (Mercenaria mercenaria), and
the oyster (Crassostrea virginica), for much the same
reasons as the fish were selected. Because of its great
importance in the bay, we have included the blue crab
(Callinectes sapidus), although virtually nothing is
known of its genetics. We have excluded estuarine
plants entirely because so few data exist for them. For
extensive references to work with terrestrial plants
and animals, the reader may consult Schonewald-Cox
et al. [35].
FINDINGS FOR FINFISH
Fundulus
Studies of the mummichog (F. heteroditus)
provide and excellent example of what can be
accomplished when the techniques of molecular
genetics are combined with those of physiology and
ecology to explain the distribution, abundance, and
evolutionary history of an organism. For a
comprehensive review of this effort, readers should
consult Powers et al. [31]. Several investigators have
demonstrated that different electrophoretic variants of
a gene can affect traits important to the survival of an
individual. A series of studies on Fundulus have
shown that respiratory stress triggers hatching [16]
and that the enzyme lactate dehydrogenase, form B
(LDH-B) influences the oxygen carrying capacity of
the blood. Finally, DiMichele et al. [17] demonstrated
that embryos with different genetic forms of LDH-B
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78
Chapter 4: Genetics
differed in their hatching times. Investigation of other
genes with electrophoretic variants shows that some
of these variants also have a very significant effect on
developmental time [32]. All of these studies and
many more on a variety of organisms suggest that
electrophoretic variants of enzymes represent
heritable variation traits important to evolutionary
fitness. Thus, populations can diverge genetically in
the process of microevolution to produce locally
adapted races.
Evolutionary biologists believe that microevolu-
tion proceeds either by random change (drift) or by
directional change (selection) in the frequencies of
these variants and of other genes [18, 35]. If this
belief is correct, then examining the pattern of gene
frequencies throughout the range of a species will
show whether that organism forms locally adapted
races. Sometimes, such information will also provide
clues about the evolutionary forces that are acting and
the amount of genetic exchange between areas [18].
Powers et al. [31] summarized investigations of the
geographical distribution of variants for 16 genes in
F. heteroclitus along the eastern coast of North
America. These investigators found a clear differen-
tiation into northern and southern populations, with a
sharp cline in gene frequencies around 41° N lat. This
pattern repeats itself within the Chesapeake Bay,
where fish from the fresh headwaters of the rivers
show the "northern" suite of variants, while fish near
the mouths of the rivers and the entrance to the bay
show the "southern" set.
Such a pattern has two possible explanations. A
series of clines might each be maintained by active
selection and represent adaptation to salinity and/or
temperature gradients. Conversely, F. heteroclitus
might form just two broadly adapted subgroups. In
that case, the northern types within the bay would
represent refugees trapped after the last glaciation as
water temperatures rose and organisms that were
adapted to low temperatures retreated north. Studies
of mtDNA sequences are well suited to distinguish
between these possibilities. Gonzalez-Villasenor and
Powers [31 and unpublished] have examined mtDNA
frequencies in four samples of Fundulus and found a
clear distinction between fish on each side of the line
at 41° N lat Fish from specific localities above and
below this line could be interrelated through a series
of single DNA alterations. Powers estimated that the
two groups differed by a minimum of nine changes in
their mtDNA. The length of these series and the
extent of the differences indicate that these two groups
have been separated for at least several thousand
years. This conclusion clearly supports the refuge
interpretation of the electrophoretic data.
Because more information is available for
Fundulus than for other species of fish, the mummi-
chog can serve as a model for other, less well under-
stood, systems. Genetic and biochemical evidence of
the sort outlined above provides guidance to anyone
interested in managing or protecting this species or
species with similar life history strategies. F. hetero-
clitus does not migrate extensively either before or
during sexual maturity, yet a large proportion of the
genetic variation in this fish exists within populations
rather than between them. Recognition of just two
long-standing subgroups in this fish, and the associa-
tion of those groups with adaptations to northerly or
southerly environments, indicates that species like this
one can be managed over broad areas rather than river
by river.
Striped Bass
Unlike Fundulus heteroclitus, the striped bass
(Morone saxatilis) has a typical anadromous life
history. The fish spawn in the low-salinity tributaries
to the Chesapeake Bay and then enter the oceanic
environment to grow and mature. Early efforts to
identify distinct subpopulations of M. saxatilis
focused entirely on morphological differences (such
as fin ray numbers and body proportions) among fish
taken from various rivers [36]. More recent studies
employing electrophoresis of serum transferrin and
other proteins as well as morphometric traits have
produced conflicting results. Using these techniques,
Morgan et al. [29] distinguished four assemblages of
bass from the Choptank, Nanticoke, Elk, and the
Patuxent-Potomac rivers. However, Sidell et al. [37]
were unable to replicate these results using the same
techniques. Berggren and Leiberman [5] were able to
discriminate fish from the major regions of the
Atlantic coast—Chesapeake Bay, Hudson River, and
Albemarle Sound—but they could not replicate the
earlier genetic distinctions among Chesapeake Bay
stocks. In part, the lack of agreement in these studies
is due to the limited amount of genetic variation
shown by M. saxatilis. In a typical species of fish,
30% of the genes will be variable. In striped bass,
only 3 of 56 genes tested (5%) had more than one
variant [38]. Very few species of any organism show
such low levels of genetic variation.
Because of the conflicting findings based on
morphology and nuclear genes, Chapman and Powers
[14] initiated a study of mtDNA variation in striped
bass from the Chesapeake Bay and from Dan River,
NC. They found that M. saxatilis from these areas
consists of 14 distinct matriarchal lines whose
frequencies differ significantly both between the
Chesapeake and North Carolina sites and among four
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Adamkewicz, Chapman, and Powers
79
major regions of the Chesapeake. On the basis of
mtDNA sequence differences, these investigators
recognized four major subpopulations in the Bay —
Worton Point, Choptank, Patuxent-Potomac, and
Rappahannock. Their data also support a separation
between fish from the Patuxent and Potomac areas,
whereas previous studies could not find significant
differences. Overall, the work on striped bass shows a
species that can be subdivided but that has limited
genetic variation and is not strikingly different from
place to place.
A more important finding in Chapman and
Powers's study concerns recruitment and the source of
each year class. Table 1 summarizes the data for the
1982 cohort of striped bass sampled over a three-year
period [14]. These fish contained a new mtDNA
sequence that had not been detected among the adult
females that presumably produced the cohort. This
sequence, designated "C," was found in 25% of the
males from the 1982 cohort taken on the spawning
grounds in 1984, 53% of the 1982 females taken near
Hart-Miller Island in February of 1985, and 50% of
the 1982 males taken on the spawning grounds in
1986. Furthermore, male mtDNA sequences differed
significantly in frequencies among the spawning areas
in the 1984 sample but not in the 1986 sample. These
data present two challenges to our understanding of
population dynamics in this fish—to understand the
origin of the new sequence and to explain the marked
change in its frequency over time.
The large increase in type C frequencies between
1984 and 1985 may be caused by migration into the
Maryland portion of the Chesapeake Bay from
localities where this type is more common. Data from
distant areas are simply not available, but the life
history of M. saxatilis [36] suggests that this explana-
tion is correct Striped bass do not migrate exten-
sively from their natal rivers during the first two years
of life. After this sedentary period, fish move into the
deeper portions of the Bay, where the histories of
males and females diverge. During their third and
fourth years, females migrate offshore, and most do
not return until age five or six when they spawn for
the first time. Males reach sexual maturity at age two.
Because they do not migrate extensively at this age,
their mtDNA frequencies in a given spawning area
reflect the contribution of matriarchal lineages within
that area two years earlier. Males do migrate exten-
sively within Chesapeake Bay between the ages of
two and five years, after which they begin oceanic
migrations. Thus, mtDNA frequencies in males that
are two years old may differ among spawning grounds
due to the influence of matriarchal ancestry, but
extensive mixing after this age obliterates these
Table 1. Numbers observed for each mtDNA
genotype in adult females during 1984 and in the
1982 cohort of fish during a three-year period [14].
Fish tested
Cohort Collected Sex
mtDNA genotype
A B C D/E F
Adult
1982
1982
1982
1984
1984
1984
1986
F
M
F
M
7
11
4
2
24
43
16
21
0
18
23
34
2
0
2
4
1
0
2
8
differences among aggregations of males. If this
hypothesis is true, a survey of females from the 1982
cohort during their first spawning run in 1987 should
reveal differences among spawning aggregations
around the bay. The samples necessary to examine
this hypothesis have been collected, but the mtDNA
data are not yet available.
The life history strategy of Morone saxatilis
produces a complex demographic structure. Whether
genetic mixing is sufficient to prevent-long term
differentiation is not yet clear, but the appearance of a
new mtDNA sequence in young fish that was not
detected in spawning females suggests that an
additional factor is acting. Clearly the C sequence
must have been present in breeding females, because
it is far too frequent in the 1982 cohort to be a new
mutation. It is plausible that the C form was suffi-
ciently rare to be overlooked in the survey by
Chapman and Powers but that females of this type
contributed a disproportionately large number of
progeny to the 1982 cohort.
The hypothesis that a large percentage of a year
class are the progeny of a small proportion of avail-
able females may, in conjunction with population
declines over the past 15 years, explain the disparity
between the findings of Morgan et al. [29] and those
of later investigators [5, 37]. The hypothesis may also
explain the remarkable lack of protein variation in M.
saxatilis. Both a declining census and an unequal
contribution to recruitment will reduce the "effective"
size of the population and increase the rate of genetic
drift (i.e., random changes in gene frequencies and
loss of variant forms) to produce a classic "genetic
bottleneck." Morgan and his group collected their
data primarily in 1967, before the decline in numbers
of striped bass began, whereas Grove's and Sidell's
work coincided with rapidly falling numbers. It is
quite possible that the conclusions of each study were
correct at the time but that the genetics of the species
changed between one study and the next. Morgan
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80
Chapter 4: Genetics
(pers. comm.) has noted the complete elimination of
infrequent transform types from the Potomac River
bass over the past 15 years. If the conclusion that M.
saxatilis has experienced a severe genetic bottleneck
can be substantiated, it will have profound implica-
tions for the protection and management of this
species. Ongoing studies of mtDNA frequency
distributions should provide an answer. If this study
substantiates the hypothesis that loss of variation is a
recent event, conserving what variation does remain
must be a very important aim of all captive breeding
programs.
White Perch
Studies of the white perch (Morone americana)
are important in their own right and also because they
provide a comparison to the closely related striped
bass. Conclusions for each species are more robust in
the context of conclusions for the other. M.
americana is abundant and widely distributed along
the eastern coast of the United States. It has a semi-
anadromous life history. Like the fully anadromous
striped bass, it migrates from the low-salinity mouths
of rivers into fresh headwaters to spawn, but, unlike
M. saxatilis, it does not migrate into the high-salinity
waters of the main bay and ocean. In a tagging study
of M. americana, Manseuti [27] found that Patuxent
River animals rarely migrated to adjacent areas and
concluded that this population should be managed as a
single stock. Woolcott [1961], using meristic and
morphometric characters, found evidence that
supported Manseuti's conclusion and extended it to
include most riverine populations in the Chesapeake
Bay.
In the early larval stages, Morone americana is
extremely difficult to distinguish from Morone
saxatilis and studies of larval recruitment in these
species are difficult. This problem led Morgan [29]
and Sidell et al. [37] to develop electrophoretic
methods for distinguishing larvae of the two fish in
mixed collections. Morgan used general protein
staining of soluble muscle proteins separated on
polyacrylamide gels, while Sidell's group employed
enzyme-specific staining and starch gel electro-
phoresis. Each method has its advantages and both
provide unambiguous and relatively rapid species
identification. However, neither study was intended
to assess the overall amount of genetic variation
present in M. americana or the genetic structure of
this species.
Studies in progress by Mulligan and Chapman
[30] and by Bowen [6] on mtDNA sequences provide
some information on the genetic background of white
perch in the Chesapeake. Mulligan and Chapman
have found evidence that M. americana forms only
three distinct groups within the bay: (1) a southern
assemblage inhabiting the James and York rivers, (2)
a Potomac population, and (3) a large upper-Bay
assemblage encompassing all western shore tributaries
from the Patuxent north and all eastern shore tributar-
ies north of the Nanticoke. Bowen's work suggests
that the white perch in the Rappahannock also form a
group distinct from all others. These mtDNA data on
genetic divergence, together with the observation that
M. americana prefers water of low salinity (<10 ppt),
suggest that water of high salinity at the mouths of the
southern tributaries is an important barrier to move-
ment in this fish. This hypothesis should be con-
firmed and the genetic structure of this species
determined by a survey of electrophoretic variants of
nuclear genes.
FINDINGS FOR INVERTEBRATES
Blue Mussel
Just as Fundulus heteroditus is among the most
thoroughly studied of the fishes, the ecological and
evolutionary genetics of the blue mussel (Mytilus
edulis) are among the best studied of any mollusk.
Much of our understanding of allozyme genetics in
natural populations of bivalves comes from the work
of Koehn's laboratory with this mussel (for a compre-
hensive review see [26]). Several studies have
demonstrated that individuals with different electro-
phoretic variants of enzymes (hence of genes) have
different tolerances for environmental factors.
Evidence is also strong that survival and vigor
increase with increasing heterozygosity (possession of
two different variants for a given gene). These
findings imply that if Mytilus forms distinctive local
races, that fact will be reflected in distinctive local
frequencies for electrophoretic variants. These studies
also suggest that inbred, highly homozygous animals
will be less fit than outbred individuals with large
amounts of genetic variation.
Virtually all of the data on geographic variation
in M. edulis comes from studies of electrophoretic
variants of enzymes, and a clear pattern has emerged
from these studies [26]. Gene frequencies are
relatively uniform over large areas, with an occasional
abrupt change in frequency over a short distance. The
evidence is clear that these sharp clines are maintained
by natural selection acting on post-larval stages
through increased mortality of less favored variants
[25]. Thus Mytilus does show adaptation to environ-
mental conditions. The geographic scope of these
adaptations is a broad one, however, and one does not
see highly differentiated local races with very re-
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Adamkewicz, Chapman, and Powers
81
stricted ranges. The adaptations themselves are not
extreme either. The same set of variants is present at
both ends of a cline with only the relative frequencies
of the individual types changed. Each cline is
opposed by the mixing of pelagic larvae of all types
from both ends, and this mixing is in most cases
sufficient to prevent the formation of distinctive local
races.
The same pattern occurs in European waters
where a second species, Mytilus galloprovencalis, is
also present. The two mussels can be distinguished
by their electrophoretic variants, but neither shows a
high degree of differentiation. The first study of
mtDNA sequences in these animals confirms the
findings of earlier studies [39]. Much less is known
about other genera of bivalves, but whenever a species
has a planktonic larval form, isolated local races seem
not to form.
Hard Clams
Like the blue mussel, the hard clam (Mercenaria
mercenaria) shows very little differentiation in the
frequencies of electrophoretic variants along the
Atlantic coast of Virginia [2]. Unlike Mytilus,
Mercenaria shows no significant changes in frequen-
cies even across an extreme gradient in salinity.
Because hard clams do not tolerate low salinities, they
inhabit only the lowest portion of the Chesapeake Bay
up to the mouth of the James River at a salinity of
about 17 ppL In the James River region, a polluted
area at the limit of its salinity tolerance, these clams
maintain gene frequencies like those along the ocean
shore [7]. Data on mtDNA variation in M.
mercenaria are now being gathered but are not
expected to contradict the conclusion that these clams
do not form genetic strains adapted to very localized
conditions.
The State of Virginia supports a long-standing
program of mariculture for hard clams, which includes
efforts to produce genetically improved strains with
increased growth rates. Although stocks that have
been maintained by hatchery culture for several
generations do show some effects of genetic drift,
these effects are very minor and do not include either
the substantial increases in homozygosity to be expec-
ted from inbreeding or a marked departure from the
gene frequencies of wild populations [1]. Experience
with these hatchery stocks demonstrates that there are
no genetic barriers either to domestication or to re-
lease of hatchery individuals into natural populations.
Oysters
Abundant data are available for American oysters
of the genus Crassostrea [23] with the commercially
important species C. virginica [9] and the Pacific
species C. gigas [10,11] particularly well studied.
Buroker [8] has made an extensive study of electro-
phoretic variants in C. virginica from 10 oyster bars
throughout the main body of the Chesapeake. He
found statistically significant differences in gene
frequencies that permitted him to separate Bay
populations into four assemblages, much as perch and
striped bass can be divided. He recognized (1) an
upper Bay assemblage of Broad Creek, Patuxent
River, and Herring Bay oyster bars, (2) a mid-Bay
region including Swan Point, Wicomico River and the
Potomac Rivers, (3) a lower Bay group covering the
Tred-Avon River, Pocomoke Sound, and the Rap-
pahannock River, and (4) near the mouth of the Bay,
the James River as a distinct site. The differences
among the groups are subtle changes of frequency, not
unique variants in each assemblage, and the genetic
distances calculated by Buroker show very little
differentiation. Buroker particularly noted that the
differences in frequencies do not correlate with any
obvious environmental variable such as salinity, water
depth, or temperature. He concluded that these
assemblages result from different mortality patterns
among variants in different areas causing divergence
in gene frequencies that are prevented from becoming
distinctive local races by larval mixing each year.
Such a process is much like the one demonstrated in
Mytilus edulis.
Buroker's data apply to the main body of Chesa-
peake Bay. Rose [33] has determined gene frequen-
cies in samples of oysters taken in transects from the
mouths to the low-salinity limits of three Chesapeake
tributaries—the Potomac, Rappahannock, and James
rivers. Rose reasoned that if oysters did form local
races in response to environmental conditions, this
fact would be most obvious moving from the main
bay to the low-salinity limit of the species. As
Buroker had, Rose found significant differences in
gene frequencies, but these differences were slight and
followed no obvious pattern. No river had the
distinctive set of frequencies that one would expect if
the population within it were isolated. Neither did the
observed differences in frequency correlate with
distance from the mouth of the river, as one would
expect if salinity were an important selective agent.
Together, Rose's and Buroker's work clearly
demonstrate that oysters do not form distinct races
within the Bay.
Because oysters are extensively cultivated, they
offer special opportunities for selective breeding as
well as special problems in preventing loss of genetic
variation in captivity. Both Pacific and Atlantic
populations have suffered serious losses from disease
-------
82
Chapter 4: Genetics
in natural populations, and oysters have been the
object of intense selective breeding efforts on both
coasts. In the State of Washington, Hershberger and
his group [24] have worked to develop stocks of C.
gigas resistant to summer die-off. They report
considerable improvement after only a few genera-
tions of selection. Both these workers and Gosling
[21] have determined that hatchery breeding and
selection have not resulted in a substantial loss of
genetic variability. This encouraging finding suggests
that artificial selection can produce improved stocks
that still maintain most of the species's natural genetic
variation.
The most important disease of C. virginica along
the Atlantic coast and in Chesapeake Bay is infection
with MSX, (the protozoan Haplosporidium nelsoni).
This disease can cause mortalities of more than 50%
in a population, and an animal once infected does not
recover. A team based at Rutgers [20] has bred for
resistance to MSX for six successive generations of
oysters from Delaware Bay. Their stocks are slower
to develop infection even with severe challenge, and
mortality in infected stock is later than in unselected
lines. These stocks have not been tested for loss of
genetic variation, but the results with C. gigas
suggest that such loss will not occur. The hatchery at
the Virginia Institute of Marine Science has combined
use of wild broodstock with efforts to breed resisiant
oysters [12 andpers. comm.]. The oysters are
released as newly set individuals in an effort to
supplement natural recruitment. This operation
spawns healthy animals taken from natural popula-
tions in an area that has experienced very high levels
of infestation and mortality in the past. These oysters
should represent survivors of intense natural selection
while still representing native genotypes.
Blue Crab
Although the blue crab (Callinectes sapidus) has
great economic and cultural importance in the
Chesapeake Bay, very little genetic information is
available for this species. What is known suggests
that this species may be genetically similar over a
wide range. In the only study available, Cole [15]
compared crabs taken from the Choptank and
Patuxent rivers within the Bay and crabs taken from
Chincoteague Bay along the Atlantic coast of Vir-
ginia. While the straight-line distance between these
sites is small, the dispersal path for marine organisms
is enormous, going as it does from oceanic salinities
to brackish ones and involving more than 100 miles of
coastline. Cole found that although some differences
in allele frequencies were statistically significant, the
two samples were similar in overall genetic content.
He attributed this fact to extensive mixing of larvae
flushed into the mid-Atlantic Bight before reentering
the Chesapeake. His findings are encouraging, but
they require substantial amplification before any
management decisions can be made. Cole's sample
sizes were extremely small (only 24 crabs from each
site), and he reported that the resolution for many of
the enzymes was poor. Because these crabs are not at
present bred in captivity, questions of hatchery impact
are irrelevant. However, much more needs to be
known about genetic variation in natural populations
before any conclusions are made about natural
subdivisions.
CONCLUSIONS AND RECOMMENDATIONS
All of the species reviewed here show a similarity
in what is known of their genetic structures. Each
shows a moderate amount of differentiation suggest-
ing that populations in the upper, middle and lower
reaches of the Bay develop slightly different adapta-
tions. However, none of the species shows evidence
of forming highly localized races adapted to fine
gradations in the environment. From a manager's
point of view, these findings mean that biological
resources in the bay can be treated as a single unit if
necessary, but that best practice is to recognize broad
subdivisions. Although the genetic data give no
special insight into what forces are endangering
species in the bay, the results do show clearly that no
damage is likely from reliance on captive breeding to
supplement recruitment. Managers should continue to
give strong support to the operation of hatcheries and
to research on hatchery techniques for all important
species.
For fish, future research should obviously focus
on the endangered striped bass. This species shows
clear evidence that its dramatic decline in population
size has caused loss of genetic varitation. Hatchery
operators should therefore take great care to have as
many parents as possible contribute their genes to any
fish bred for release into the wild. Genetic monitoring
to ensure that inbreeding does not increase is very
desirable. In conjunction with hatchery efforts, a
study should go forward immediately to confirm or
disprove the mtDNA evidence that a small proportion
of adult fish contributes disproportionately to recruit-
ment. Techniques exist to slow the rate of inbreeding
in the smallest of populations [35], but these methods
should not be applied until it is known whether great
differences in fecundity exist between genetic
variants.
Although this review has concentrated on large,
economically and esthetically important species such
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Adamkewicz, Chapman, and Powers
83
as the striped bass, no one should forget that plants
and many individually insignificant animal species
form the ecological foundations of the Chesapeake
Bay. No amount of effort will sustain healthy
populations of bass if that foundation is destroyed.
For this reason, some research effort must go to
obscure species rather than all effort being concen-
trated on only a few organisms. The program to
protect the Bay would benefit from study of important
forage species such as herring, anchovies, and
menhaden.
Among invertebrates, the release of hatchery-
bred animals to supplement natural recruitment is
essential for oysters. The present research on remote
setting of oyster larvae [12] should continue to be
given a very high priority in combination with the
development of disease-resistant broodstock. The
efforts to breed MSX-resistant animals should be
supplemented with efforts to develop resistance to
another diseases and pollutants. The greatest assis-
tance to existing programs of genetic selection for
disease resistance would come from the development
of a standardized assessment of larval performance
and infection/resistance to MSX and other agents.
Once such an assay exists, it should be used in a
vigorous selective breeding program. Continued
support should be given to programs that investigate
the mechanisms of infection and mortality in oysters.
The facts that oysters do not form highly differenti-
ated populations and that they are routinely cultivated
in hatcheries without loss of genetic variation offer a
great opportunity to produce and release genetically
improved animals without fear of damaging natural
populations.
The need for basic research in population
genetics is obviously greatest in blue crabs. While
research on oysters has progressed into the area of
practical applications, work on the genetics of crabs
has been almost nonexistent. A survey of basic
genetic variability, fitness differentials, and geo-
graphic differentiation should begin immediately so
that data are available before management decisions
are necessary. This information will be important
when, as seems inevitable, new decisions about
restriction of harvesting and protection of breeding
stocks must be made. The most productive research
would be an initial study of electrophoretic variation
to discover whether locally adapted populations do
exist. Such a study could examine a far larger number
of individuals in a relatively short time than could a
DNA-based survey, and Cole's work shows that such
studies are possible. Any pattern of differentiation
that the electrophoretic study reveals can then serve as
a guide to the most promising areas of investigation
for later DNA work.
Genetics can make a real contribution to the
management and conservation of living resources in
Chesapeake Bay. Because decisions are already being
made that affect the genetic structure and perhaps
future evolution of species in the bay, an informed
understanding of these issues is essential. No amount
of genetic manipulation will compensate, however, for
lethal degradation of the environment. Therefore the
most important task is simply to make the bay a
cleaner, healthier habitat for all estuarine organisms.
ACKNOWLEDGEMENTS
We thank Janice Meadows, assistant librarian,
and other staff members of the Virginia Institute of
Marine Science for their assistance in the preparation
of this review. We are also grateful to the many
colleagues mentioned throughout these pages who
generously shared their unpublished results with us.
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and Saccostrea. Mar. Biol. 54:171-184; 1979.
12. Castagna, M. Mollusk culture for the Chesapeake
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of male striped bass, Morone saxatills, spawning
in the three areas of the Chesapeake Bay from
1984 to 1986. Fish. Bull. 85:167-170; 1987.
14. Chapman, R.W.; Powers, D.A. Mitochondrial DNA 30.
analysis of striped bass, Morone saxatilis,
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15. Cole, M.A.; Morgan, R.P. Genetic variation in two 31.
populations of blue crab, Callinectes sapidus.
Estuaries 1:202-205; 1978.
16. DiMichele, L.; Taylor, M.H. The mechanism of
hatching in Fundulus heteroclitus: development
and physiology. J. Exp. Zool. 217:73-79; 1981. 32.
17. DiMichele, L.; Powers, D.A.; DiMichele, J. Develop-
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variation at enzyme synthesizing loci in Fundulus
heteroclitus. Amer. Zool. 26:201-208; 1986. 33.
18. Endler, J.A. Geographic variation, speciation, and
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19. Epperly, S.P. A meristic, morphometric, and 34.
biochemical investigation of Atlantic menhaden,
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patterns in strains of oysters Crassostrea 35.
virginica selected for resistance to the parasite
Haplosporidium nelsoni (MSX). J. Parasit.
73:368-376; 1987.
21. Gosling, E.M. Genetic variability in hatchery 36.
produced Pacific oysters (Crassostrea gigas
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22. Grove, T.L.; Berggren, T.S.; Powers, D.A. The use
of innate genetic tags to segregate spawning
stocks of striped bass (Morone saxatilis). In: 37.
Cronin, L.E. (ed.). Estuarine Processes, pp 166-
176. Academic Press, New York, 1976.
Hedgecock, D.; Okazaki, H.B. Genetic variation
within and between populations of American
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1984.
Hershberger, W.K.; Perdue, J.A.; Beattie, J.H.
Genetic selection and systematic breeding in
Pacific oyster culture. Aquaculture 39:237-245;
1984.
Hilbish, T.J.; Koehn, R.K. Exclusion of the role of
secondary contact in an allele frequency cline in
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1985.
Koehn, R.K.; Hilbish, TJ. The adaptive importance
of genetic variation. Am. Scientist 75:134-141;
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Manseuti, R.J. Movement, reproduction and
mortality of the white perch (Roccus americanus)
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Morgan, R.P.II; Koo.T.S.Y.; Krantz, G.E. Electro-
phorctic determination of populations of striped
bass, Morone saxatilis, in the upper Chesapeake
Bay. Trans. Amer. Fish. Soc. 102: 21-32; 1973.
Mulligan, T.J.; Chapman, R.W. Stock identification
of white perch, Morone americanus, based on
mitochondrial DNA analysis. ICES publication
C.M. 1986/M-.20.
Powers. D.A.; Ropson, I.; Brown, D.C.; Van-
Beneden, R.; Cashon, R.; Gonzales-Villasenor,
L.; DiMichele, J.A. Genetic variation in Fundu-
lus heteroclitus: geographic distribution. Amer.
Zool. 26:131-144; 1986.
Powers, D.A.; Dalessio, P.M., Lee, E.; DiMichele, L.
The molecular ecology of Fundulus heteroclitus:
hemoglobin-oxygen affinity. Amer. Zool.
26:235-248; 1986.
Rose, R.L. Genetic variation in the oyster Crassos-
trea virginica (Gmelin) in relation to environ-
mental variation. Estuaries 7:128-132; 1984.
Russell, H.A.; Jefferies, J.E. Serum transferrin
polymorphism in grey trout, Cynoscion regalis,
from the lower Rappahannock River. Estuaries
2:269-270; 1979.
Schonewald-Cox, C.; Chambers, S.; MacBryde, B.;
Thomas, W. editors. Genetics and Conservation.
The Benjamin/Cummings Publishing Company,
Inc., Menlo Park, CA 1983.
Setzler, E.M.; Boynton, W.R; Wood, K.V.; Zion,
H.H.; Lubbers, L.; Mountford, N.K.; Frere, P.;
Tucker, L.; Mihursky, J.A. Synopsis of biological
data on striped bass, Morone saxatilis. FAO
Synopsis. 121: 69 pp.; 1980.
Sidell, B.D., Otto, R.G.; Powers, D.A.; Karweit, M.;
Smith, J. A reevaluation of the occurrence of
subpopulations of striped bass (Morone saxatilis,
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Walbaum) in the Upper Chesapeake Bay. Trans.
Amer. Fish. Soc. 109:99-107; 1980.
38. Sidell, B.D.; Otto, R.G.; Powers, D.A. A biochemi-
cal method for distinction of striped bass and
white perch larvae. Copeia, No. 2, pp. 340-343;
1978.
39. Skibinski, D.O.F. Mitochondrial DNA variation in
Mytilus edulis L. and the Padstow mussel. J.
Exp. Mar. Biol. Ecol. 92:251-258,1985.
40. Sullivan, J.B. Stock identification of Atlantic
croaker, Micropogon undulatus, and bluefish,
Pomatomus saltatrix, using biochemical
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41. Williams, G.C.; Koehn, R.K. Population genetics of
North American catadromous eels (Anguilla). In
Turner, B J., editor. Evolutionary Genetics of
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Sci. 3:94-113; 1962.
-------
Chemical and Physical Processes Influencing Bioavailability
of Toxics in Estuaries
James G. Sanders and Gerhardt F. Riedel
The Academy of Natural Sciences
Benedict Estuarine Research Laboratory
Benedict, Maryland 20612
INTRODUCTION
The transport, availability, and impact of toxic
substances in the Chesapeake Bay are matters of
major concern to those charged with the management
of the Bay. The human attraction to the coastline
surrounding the Bay ensures that population and
industrial growth will continue, and with it the loading
of toxic substances. It is imperative, therefore, that
we as a regional society address the potential toxic
effect of anthropogenic inputs to the Chesapeake Bay.
A number of factors, however, control the transport
and impact of such substances. This chapter ad-
dresses several such factors, primarily the chemical
and physical controls of pollutant availability.
Among the many factors altering the bioavaila-
bility of toxic substances, we may generally consider
two major groupings. In the first group are processes
that change the availability of the contaminant without
altering the total amount present. This group of
factors includes organic and inorganic complexation,
ionic strength, pH, redox reactions, and competition
with similar compounds. In the second group are
processes that do alter the concentration of the
contaminant within the system. Factors in this group
include sources and sinks, such as adsorption,
flocculation, sedimentation, and remobilization.
We also include a separate section discussing the
direct and indirect effects that organisms have on the
bioavailability of these compounds, even though
individually these effects could be legitimately be
considered as belonging to the first or second group of
factors.
General Mechanisms of Uptake
It is useful to consider groupings of compounds
that have similar patterns of availability to organisms.
A focus of recent research has been the effect of
chemical form, or speciation, of trace elements on
their availability to aquatic biota. Generalizations
concerning bioavailability are available for three
groups of contaminants: cationic trace elements,
anionic trace elements, and organic compounds
(Figure 1).
For cationic elements, free ion activity appears to
be the major factor determining bioavailability [4, 5,
36,61,96, 125, 150, 152, 153, 181]. The hypothesis
that the free ion activity represents the sole (or most
important) biologically available form can be derived
from a thermodynamic model in which the ion is
complexed by the active site of the enzyme where it
interacts or where it passes across the cell membrane.
Other chemical forms, however, may be available.
For example, an uncharged complex may provide a
second route of access to the cell. For grass shrimp
(Palaemonetes pugio) [37] and several species of
phytoplankton [132, 133], it appears that the calcu-
lated activity of the uncharged species AgCl° better
explains the availability and toxicity of silver than
does the calculated free silver ion activity.
For anionic trace elements (e.g., chromate,
molybdate, selenate, arsenate, and germanate),
bioavailability is most often controlled by competition
for uptake with a similar nutrient ion of greater
abundance (i.e., sulfate, phosphate, or silicate). For
those elements whose availability varies inversely
with sulfate (e.g., chromate [119,120], selenate [136,
170], molybdate [26, 28]), salinity plays a dominant
role in their bioavailability. Arsenate behaves as a
phosphate analog [18, 115, 126], and germanate as a
silicate analog [9], so the availability of these ele-
ments varies according to the concentrations of their
-------
88
Chapter 5: Bioavailability of Toxics
ANIONS: COMPETITION
FOR TRANSPORT SITES
CrO42- SO4
2-
CATIONS:
TRANSPORT CuCO
OF FREE IONS
CuL
%^
CrO4
2-
INGESTION
OF PARTICLES
CONTAINING
CONTAMINANTS
PCB
NEUTRAL ORGANICS: SOLUBLE
IN LIPID BILAYERS
Figure 1. Dominant routes of uptake for three important contaminant classes.
analogs, which in turn are highly dependent on
biological activity.
Many of the toxic organics are neutral xenobiotic
compounds, and accumulation is presumed to occur
via passive diffusion through semipermeable lipid bi-
layer membranes, such as epidermis, gills, gut wall, or
in the case of microorganisms, cell membranes [117].
Therefore, the most available form for such com-
pounds is often the un-ionized form. Ionized forms,
as with trace elements, must rely on active transport
sites, for which evolution has not provided many
models. For neutral molecules, the lipid solubility of
compounds is commonly measured in terms of their
octanol/water partition coefficient (Pow). This coeffi-
cient correlates well with bioaccumulation for a wide
variety of organics, including polyaromatic hydrocar-
bons (PAH's), polychlorinated biphenyls (PCB's),
and neutral pesticides [24, 55, 97,108, 145,163].
CHEMICAL FACTORS
In the Chesapeake Bay, salinity varies from fresh
water at the river heads to 30 parts per thousand (ppt)
at the entrance of the Bay; many other chemical
parameters show similar gradients or local variations
due to biological activity. Therefore, salinity and
changes in other chemical parameters are major
factors in the availability of many substances to
organisms.
Ionic Strength
In concentrated salt solutions, the activity of an
ion is reduced because of the interaction of its charge
with other nearby ions. In general, the individual ion
activity coefficient ranges from 1.1 for uncharged
species to approximately 0.1 for triply charged species
in full-strength seawater [148]. In practice, changes in
bioavailability due to ionic strength alone must be
considered in conjunction with changes due to
inorganic and organic complexation. Powerful
computers have made these calculations routine;
several computer programs exist to calculate specia-
tion (see [102] for a review of several such programs).
Inorganic Complexation
A number of ions are present in seawater in
almost constant ratio to one another. Although these
ions (sodium, magnesium, potassium, calcium,
strontium, chloride, sulfate, bromide, and borate) and
carbonate, bicarbonate, hydrogen, and hydroxide ions
(which are not conservative, but rather vary somewhat
with pH and biological activity), are not particularly
strong complex formers, their concentrations in
seawater are high enough that the most predominant
chemical species of most trace elements and many
other inorganic ions are ion pairs or other complexes
with these ions. Other ions are of biological origin
(e.g., sulfide) and have wide variation independent of
salinity. They may be important, however, in deter-
-------
Sanders and Riedel
89
mining the chemical speciation of inorganic pollutants
under some circumstances.
Complexation of the various inorganic pollutants
differs greatly. For example, the inorganic speciation
of copper is dominated by complexation by hydroxide
and carbonate species; it is only slightly complexed by
the very abundant chloride ion [149]. However, silver
and cadmium have comparatively high binding
constants for chloride, and their inorganic speciation
is dominated by chloride complexes [37, 76,133].
The binding of various ions is a highly interdependent
process since they compete for common ligands;
however, the majority of the important equilibrium
constants are well enough known that computer
models can readily estimate the activity of most ions
from their measured concentrations.
Organic Complexation
In addition to the inorganic ligands present in
aquatic systems, organisms produce, through excre-
tion, leakage, or decay, a variety of organic com-
pounds that may also complex and reduce the free ion
activity of trace elements. Levels of dissolved organic
compounds (DOC) range from 0.1 mg/1 in unpolluted
fresh water to 1-2 mg/1 in seawater to >10 mg/1 in
highly productive or polluted water. Concentrations
in the Chesapeake Bay range from 1.9 to 13 mg/1 [98,
99,128, 142, 181]. The composition of the organic
matter varies depending on the source, but it contains
many functional groups with metal complexing
capabilities. Unlike the situation with inorganic
complexing ions, natural organics contain a large
number of compounds of varying complexing
strengths [34, 38]. Therefore, except for a few studies
on specific chelators found to be excreted by algae, or
released into the environment by human activity, the
studies of complexation by natural DOC are largely
empirical. Many biological and analytical methods of
measuring the organic fractions of metals (or alterna-
tively the metal-complexing capacity) have been
devised [e.g., 86,99], some of which are specific for
certain types of complexes (e.g., polar vs. non-polar
organics). It is extremely difficult to compare one
method with another, and one site with another.
Only a few metals (iron, copper, zinc, and
mercury, which have the greatest affinity for ligands)
are complexed significantly by natural DOC. Some
species of algae can secrete special iron-chelating
compounds (siderophores), which solubilize iron and
facilitate uptake [157]. The types of algae that exhibit
this behavior include dinoflagellates [157], cyano-
phytes [83], and diatoms [158], which are common in
the Chesapeake Bay, but the extent of production of
these compounds in the Bay and their importance to
the system as a whole is unknown. Copper is also
organically complexed to a large extent (50-98%) [56,
92,151,187] by a variety of dissolved organic
compounds, some of which are relatively labile and
some of which are kinetically inert. In fresh water,
where the source of DOC is highly variable, there is
little correlation between DOC and copper complexa-
tion capacity between different sites; however, in
marine and estuarine waters, including the Chesa-
peake Bay, where the DOC is largely of autoch-
thonous origin and is more homogeneous on spatial
and temporal scales, there is a significant correlation
between DOC concentration and copper complexation
capacity [99, 160].
Lipid-soluble copper complexes have been
demonstrated to have exceptionally high bioavailabil-
ity and toxicity [8, 19,40]. The organic compounds
used in these studies were synthetic, however, and the
existence of similar compounds in the environment is
unconfirmed. This mechanism could be important in
the event of combined trace element and organic
pollution.
The uptake of toxic organics can also be modified
by the presence of natural dissolved organic com-
pounds. The uptake of DDT by Daphnia magna from
water is reduced by association of the DDT with
dissolved humic material (S. Friant, pers. comm.).
Presumably, the DDT-humic complex is more soluble
and less lipophilic than DDT alone, so the uptake of
DDT is controlled by the remaining unbound DDT.
PH
Many of the inorganic compounds important in
the complexation of some toxic metals (e.g., carbon-
ate, phosphate, and sulfate), as well as many of the
pollutants, have equilibria dependent on pH within the
pH range of aquatic systems [184]. Control of the
relative abundance of such species through these
equilibria is perhaps the most direct way in which pH
determines the bioavailability of pollutants. In
addition, the active site of most organic complexing
agents are pH-sensitive groups (e.g., carboxylic acids,
amines, and sulfhydral residues), so that pH has a
strong influence on the extent to which organics
complex toxic trace elements [86]. Finally, the active
sites of enzymes and trace element uptake sites
contain the same variety of pH-sensitive groups, and
the activity of these systems varies with pH as well.
In the case of enzyme systems, most have an optimal
configuration over a narrow range of pH [137]. Thus,
the effect of pH on the availability of a given com-
pound to a particular organism in a specific site
depends on a variety of effects, some of which (e.g.,
the effect of pH on inorganic complexation) are rather
-------
90
Chapter 5: Bioavailability of Toxics
straightforward to predict, while others (the effect of
pH on organic complexation and on uptake) are rather
difficult
The input of acidic precipitation to the Chesa-
peake Bay watershed may be an important process.
Rainfall with a pH as low as 3.8 has been measured
within Maryland; within small streams such rainfall
causes a pH reduction of up to 1.1 pH unit within four
hours [6, 7]. Such events may lead to the mobilization
of metals, or changes in the uptake of metals by
organisms, but the duration of such events is limited,
and their importance to the ecosystem is still being
investigated.
pH can also play an important role in the bio-
availability of ionizable toxic organic compounds.
The uptake and toxicity of pentachlorophenol to
goldfish (and probably most organisms), for example,
is determined by pH, because of the equilibrium
between hydrogen ion, pentachlorophenol, and the
pentachlorophenolate ion [79]. Atrazine, an herbicide
of interest in the Chesapeake Bay, is a slightly basic
compound, so its solubility and bioavailability are also
dependent on pH [177].
Redox Reactions and Equilibria
Several of the toxic trace elements participate in
redox reactions that alter their biological availability.
Most of the important redox reactions of trace
elements are relatively slow compared to inorganic
and organic complexation reactions. Most of the
important redox couples are directly or indirectly the
result of the activity of organisms, for without the
constant formation of reduced compounds and oxygen
by autotrophs from external energy, the world would
eventually reach redox equilibrium.
In the sediments and bottom waters of the
Chesapeake Bay, a strong oxygen demand is fueled by
organic carbon from the highly productive surface
waters. This demand produces anoxic, sulfide-rich
interstitial water and seasonally anoxic bottom waters
in the deep channel of the Bay. In such waters a
variety of metals (arsenic, chromium, iron, manga-
nese, and selenium) can be reduced to forms different
from the commonly oxidized form. Arsenic, for
example, is most stable in oxidized waters as the
arsenate ion. However, in anoxic bottom and intersti-
tial waters, arsenate is largely reduced to arsenite
[111, 122,129]. Arsenite in the sediments and bottom
waters can diffuse or be mixed up into the surface
waters, where it slowly oxidizes back to arsenate.
Arsenate is primarily available and toxic to phyto-
plankton, which have a requirement for phosphate,
while arsenite is more toxic to fauna [94,109].
Selenium, which has selenate and selenite as the
major oxidized and reduced species, probably
undergoes a similar cycle, although selenite is
apparently more stable under oxidizing conditions
than arsenite [31]. Selenite is generally the more
biologically active of the two forms of selenium,
satisfying trace requirements at lower concentrations
[116]. Anoxic waters also reduce the highly toxic (but
relatively unavailable) chromate ion to chromite,
which is much less toxic.
Indirect evidence suggests that sulfide is present
in oxidized seawater and in the Chesapeake Bay at
concentrations of 10'10 to 10"11 M (G. Cutter, pers.
comm.) [35], enough to have significant effects on the
speciation of trace metals such as silver and mercury.
However, standard analytical techniques are not
sensitive enough to measure sulfide at those concen-
trations. Moreover, in the Chesapeake Bay, where
there is occasional anoxia and a constant source of
sulfide diffusing from the sediments, concentrations
are likely to be more variable.
Another source of reduced metal ions is photo-
reduction. The interaction of light or ultraviolet
radiation on DOC or on metal organic complexes has
been shown to directly reduce certain metal ions and
to produce redox-active compounds such as peroxide
or superoxide that can reduce metals. It has been
shown that peroxide concentrations in the Patuxent
River undergo a diurnal cycle, with concentrations
increasing in the daylight hours and decreasing at
night [77]. Peroxide, in turn, has been shown to
reduce chlorine from industrial sources to chloride
[63], reduce some copper(II) to copper(I) [93], and
reduce and solubilize manganese dioxide to man-
ganous ion [154]. DOC and light can reduce ferric
iron to ferrous iron [67, 166]. Chromate is reduced
photochemically in the presence of high DOC levels
[121] (G. HelzandR. Kieber, personal communica-
tion), but it is unknown whether the photoreduction
proceeds through peroxide or a similar intermediate.
The importance of photoreduction in modifying the
biological and chemical behavior of pollutants is not
yet clear; however, since photochemical reactions are
favored by the presence of high DOC and abundant
light, shallow, organic-rich estuarine systems such as
the Chesapeake Bay are likely to be sites where it is
most important.
Organo-metallic Compounds
Several metals (antimony, arsenic, lead, mercury,
selenium, and tin) are present in the Chesapeake Bay
partly as covalently bound compounds (e.g., methylar-
sonic and dimethylarsinic acids [129,130], alkyl lead,
methylmercury [185], organic selenium [156], methyl
and butyltin compounds [47, 52, 53, 54, 169]). In
-------
Sanders and Riedel
91
SURFACE MICROLAYER
RIVER_
INPUT
SEAWARD
ADVECTION
(f^ f .. *PV*HXNP1
ADSORPTION •* •• DESORPTION
FLOCCULATION
REMOBILIZATION
4
SEDIMENTATION
J.ANDWARD
ADVECTION
Figure 2. Processes affecting transport and availability in the Chesapeake Bay.
some cases (methylarsenic compounds, methyl
mercury, organic selenium, and methyltins), these
compounds are formed in the environment from
inorganic metal by the local biota. In others (alkyl
lead and butyltins) the compound is primarily anthro-
pogenic, and the biota ultimately act to degrade the
compound [106]. Metals can also be methylated by
exchange reactions with other methylated compounds
[13,20,85].
There are few direct comparisons of the bio-
availability of alkyl compounds and their inorganic
precursors, but in general it appears that their trans-
port is enhanced by their greater lipid solubility. For
example, in a study of the uptake of inorganic and
methylmercury by the diatom, Skeletonema costatum,
and the copepod, Acartia clausi, exposure to methyl-
mercury resulted in much greater uptake of total
mercury than exposure to inorganic mercury [43].
Certainly, toxicity is enhanced [23]. Methylated tin
species are more toxic than inorganic forms, and di-
and tri-substituted alkyls are more toxic than mono-
substituted ones (JJ. Cooney, pers. comm.). For
inorganic and organo-tin compounds, there is a
general correlation between the P^ an index of the
lipophilicity of the compound, and the toxicity to
phytoplankton [174] and animals [186]. In Chesa-
peake Bay sediments, concentrations of tin are high
enough to affect the microbial community, inhibiting
sensitive species and probably selecting for resistant
species and those able to methylate tin [53, 54].
As an example of how chemical factors combine
to affect the bioavailability of contaminants in the
Chesapeake Bay, consider the following. A strong
negative correlation has been observed between
salinity and the copper uptake and accumulation in
natural Patuxent River oysters. This observation fits
the expectation of increased copper availability at low
salinity due to low ionic strength and less inorganic
complexation [176,181]. A similar negative correla-
tion between salinity and copper in oysters has been
noted Bay-wide [72]. Further research, however,
using nitrilotriacetic acid-copper buffer systems, has
shown a greater effect due to salinity than could be
explained on the basis of free ion activity alone, which
might be explained by competition for uptake with
calcium or magnesium.
CONTAMINANT SOURCES AND PROCESSES
AFFECTING DISTRIBUTION
In addition to the above-discussed factors which
alter the bioavailability of toxic substances, there are
also factors that control the total concentration to
which organisms are exposed, factors that relate to the
sources, sinks, and cycling of toxics in the Chesa-
peake Bay. These are discussed below and are dia-
grammed in Figure 2.
Sources
There are several sources of toxic substances to
the Chesapeake Bay (Table 1). One of the most
important is runoff of fresh water from the land.
Industry, agriculture, and municipal effluents each
contribute a significant load of compounds. Many
toxic substances are present in the air as suspended
particles (e.g., copper, zinc, and PAH) or as vapors
(e.g., mercury, lead, selenium, arsenic, and volatile
organics), and can be deposited by rain or through dry
fall. Sampling aeolian deposition, however, is a
difficult exercise, as it is extremely variable in time
and space. Aeolian flux of some metals has been
estimated for the Chesapeake Bay (Table 1) and the
nearby Delaware Bay [25], and while the loadings are
not large compared to runoff, they are significant.
Aeolian inputs may be a particularly significant
source to the surface microlayer (see below) where a
variety of toxics are concentrated, and to areas remote
from industry or other concentrated sources.
-------
92
ChapterS: Bioavailability of Toxics
Table 1. Estimated sources of toxic trace elements, in metric tons per year, entering the Chesapeake Bay. Taken
fromBierietal. [17].
Source
Cd Cr
Cu
Ni
Zn
Fe
Mn
Major tributaries
Industry
Municipal wastewater
Urban runoff
Shore erosion
Wet fall
75
178
6
7
1
3
551
200
200
10
83
ND
517
190
99
9
29
28
402
ND
ND
20
ND
25
307
155
68
111
28
34-
1444
167
284
63
3
825
199683
2006
625
977
57200
87
19000
ND
ND
22
ND
15
ND=not determined.
Distribution
The Chesapeake Bay and similar estuaries are
extremely dynamic systems, moving and changing
constantly in response to winds, tides, and runoff, so it
is difficult to discuss the distribution of a given toxic
compound in the water except on a statistical basis.
Nevertheless, some general trends in contaminant
distributions can be discussed. Compounds whose
source is in fresh water are inversely correlated with
salinity, i.e., mean concentrations decrease from the
head of the Bay to the ocean, and from tributaries to
the main Bay. This pattern has been observed for
metals such as cadmium, copper, and lead [78].
Conversely, for contaminants whose concentrations in
seawater are normally higher than those in fresh
water, a reverse trend is observed, with bottom water
and more saline waters having higher concentrations
[78,129]. Elements that are reduced and solubilized
in sediments (e.g., iron, manganese, arsenic, and zinc)
also tend to be more concentrated in deep waters of
the Bay [22, 78, 129, 159].
The phenomenon of seasonal anoxia in the
Chesapeake Bay is undoubtedly a dominant factor in
the spatial and temporal distributions of redox-
reactive elements such as arsenic, chromium, iron,
manganese, and selenium. Unfortunately, virtually no
published studies on this aspect of the anoxia are
available yet. This should be an active area of
research in the near future.
Surface Films
Contaminants can concentrate in two different
areas within water bodies: the sediment and the
surface microlayer. The surface microlayer is the
upper 50 jam to 1-mm surface film that lies between
the water and the atmosphere. Many natural organic
compounds and contaminants concentrate there, due
to their hydrophobic nature [11,57, 58]. This film
also contains a variety of particles, including a unique
fauna and flora [58]. In addition, the presence of
natural organics ensures the presence of trace con-
taminants that normally bind to organic complexing
agents [44].
An interesting question for research is whether
the high concentrations of contaminants found in the
surface microlayer pose a particular threat to any part
of the ecosystem. The higher concentrations in the
microlayer are offset to some extent because the
processes that cause the concentration there also
inhibit uptake by organisms.
Recent studies in the Chesapeake Bay have found
high concentrations of many pollutants, including
metals, alkanes, and aromatic hydrocarbons in the
surface microlayer [12,50, 52,59]. Atrazine was
found to concentrate by a factor of up to 110 in the
surface microlayer of the Chesapeake Bay [177, 178].
When seagrass microcosms received additions of
atrazine dissolved in water, associated with particles,
or as a component of surface film, differences in
toxicity were not observed [29].
Because of the high concentrations and the
potential for direct uptake by biota, this layer may
represent an efficient and important transfer mecha-
nism for pollutants. This process might be particu-
larly important for periodically emergent organisms,
such as sedentary animals in the tidal zone or emer-
gent vegetation or neuston, including commercially
important fish eggs. Certainly, concentration in the
microlayer reduces to some extent the exposure of
organisms that do not come into that boundary, and
may expose the contaminant to a higher rate of
removal through photooxidation or adsorption. For
example, substantial photooxidation of benzanthracine
and similar compounds has been observed in surface
waters [81].
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Sanders and Riedel
93
Adsorption
Adsorption of inorganic pollutants to a particle is
analogous to complexation as discussed in the
previous section, and is subject to most of the same
controls, including the effects of pH, ionic strength,
and competing ions. Adsorption reactions may be the
important controlling factors for the distribution of
many toxic trace metals in estuaries, including copper,
lead, and zinc [60]. These reactions are partly
reversible, so that the material can be desorbed from
solids by lowering the concentration of the pollutant
in solution, or by changing the chemical system to
favor the release (a change in pH, or an increase in
competing ions, for example).
Adsorption of organic compounds to suspended
particulates is a function of the hydrophobic interac-
tions much like the binding of organics to cells [87],
so that nonliving particles compete with living cells
for these substances as well. In a study of the fate and
effects of the herbicide atrazine in microcosms with
Chesapeake Bay vegetation, 78% of the atrazine
applied became associated with either suspended
particulates or sediments [29]. A number of herbi-
cides have been shown to bind strongly to colloids in
the Chesapeake Bay [88, 89]. Preferential binding to
such small particles will greatly affect the transport
and eventual fate of such compounds, allowing longer
residence times within the water column.
Flocculation
Suspended solids and colloids carrying adsorbed
trace elements or organics can aggregate to form
larger particles, which become more susceptible to the
processes of sedimentation and filtration. An impor-
tant component of the suspended solids in most river
systems is clay particles. In fresh water, the aggrega-
tion of clay particles is quite slow, whereas in
seawater, clay particles aggregate readily [148].
Flocculation can lead to segregation and concentration
of trace metals. In studies of metal interactions with
Chesapeake Bay floes, most metals were concentrated
in the smaller size fractions [45,46].
In a typical estuarine circulation, particles are
carried into the estuary by the river; as the water
moves downstream and becomes more saline, the
particles agglomerate, and eventually settle to the
bottom water. The bottom water, carrying more saline
water toward the river, returns the particulates
upstream, where the lower salinity tends to disaggre-
gate them. They may then reenter the surface water
through mixing and begin the cycle again. In this
way, a zone called the turbidity maximum is formed.
The high concentration of suspended solids in this
area may be both an important barrier to the move-
ment of toxic compounds through the estuary and a
primary reservoir of the same compounds. Because of
the high concentrations of toxics in this zone, it is
likely to be a primary site for any effects of the
toxicant.
Sedimentation
The ultimate repository of most particles is the
sediment. In particular, the sediments of the turbidity
maximum zone downstream of a major source are
often a major repository of toxic trace elements and
organic compounds. For example, high concentra-
tions of kepone (a halogenated pesticide released into
the James River in the 1960's and 1970's) have been
incorporated into the sediment in the James River
turbidity maximum zone, and buried under as much as
50 cm of cleaner, more recent sediment [62,69].
Within the Chesapeake Bay, Officer et al. [103]
and Helz et al. [65] determined that most of the
particulates and associated toxics entering from the
Susquehanna River are deposited at the head of the
Bay.
In an estuary a variety of solids may be important
in adsorption, flocculation, and sedimentation,
including clay particles, iron and manganese oxides,
sulfide minerals, organic films, and plant debris.
Since adsorption is a surface phenomenon, finer solids
have a greater adsorptive capacity than coarser solids
of the same type. Unlike complexation, adsorption
removes the pollutant from solution, and if the solid
leaves the water column, the pollutant is also re-
moved, and may become completely unavailable to
organisms in the water column. This process is no
doubt largely responsible for the high concentration
of toxics found in the sediments near sources of
pollution [16, 32, 33,49,70, 71, 90,107, 110,112,
143, 164, 165,173,175]. For example, the concentra-
tions of arsenic, copper, manganese, nickel, lead, tin,
and zinc in Chesapeake Bay sediments decline
seaward from maxima at the Baltimore Harbor,
Susquehanna River, and Elizabeth River [62, 100],
and a similar distribution is seen with PAH's and
PCB's [62, 70, 175].
Sedimentation and resuspension by storm events
may be a major factor in the sedimentation patterns of
a shallow estuary such as the Chesapeake Bay. It has
been estimated that 80% of the sedimentation between
1966 and 1976 occurred during two storm events and
that approximately 50% of yearly sedimentation
occurs during a short period in spring [103]. The
importance of such episodic events in controlling
contaminant availability is not well documented,
however, probably because of sampling difficulties.
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Chapter 5: Bioavailability of Toxics
Remobilizatfon
Incorporation into sediments and subsequent
burial does not forever remove toxics from possible
uptake by organisms. Many organisms reside in the
sediments and are thus exposed to the toxic material.
Toxic substances may be present in the pore waters of
sediment in higher concentration than in the water
column. For example, in experimental microcosms
the concentration of atrazine in pore waters was
fourfold higher than surface water concentrations
[29]. Because the benthic biota can be the majority of
the biomass in such a system, the net effect of
incorporation of atrazine into the sediments might
actually be a net increase in its availability to organ-
isms.
Other toxics, normally less available to organisms
in an oxidized state, can be reduced in anoxic sedi-
ments and their bioavailability enhanced. For
example, arsenic present in water as arsenate can be
adsorbed onto surface sediments. When buried in
reducing sediments, the arsenic is reduced to arsenite,
which is present in much higher concentration in the
interstitial waters than in the water column. In the
Patuxent River, the arsenite concentration in the pore
waters is up to 50 times higher than the arsenate
concentration in the surface waters [122]. Other
metals coprecipitated with iron and manganese
oxyhydroxides, such as copper, zinc, and lead, can be
remobilized in pore waters, particularly in the north-
ern Chesapeake Bay, where low sulfate concentrations
result in less sulfide formation under reducing con-
ditions [22, 159]. Cadmium also may be remobilized
from Chesapeake Bay sediments [64].
Toxic compounds in sediment pore waters can
also be returned to the water column by diffusion,
resuspension of the sediments, or activities of benthic
organisms. Detectable fluxes of arsenic from con-
taminated Chesapeake Bay sediments have been
measured, with a variety of chemical forms (arsenate,
arsenite, and methyl arsenic) detected [122]. Burrow-
ing organisms (Nereis succinea) at levels approximat-
ing natural abundances caused an enhancement of the
fluxes by a factor of five over molecular diffusion;
this biological enhancement was slightly less than the
enhancement caused by daily resuspension of 8 cm of
sediment, and more than the enhancement caused by
resuspension of the top 2 cm.
BIOLOGICAL PROCESSES
Many inorganic compounds are capable of
participating in the oxidative and reductive reactions
that comprise cellular metabolism. Some elements are
required in small quantities but toxic in larger
quantities. Both inorganic and organic substances can
be incorporated by the organism after uptake, and
partitioned according to their chemical structure.
Incorporation of a toxic compound markedly affects
the compound's transport through the Chesapeake
Bay, depending on the organism, its life cycle,
behavior, mobility, and persistence. Materials
associated with phytoplankton and pelagic micro-
organisms behave in a manner somewhat similar to
particulates, as discussed in the previous section.
Toxic compounds may pass to successively higher
trophic levels through predator/prey relationships.
Active participation of toxic substances in
biological processes can lead to chemical transforma-
tion of the compound into a form quite different from
the original. If these transformed chemical species
have biological affinities or chemical stabilities that
differ from the original compound, this transformation
can alter the compound's reactivity and perhaps its
transport through the estuary. In addition, biological
uptake and transformation may enhance or alleviate
its toxicity to other organisms, increasing or minimiz-
ing the potential for harm.
In addition to direct uptake and modification of
toxic substances, biological activities can substantially
modify the chemical environment surrounding them,
thus altering the bioavailability of some contaminants.
For example, algal photosynthesis, particularly during
bloom conditions, can significantly alter the pH and
the oxygen concentration of the water column.
Microbial activities within sediments and in stratified
water columns cause anoxia, altering redox condi-
tions. A number of organisms are capable of excret-
ing large quantities of organic compounds that can
complex toxic substances, altering their toxicity and
transport. The production of feces in higher trophic
levels leads to the aggregation and sedimentation of
particles and associated toxics. Finally, the movement
of infaunal organisms affects the cycling of contami-
nants between estuarine sediments and the overlying
water column.
All of the above processes, both those associated
with direct uptake and modification of toxic sub-
stances and those associated with biologically induced
changes in the environment, are discussed in more
detail below.
Uptake and Incorporation
The ability of marine organisms to accumulate
contaminants has been well documented; the literature
is replete with studies of pollutant levels in organisms
from the Chesapeake Bay [1,2,16,17,32, 33, 39, 51,
66, 70,74, 75, 82, 95,113, 114, 123,141, 146, 147,
161,162, 175,180]. Less well known, and currently
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Sanders andRiedel
95
under study, is the role of various processes that affect
contaminant availability and uptake. Organisms have
the ability to extract contaminants dissolved in the
water column, adsorbed on suspended particles and
detritus, and contained within their food. Uptake
occurs either by adsorption of the contaminant onto
cell surfaces or by absorption across body surfaces
such as gills and guts, or through a combination of
both.
For those organisms (primarily animals) that live
in a medium containing a dissolved toxic compound,
contact sediment with high concentrations of the toxic
substances, and ingest other organisms that have
already incorporated the toxic substance, the logical
query is which source contributes most to the uptake
of the toxic. The answer, of course, varies for
different organisms and pollutants; in general,
however, water appears to be the most important
source of most toxics to the greatest number of
organisms. Much of the metal present in food is not
passed across the gut wall. More lipophilic materials
such as alkyl compounds are generally more readily
incorporated from food, but they are also more readily
taken up from water [43]. The solubility of the most
lipophilic compounds (log P^ > 8) is such that water
is a poor source, and food becomes more important.
Generally speaking, uptake of contaminants from food
is more important for larger animals, for which
surface absorption is less important. Animals have
been shown to take up large amounts of metals from
contaminated sediments, but it is usually not clear
whether the materials have been ingested or absorbed
directly from interstitial water.
The uptake of silver by oysters from Chesapeake
Bay is solely from silver dissolved in water. Uptake
from either algal food or sediment is insignificant in
comparison [131, 134]. Copper uptake by the same
organism is more complex. Uptake of dissolved
copper occurs readily, largely controlled by the
availability of the free copper ion [182]. Copper is
also taken up from food, although to a lesser extent
[181]. It appears that colloidal organo-cupric com-
plexes are also available [183].
Through the predator/prey relationship, the
potential exists for pollutants to be passed up the food
chain from autotrophs to herbivores to carnivores. In
past years, such food chain transfer was thought to
lead to increased body burdens at each step in the
chain, resulting in top carnivores with extremely high
concentrations of toxic substances, toxic responses
within the population, and even potential toxicity to
human consumers. Recent work has shown, however,
that most contaminants do not get magnified as they
are passed up the food chain [179]. Available studies
from the Chesapeake Bay support this claim; accumu-
lation of kepone from the contaminated James River
showed no evidence of biomagnification, but rather an
increase in body burden proportional to the
individual's residence time in the vicinity of the spill
[15, 73], with little evidence of uptake from food [14].
In a study of the passage of kepone through a simple
plankton-mysid-fish food chain, spot (Leiostomus
xanthurus) obtained only 0.2% via food [10]. Grass
shrimp (Palaemonetes pugio) fed cadmium-"loaded"
brine shrimp (Anemia salina) obtained only 1.2% of
the cadmium that they would obtain from the water
used to load the brine shrimp [101]. Biomagnification
through food appears to be a significant factor for
very few compounds; DDT [84] and methylmercury
[42] appear to be notable exceptions. Again, biomag-
nification correlates with very high lipophilicity and
low biodegradability.
The incorporation of a contaminant can drasti-
cally alter its transport and impact within the Chesa-
peake Bay. In general, because organisms try to
maintain position within a specific environment,
incorporation within tissues will lead to the contami-
nant remaining within the estuary. For example,
silver has a high affinity for particle surfaces and is
rapidly taken up and incorporated by phytoplankton as
well as by suspended sediments. As these different
particles move down the estuary, the silver is
desorbed from the suspended particles and remains in
a dissolved state, complexed with chloride. Silver
associated with phytoplankton, however, does not
desorb, thereby remaining in the Chesapeake Bay and
probably recycling [132, 133]. Many organisms
within the estuary are sedentary; contaminants
incorporated by these species will accumulate within
the population's growth zone. Uptake and incorpora-
tion of toxics by migratory organisms such as pelagic
fish presumably will result in transport of Chesapeake
Bay toxics into other ecosystems. By the same
process, toxic substances can be transported into the
Chesapeake Bay region. On the whole, however, an
overwhelming percentage (97%) of the carbon
produced within the Bay remains within the Bay [144,
155]; very little is exported. Also, only a small
percentage of the total biomass of the Chesapeake Bay
system is composed of migratory species; thus any
transfer via this mechanism is likely to be minor. For
example, it has been estimated that less than 0.01% of
the arsenic contained within the coastal shelf off
Georgia was exported from the system by biological
uptake and subsequent migration [127].
Many marine animals detoxify trace metals by
binding the metals to metal-binding proteins. These
proteins have a high affinity for such metals as silver,
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96
Chapter 5: Bioavailability of Toxics
cadmium, copper, mercury, and zinc. A cadmium-
and copper-binding protein has been isolated from
American oysters (Crassostrea virginicd) [41]. The
ratio of cadmium to copper bound to this protein
varied considerably through the year, with higher
copper binding noted during the summer spawning
period. An inducible protein binding cadmium,
copper, and zinc has also been found in three Chesa-
peake Bay fish: striped bass (Morone saxatilis), white
perch (Morone americana) and channel catfish
(Ictalurus punctatus) [3]. Similar proteins appear
widespread, also occurring in cyanobacteria [105] and
algae [48]. These proteins not only reduce the toxic
impact of the metal on the individual organism, but
also sequester the metal from the environment and
reduce the chance that another organism will encoun-
ter the same metal. In addition, prior exposure to
sublethal levels of one metal can enhance an
organism's ability to withstand another toxic metal
through the induction of metal-binding proteins that
are broadly specific [124].
Transformation of Chemical Form and Modification
of Toxicity
Another important aspect of biological uptake is
the transformation of chemical form of a contaminant.
Aside from the concentration and sequestering of
toxic substances, organisms have the capability to
alter the chemical form or partitioning of many
inorganic elements and organic compounds. Such
shifts alter biological reactivity and toxicity of these
compounds and can alter their rate of transport
through the estuary as well as their eventual fate.
Potential metal/phytoplankton transformations are
illustrated in Figure 3. For example, biological uptake
of arsenic leads to the production of reduced and
methylated arsenic compounds, some of which are
more toxic to higher trophic levels than was the
original arsenic compound [94,109,127,129].
Within the Chesapeake Bay, large fractions (up to
80%) of the arsenic may be present in these reduced
or methylated forms [129]. The production of such
compounds also increases the rate of transport of
arsenic through the coastal zone by altering its
biological and geochemical reactivity [130].
Toiler et al. [185] measured elevated levels of
inorganic and organic mercury species around the
Chalk Point Steam Electric Station on the Patuxent
River, and concluded that inorganic mercury released
by the plant was being biologically transformed.
A number of microorganisms, particularly those
in anoxic sediments, are capable of methylating trace
metals. Within the Chesapeake Bay, a slow, stepwise
methylation of tin occurs in sediments [47]. Other
INCORPORATION
AND SEDIMENTATION
(Cr)
DISSOLVED
METAL ION
c.
LIGHT
RELEASED
ORGANOMETAL
(As, Hg)
RELEASED
REDUCED METAL
(As)
PHOTOREDUCTION
VIA EXCRETED ORGANICS
(Cr. Cu. Mn. Fe)
COMPLEXATION
BY EXCRETED ORGANICS
(Fe, Cu)
Figure 3. Possible interactions of dissolved metal ions
with phytoplankton, including changes in partitioning (A),
direct transformation of chemical speciation (B), and
indirect facilitation of complexation and photoreduction (C).
metals, such as mercury, lead, and arsenic, have been
shown to undergo methylation in similar aquatic
systems or in experimental enclosures (C.C. Gilmour,
pers. comm.) [23, 122, 172]; although not reported for
the Chesapeake Bay, these reactions presumably are
occurring.
Degradation
Degradation of toxic compounds by biota is an
equally important process. A number of organisms
have been shown to possess enzyme systems capable
of degrading trace organics, leading to their transfor-
mation and presumably their detoxification. Such
reactions are more complex than those discussed
above for inorganic compounds, as the initial organic
parent compounds are metabolized through several
steps into a variety of metabolites with varying
solubility, reactivity, and toxicity. Often degradation
is not complete. Four different petroleum products
showed very different rates of degradation when
presented to microorganisms and sediments from
polluted Baltimore Harbor [168]. None of the oils
was degraded completely; thus, the overall process is
quite uneven.
There is evidence that prior exposure to the same
pollutant selects for a microbial community better
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Sanders and Riedel
97
able to degrade new contaminants. Such selection
may, however, simply increase the rate at which a
substance is degraded, and not affect ultimate fate. In
a comparison of microorganisms from Baltimore
Harbor and the eastern Bay, Colwell and Walker [27]
found that although rates of degradation were much
more rapid in Baltimore Harbor sediments, similar
compounds were degraded eventually by the micro-
bial community from the eastern Bay.
Rates of degradation, as with most other micro-
bial processes, vary greatly. In general, temperature,
nutrient concentrations, and prior exposure appear to
be the most important factors limiting degradation
[27,106]. The presence of toxic substances within the
organic fraction can also be important; for example,
the preferential partitioning of metals such as mercury
in the oil component of sediments [135] can retard
degradation because of the toxicity of mercury to the
microbial community [167]. Some organic com-
pounds may be degraded through co-metabolic
processes. Phenanthrene is both degraded directly and
co-metabolized in the presence of a variety of carbon
sources in Chesapeake Bay sediments, and micro-
organisms capable of such degradation are present in
elevated numbers in sediments polluted with hydro-
carbons [28,80,140].
Herbicides associated with Chesapeake Bay
sediments were found to degrade at rates 2-30 times
the reported rates in soils and lake sediments, suggest-
ing that estuarine sediments have an increased
capacity for degradation of synthetic organics [89].
In general, although transformation of organic
and inorganic substances plays an important role in
contaminant transport and impact, the processes are
not well understood. Further studies are warranted to
fully understand the role of biological transformation
in the transport and reactivity of toxic substances.
Alterations of contaminant geochemistry by biota are
extremely important to the estuarine ecosystem as a
whole, and must be considered when impact predic-
tions and management assessments are made. Too
often, such biogeochemical concerns are ignored.
Modification of the Environment
In addition to direct uptake and modification of
contaminants, the activity of biota can substantially
modify the chemistry of the surrounding environment.
Algal photosynthesis results in the removal of
inorganic carbon from the water column and a
resulting increase in pH; during bloom events, such
increases can be substantial. In the freshwater section
of the Potomac River during 1984, a persistent bloom
of Microcystis sp. caused substantial variability in
pH and maximum pHs of >10 [138]. Such a large
shift in pH will greatly affect the geochemistry of
many trace substances, as discussed in CHEMICAL
FACTORS. In this instance, the persistently high pH in
the water column increased the mobility of phos-
phorus in the sediments and led to increased flux of
phosphorus across the sediment/water interface,
maintaining the high levels of algal growth and the
high pH. Although not measured, presumably other
pollutant fluxes were affected as well.
Large pH shifts are found not just within the
freshwater portions of the Chesapeake Bay; large
algal blooms can significantly affect pH even in areas
of moderate salinity. A persistent bloom of Katodin-
ium rotundatum in the mesohaline portion of the
Patuxent River during January-March 1987 drove the
surface pH of the river to above 9 on a daily basis, and
caused a maximum pH of >9.5, approximately 1.5 pH
units higher than normal (J.G. Sanders and SJ.
Cibik, unpublished data).
Another effect of biota is the variability in
dissolved oxygen concentrations. As a consequence
of nutrient enrichment, high algal productivity and
increased respiration rates drive large swings in the
diurnal oxygen cycle, with higher oxygen concentra-
tions during the day and lower concentrations during
the night In the Patuxent River, the daily swing in
oxygen concentration increased during the late 1960's
from approximately 3 mg/1 to 7 mg/1 [30]. Similar
trends have been shown for the entire estuarine
portion of the Patuxent River [91]. In stratified water
columns, high levels of respiration can deplete oxygen
concentrations. In the mainstem of the Chesapeake
Bay and some of the tributaries, water column strati-
fication during summer and high respiration rates lead
to depletion of oxygen in the bottom waters, often for
long periods [104,139]. Oxygen depletion and result-
ing sulfide formation causes large changes in the
mobility and complexation of some compounds, par-
ticularly metals that form insoluble hydroxide or
sulfide complexes (see CHEMICAL FACTORS).
Autotrophs have been shown to excrete DOC
during normal growth. Under conditions of high cell
densities, such excretions can become quite signifi-
cant. For example, concentrations of DOC commonly
vary from 1.9 to 6.0 mg/1 in the Patuxent River [99,
128, 142, 181]. During dense algal blooms, however,
concentrations are much higher (K.G. Sellner,
unpublished data). These very high concentrations of
dissolved organics have the potential to alter specia-
tion of both inorganic and organic pollutants. The
effects of these enhanced organic compounds should
be a focus of further research.
Organisms within the water column and on the
bottom produce aggregated fecal pellets containing
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98
ChapterS: Bioavailability of Toxics
unassimilated organic matter and inadvertently
ingested sediment particles. Fecal pellets provide
perhaps the most efficient mechanism for transfer of
both inorganic and organic pollutants from the water
column to the sediments, because they are produced at
a high rate, sink relatively rapidly, contain high
concentrations of contaminants, and break down
slowly (see review [42]). For example, studies of zinc
have indicated that rapid recycling between the
sediment and water column takes place in the northern
Chesapeake Bay. This cycle has been attributed to
uptake by phytoplankton, ingestion by zooplankton,
and rapid return to the sediments in fecal pellets [22].
In addition to providing a rapid mechanism for
transport, fecal pellets presumably alter the availabil-
ity of the pollutants as discussed in CONTAMINANT
SOURCES AND PROCESSES AFFECTING DISTRIBUTION.
Little work within the Chesapeake Bay, however, has
focused upon this problem.
Benthic infauna may change the distributions of
contaminants in sediments in a variety of ways,
including mechanical disturbance (mixing, pelletiza-
tion, and sorting), chemical changes (increased
oxygen penetration or organic enrichment), increased
microbial activity, the effects of ingestion on sediment
constituents, increased surface area, and the uptake of
metals by the organisms themselves [118]. Burrowing
activities of Nereis succinea in contaminated
sediments, for example, increased the flux of arsenic
by a factor of five [122]. This enhancement corre-
sponded to an approximately equal change in the
surface area of the sediment due to worm burrows.
SUMMARY AND RESEARCH DIRECTIONS
In this chapter, we have attempted to outline the
various chemical and physical factors that affect
contaminant input, distribution, and availability, and
to discuss our current knowledge of these factors
within the Chesapeake Bay. Ideally, we would like to
be able to summarize our discussion with a general
ranking of the various factors and their relative
importance to the movement of toxic substances. Un-
fortunately, such a summary is not possible. There
are many different compounds, each with a different
reactivity and behavior. Factors that control one
compound or group of compounds may well be
unimportant for another. All of the processes dis-
cussed here are important for some contaminant, but
no one process is a major factor for all contaminants.
The best that we are able to provide here is a general
survey of different kinds of compounds and the
processes that are important in determining their
distribution and rate of transport
In contrast, the general pathways of biological
uptake for most compounds can be summarized, as in
Figure 1. Most contaminants can be placed in one of
the three groups illustrated. Again, however, the
relative importance of these various pathways in the
natural system is not fully understood for many
compounds; we must continue to support investiga-
tions into the mechanisms of contaminant uptake. In
addition, we must remember that the number of
contaminants for which information is available is
quite limited. Thus, the application of our current
knowledge, gained through laboratory studies, to the
Chesapeake Bay could lead to serious errors in
estimation of impact. For example, we cannot directly
measure the free ion concentration of most cationic
elements, and therefore we cannot predict rates or
amounts of uptake, nor can we predict the potential
availability of all sediment-bound organics from data
available at present.
Organisms play an important role in the bio-
availability of toxic compounds, both through direct
modification of the chemical structure of the toxic
compound, and through modification of the surround-
ing environment (pH, oxygen changes). Unfortu-
nately, except for a relatively few compounds, we do
not know the extent of such modifications.
The overall conclusion must be that the complex-
ity of the interactions between physical, chemical, and
biological factors is extreme. This complexity
hampers our ability to fully understand (and more
important, predict) pollutant transport, availability,
uptake, and impact.
Past efforts, to a large extent, have focused upon
determining the concentration of contaminants within
the various compartments of the system. Although
this research has value, we must move toward careful
study of the processes themselves. It is only through
the latter research that effective management of
resources and protection from contamination can be
achieved. We have identified a number of areas in
which further research is needed. In no particular
order, they are presented below.
(1) We need to determine how partitioning of a
contaminant affects its toxicity, for a number of
important model compounds. For example, is organic
complexation a detoxification mechanism for many
inorganic elements? What about organic compounds?
Which contaminants are still available and toxic when
associated with particulates? How important is the
concentration of contaminants in the surface micro-
layer?
(2) We need to better understand the role of biota
in the transfer of pollutants within the system, in-
cluding transport across the sediment/water interface,
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99
the importance of fecal pellet production, and the fate
of contaminants associated with the biota themselves.
An interesting question is what happens to tissue-
associated toxics when an organism dies. Also, does
the sequestering of pollutants by proteins protect
predators when they consume prey containing these
compounds?
(3) We must determine the importance of
seasonal anoxia on the transport and availability of
important contaminants. Does anoxia in general
increase or decrease contaminant availability? Do
anoxic sediments differ greatly from oxidized sedi-
ments in their ability to act as a pollutant reservoir?
(4) We do not understand the extent to which
sediments control the movement and availability of
contaminants. Sediments act as a sink for many
substances; they can also retain toxics in the estuary
and prolong the exposure. An integrated, wide-scale
study of the flux of carbon and toxic substances across
the sediment/water interface, and the processes that
control this flux, is necessary.
(5) We need to collect better (and more) informa-
tion on the role of communities in the transport and
transformation of contaminants. Too much of our
information is based on laboratory experiments with
single species of organisms under controlled condi-
tions. We need more realistic experimentation using
more sophisticated methods, with microcosms, or
using other approaches. It is data of this sort that will
be most amenable to the modeling efforts that must
eventually form the basis of predictive capabilities.
(6) Finally, we need to continue descriptive
studies of the sources of toxic compounds entering the
Chesapeake Bay, their location within the system, and
their availability to important organisms. This type of
research will always be useful. In particular, we need
information on the sources and reservoirs of key
organic compounds.
ACKNOWLEDGEMENTS
The preparation of this report was supported by a
grant from the Environmental Protection Agency to
the Chesapeake Research Consortium and by the
Academy of Natural Sciences.
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