EPA-660/3-75-025
JUNE 1975
Ecological Research Series
Tidal Flats in Estuarine Water
Quality Analysis
National Environmental Research Center
Office of Research and Development
U.S. Environmental Protection Agency
Corvallis, Oregon 97330
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RESEARCH REPORTING SERIES
Research reports of the Office of Research and Development,
U.S. Environmental Protection Agency, have been grouped into
five series. These five broad categories were established to
facilitate further development and application of environmental
technology. Elimination of traditional grouping was consciously
planned to foster technology transfer and a maximum interface in
related fields. The five series are:
1. Environmental Health Effects Research
2. Environmental Protection Technology
3. Ecological Research
4. Environmental Monitoring
5. Socioeconomic Environmental Studies
This report has been assigned to the ECOLOGICAL RESEARCH STUDIES
series. This series describes research on the effects of pollution
on humans, plant and animal species, and materials. Problems
are assessed for their long- and short-term influences. Investigations
include formation, transport, and pathway studies to determine
the fate of pollutants and their effects. This work provides
the technical basis for setting standards to minimize undesirable
changes in living organisms in the aquatic, terrestrial and atmospheric
environments.
This report has been reviewed by the Office of Research and
Development, EPA, and approved for publication. Approval does
not signify that the contents necessarily reflect the views and
policies of the Environmental Protection Agency, nor does mention
of trade names or commerical products constitute endorsement or
recommendation for use.
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EPA-660/3-75-025
JUNE 1975
TIDAL FLATS IN ESTUARINE WATER QUALITY ANALYSIS
by
David A. Bella
Department of Civil Engineering
Oregon State University
Corvallis, Oregon 97331
Grant No. 16070 DGO
Program Element 1BA025
ROAP Task No. 21A1T/01
Project Officer
Richard Callaway
Pacific Northwest Environmental Research Laboratory (PNERL)
National Environmental Research Center (NERC)
Corvallis, Oregon 97330
National Environmental Research Center
Office of Research and Development
U.S. Environmental Protection Agency
Corvallis, Oregon 97330
For Sale by the National Technical Information Service,
U.S. Department of Commerce, Springfield, VA 22151
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ABSTRACT
This report summarizes the results of a research project entitled "Tidal
Flats and Estuarine Water Quality Analysis." The initial phases of the study
involved mixing processes and tidal hydraulics, however, the study emphasis
shifted to estuarine benthic systems as the importance of these systems
became more apparent. The sulfur cycle was given particular emphasis because:
(1) sulfides, resulting from sulfate reduction within the benthic
systems, can influence the benthic oxygen uptake rate,
(2) free sulfides are highly toxic to a variety of organisms, and
(3) the release of hydrogen sulfide may contribute to a deterioration
of air quality.
The sulfur cycle is of particular importance in tidal estuaries because
of the high sulfate concentrations of saline waters in comparison to fresh
waters. A conceptual model of estuarine benthic systems was developed and a
classification system of estuarine benthic deposits which is based on the
availability of hydrogen acceptors and reactive iron was developed.
Field studies demonstrated that estuarine waters overlying organic rich
tidal flat deposits could contain significant concentrations of free sulfides
even when dissolved oxygen was present. Field studies of benthic oxygen
uptake and benthic sulfide release were conducted. Water quality profiles
within the deposits were also determined. A number of laboratory studies
were conducted to determine the rate of sulfate reduction. Results from
experiments using extracts from benthic deposits and algal mats demonstrated
a close relationship between the rate of sulfate reduction and the sulfate
11
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and soluble organic carbon concentrations. A general systems model of
estuarine benthic systems was developed, however, specific definition of
all processes was not possible without further experimental results. A
variety of activities which could contribute to significant environmental
changes with estuarine benthic systems were identified.
Methods of determining dispersion coefficients from salinity profiles
were examined and an improved method was developed. The build-up of a
pollutant in the vicinity of the outfall during the slack water period of
the tide was studied through a field experiment and mathematical model study.
This report was submitted in fulfillment of Grant 16070 DGO, by Oregon
State University in Corvallis under the sponsorship of the Environmental
Protection Agency. Work was completed as of November, 1973, with minor
revisions made in August, 1974.
111
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CONTENTS
Section Page
I CONCLUSIONS l
II RECOMMENDATIONS 4
III INTRODUCTION 6
General Background ^
General Approach °
Detail and Perspective 8
The Evolution of the Study 10
IV DESCRIPTION OF BENTHIC SYSTEM 12
Introduction 12
General Benthic System 12
Larger Plants and Animals 20
V DESCRIPTION OF SITES USED FOR FIELD STUDIES 26
VI STUDIES OF BENTHAL OXYGEN UPTAKE 32
Initial Laboratory Studies 32
Experimental Design and Procedure for
Field Studies 32
Mathematical Model of Benthal Respirometer
System 34
Calculation of Leakage 37
DO Uptake Rate Calculations 37
DO Uptake Results 38
Sensitivity Study 45
VII FREE SULFIDE IN OVERLYING WATER 47
General 47
Model of Free Sulfide Transfer Through
Aerobic Zone 48
Model of Free Sulfide in Overlying Water 50
Free Sulfide Measurements at Other Sites 57
VIII CONDITIONS WITHIN BENTHIC SYSTEMS 58
Field Measurements 58
A Classification of Estuarine Benthic Systems 59
IX STUDY OF SULFATE REDUCTION USING EXTRACTS 66
General 66
Media Preparation 68
Methods 71
Relationship Between Sulfide, Sulfate and
Organic Carbon 73
Rates of Sulfate Reduction 76
IV
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Section Page
X SULFATE REDUCTION STUDY USING S-35 95
General Approach 95
Sample Preparation 95
Initial Conditions 97
Collection of Sulfide 97
Rates of Sulfate Reduction 101
XI OTHER ESTIMATES OF SULFATE REDUCTION 103
Diffusion of Sulfate 103
Incubation of Benthic Cores 105
Summary of Sulfate Reduction Rates 107
XII BENTHIC SULFIDE RELEASE 109
»
General 109
Modified Benthic Respirometer 109
Results of Sulfide Release Measurements 110
' Profiles of Free Sulfide 114
Summary of Benthic Sulfide Release Experiments 114
XIII MIXING WITHIN DEPSOITS 116
t
General 116
Tidal Mixing 117
Periodic Scour 122
XIV GENERAL BENTHIC DEPOSIT MODEL 125
General 125
Principal Assumptions 126
General Description of Soluble Materials 126
General Equation for Insoluble Materials 128
Biochemical Model I . 128
Biochemical Model II 135
XV ENVIRONMENTAL IMPLICATIONS FOR ESTUARINE BENTHIC SYSTEMS 141
General Implications 141
Changes in Organic Deposition 142
Changes in Inorganic Deposition 142
Construction of Dikes, Jetties, Wharves, etc. 142
Hydrodynamic Changes 143
Tideland Filling 143
Transient Conditions Due to Dredging 145
Long Term Particle Size Change 146
Spoil Disposal 146
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Section
XVI LOWER LEVEL RESOLUTION STUDIES 148
General 148
Advection Errors 148
Estimating Dispersion Coefficients in Estuaries 149
Slack Water Build-up in Estuaries 149
Tidal Measurements 150
XVII REFERENCES 151
XVIII PUBLICATIONS 158
XIX APPENDIX - FINITE DIFFERENCE MODEL 160
VI
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FIGURES
Page
1 AREAS OF FEEDBACK 11
2 GENERAL BENTHIC SYSTEM 13
3 CONCEPTUAL MODEL OF LARGER ANIMALS WITHIN BENTHIC SYSTEM 21
4 SITES LOCATED ON YAQUINA ESTUARY " 28
5 TRANSECT A SITE 1 - SUMMER 1970 29
6 SUMMARY OF CORRECTED BENTHAL OXYGEN UPTAKE RATES AT
SITES 3 AND 4. (1969) 39
7 VARIATION OF BENTHAL UPTAKE RATES AT SITE 4 (1969) 41
8 RANGE OF FREE SULFIDES IN OVERLYING WATER 52
9 DISSOLVED OXYGEN AND FREE SULFIDE PROFILES AT SITE 5 55
10 DISSOLVED OXYGEN AND FREE SULFIDES AT SITE 5 56
11 EXAMPLES OF ESTUARINE BENTHIC TYPES 60
12 QUALITATIVE DESCRIPTION OF ESTUARINE BENTHIC SYSTEM RESPONSE
TO CONSTANT DEPOSITION CONDITIONS 62
13 EXPERIMENTAL RESULTS OF CULTURE 1 COMPARED WITH EQUATIONS 23,
24 AND 25 81
14 EXPERIMENTAL RESULTS OF CULTURE 2 COMPARED WITH EQUATIONS 23,
24 AND 25 82
15 EXPERIMENTAL RESULTS OF CULTURE 3 COMPARED WITH EQUATIONS 23,
24 AND 25 83
16 EXPERIMENTAL RESULTS OF CULTURE 4 COMPARED WITH EQUATIONS 23,
24 AND 25 84
17 EXPERIMENTAL RESULTS OF CULTURE 5 COMPARED WITH EQUATIONS 23,
24 AND 25 85
18 EXPERIMENTAL RESULTS OF CULTURE 6 COMPARED WITH EQUATIONS 23,
24 AND 25 86
19 EXPERIMENTAL RESULTS OF CULTURE 7 COMPARED WITH EQUATIONS 23,
24 AND 25 87
20 EXPERIMENTAL RESULTS OF CULTURE 8 COMPARED WITH EQUATIONS 23,
24 AND 25 88
vii
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FIGURES (cont)
Page
21 EXPERIMENTAL RESULTS OF CULTURE 9 COMPARED WITH EQUATIONS 23,
24 AND 25 89
22 EXPERIMENTAL RESULTS OF CULTURE 10 COMPARED WITH EQUATIONS 23,
24 AND 25 90
23 EXPERIMENTAL RESULTS OF CULTURE 11 COMPARED WITH EQUATIONS 23,
24 AND 25 91
24 EXPERIMENTAL RESULTS OF CULTURE 12 COMPARED WITH EQUATIONS .23,
24 AND 25 92
25 EXPERIMENTAL RESULTS OF CULTURE 13 COMPARED WITH EQUATIONS 23,
24 AND 25 93
26 EXPERIMENTAL RESULTS OF CULTURE 14 COMPARED WITH EQUATIONS 23,
24 AND 25 94
27 RELATIONSHIP OF SULFATE REDUCTION RATE TO SULFATE DIFFUSION
ASSUMING STEADY STATE 104
28 BENTHIC SULFIDE RELEASE WITHIN RESPIROMETER - SITE 5 - 8/3/71 111
29 BENTHIC SULFIDE RELEASE WITHIN RESPIROMETER - SITE 5 - 8/9/71 111
30 BENTHIC SULFIDE RELEASE WITHIN RESPIROMETER - SITE 5 - 8/20/71 112
31 HYDRAULIC SURFACE DURING EXPERIMENTS AT SITE 1 119
32 CONDITIONS AT SITE 1 DURING DRAINAGE EXPERIMENT 119
33 FREE WATER ELEVATIONS DURING EXPERIMENTS AT SITE 1 120
34 RELATIONSHIP BETWEEN TOTAL SULFIDE (ACID SOLUBLE) AND REDOX
POTENTIAL WITHIN DEPOSITS AT SITES 1, 2, 4, AND 5 123
35 TEMPORAL CHANGE OF TOTAL SULFIDE (ACID SOLUBLE) PROFILES
AT SITE 4 123
36 BIOCHEMICAL MODEL I FOR ESTUARINE BENTHIC SYSTEM 129
37 BIOCHEMICAL MODEL II FOR ESTUARINE BENTHIC SYSTEM 136
viii
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TABLES
1 PARTICLE SIZE FOR SAMPLING SITES 30
2 BACTERIAL COUNTS PER GRAM OF WET SEDIMENT
(SUMMER) 1970 31
3 DEFINITION OF TERMS FOR RESPIROMETER MODEL 36
4 SUMMARY OF BENTHAL OXYGEN UPTAKE RATE 44
5 ESTIMATED PARAMETERS APPLICABLE TO SITE 5
DURING LATE SUMMER AND EARLY FALL PERIOD 53
6 CLASSIFICATION OF ESTUARINE BENTHIC SYSTEMS 61
7 SUMMARY OF SOLUBLE ORGANIC EXTRACT PROCEDURES 70
8 EXPERIMENTAL YIELD RATIOS AND MAXIMUM RATES OF
SULFIDE PRODUCTION FOR EACH CULTURE 74
9 SUMMARY OF YIELD RATIOS 75
10 SUMMARY OF S-35 RECOVERY 102
11 SULFATE REDUCTION IN CORES AT ROOM TEMPERATURE (25°C) 106
12 COMPARISON OF THE RATES OF SULFIDE PRODUCTION
MEASURED IN THIS STUDY WITH THOSE OF
PREVIOUS INVESTIGATORS 108
13 ESTIMATES OF SULFIDE RELEASE RATES AND OXYGEN
UPTAKE RATES AS MEASURED IN RESPIROMETER
EXPERIMENTS AT SITE 5 113
14 POTENTIAL BENTHIC SULFIDE RELEASE RATES 115
15 PERMEABILITY AND VOID RATIO AT SITE 1 121
16 DEFINITION OF TERMS FOR BIOCHEMICAL MODEL I
OF ESTUARINE BENTHIC SYSTEM 130
17 DEPENDENCE OF REACTIONS OF BIOCHEMICAL MODEL I
ON INTERNAL BENTHIC CONDITIONS 133
18 DEFINITION OF TERMS FOR BIOCHEMICAL MODEL II
OF ESTUARINE BENTHIC SYSTEM 137
19 DEPENDENCE OF REACTIONS OF LOWER RESOLUTION MODEL
(MODEL II) ON BENTHIC CONDITIONS 138
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ACKNOWLEDGMENTS
The writer wishes to acknowledge with appreciation the support of this
grant by the Environmental Protection Agency. The assistance of Mr. Richard
Callaway who served as EPA Program Officer is particularly acknowledged.
The contributions of research collaborators whose names appear as authors
in the publication list of section XVIII were essential to the success of
this project. I express my gratitude for their dedication, and enthusiasm
and for the capable work which they performed. Special appreciation is given
to William J. Grenney, Alan E. Ramm, Paul Peterson, Peter Wong and Ralph C.
Olsen.
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SECTION I
CONCLUSIONS
1. Benthic systems are significant regions of estuarine systems and should
not be merely treated as boundary conditions to the overlying waters. The
processes occurring within estuarine benthic systems and, in particular,
the sulfur cycle, are of major importance with regard to sound environ-
mental management of estuaries. A description of estuarine benthic
systems is provided in sections IV and XII.
2. Free sulfide (produced within deposits) can be found at concentrations
of approximately 1 mg/L in oxygenated tidal flat waters overlying high
organic deposits. Free sulfide concentrations of 50-100 mg/L and higher
can be found within the interstitial waters of high organic benthic
deposits within several centimeters of the deposit surface. Such con-
centrations can be toxic to a wide variety of organisms.
3. The presence of high organics within the deposits, available sulfates,
low concentrations of available iron, poor drainage, low water velocities
and the absence of significant scour are conditions which lead to the
build-up of free sulfides within estuarine benthic systems. A variety
of human actions can contribute toward these conditions. (See Section XVI)
4. In addition to the conditions described in conclusion 3, low dissolved
oxygen concentrations and shallow water depths are conditions which
favor higher free sulfide concentrations with the overlying waters.
5. Rates of sulfide production ranging from approximately 10 mg(S)/l-day
to 70 mg(S)/l-day were measured by a variety of methods.
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6. Rates of sulfate reduction within laboratory experiments using algal
extracts were found to be primarily dependent on the sulfate and
organic carbon concentrations. The rate of sulfide production in mg(S)/
1-day (4^-) is given by the following equation
. _ 77
dt "
in which L is the sulfate concentration and C is the soluble organic
carbon concentration. The rate of soluble organic carbon utilization is
given by
f - -*•»<&
The rate of sulfate removal is given by
dt
7. In situ benthic oxygen uptake rates within tidal flat areas without burrow
2 2
holes ranged from approximately 1.4 gm/m -day to 2.1 gms/m -day with the
higher rates associated with higher water velocities. Benthic oxygen uptake
2
rates up to approximately 8-9 gms/m -day were measured in regions with
large numbers of burrow holes. Within such regions, correction of respiro-
meter leakage had to be made.
8. Estimates of benthic sulfide release in tidal flat regions of high organic
content were limited and highly variable. These limited results, however,
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2
suggest that benthic sulfide release rates of approximately 1 gm(S)/m -day
and higher are not unreasonable in tidal flat regions displaying the
characteristics described in conclusion 3. It is not now unreasonable to
suspect sulfide release from estuarine benthic systems, particularly in
tidal flat regions, as a major contributor to atmospheric sulfur.
9. The errors associated with common, finite difference models of advection
can be classified into three general categories: (1) oscillation errors;
(2) skewness errors; and (3) dispersive errors. The use of the upstream
difference method permits correction of these errors.
10. The use of the steady state assumption for measuring longitudinal dis-
persion coefficients from salinity profiles can lead to a false rela-
tionship between fresh water flow and the magnitude of the dispersion
coefficients. An improved method of estimating dispersion coefficients
from salinity profiles was developed.
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SECTION II
RECOMMENDATIONS
1. The magnitude and extent of benthic sulfur release to the atmosphere
should be examined on a large scale. It is possible that increases in
this release due to a variety of human activities can result in major
inputs of atmospheric sulfur.
2. The extent of estuarine benthic deposits containing significant amounts
of free sulfide and the influence of human activity on this extent should
be examined.
3. A more comprehensive (low level resolution) understanding of the system
properties of estuarine ecosystems must be pursued. Only with such an
improved understanding can the significance of the more common environ-
mental concerns (e.g. low dissolved oxygen, higher free sulfides, stream
flow regulation) be appreciated with regard to the functioning of
estuaries within the biosphere. Such studies must place a greater
emphasis on the long term consequences of human activities.
4. A more qualitative description of the formation of combined sulfides
within benthic deposits should be developed. Such a description would
enable a refinement of the systems model presented in section XV. Such
a study should proceed at two levels of resolution. The finer resolution
would likely deal with the reactions of iron within benthic systems. The
lower resolution model should deal with a measurement of the "chemical
sulfide demand" (CSD) of a deposit. The CSD would be a measure of the
sediment capacity to tie up sulfides in insoluble forms.
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5. Studies are needed to quantitatively define the rate of sulfate reduc-
tion particularly in the top regions of the anaerobic portions of deposits
and immediately below algal mats. Attempts should also be made to
develop a measure of the "biochemical sulfate demand" (BSD) of deposited
material. The BSD might be a useful concept to incorporate into the
lower resolution model of section XV. The ratio, BSD/CSD, would be an
indicator of potential sulfide release for a given deposit (generally
within the top several centimeters).
6. The concentrations of free sulfides within waters overlying deposits and
the amount of hydrogen sulfide released to the atmosphere will depend,
in part, on the rate of oxidation of free sulfide. Additional study is
needed to better estimate the rate of sulfide oxidation within estuarine
waters, particularly at low sulfide concentrations. The age of the water
sample should be an important consideration as it appears that recently
collected estuarine waters display a more rapid oxidation rate than
waters stored for a period of time after collection.
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SECTION III
INTRODUCTION
GENERAL BACKGROUND
This is the final report on a three year study entitled "Tidal Flats in
Estuarine Water Quality Analysis." This study was supported by the Environ-
mental Protection Agency through Research Grant No. 16020DGO.
GENERAL APPROACH
The technical material presented within this report is generally arranged
in logical order rather than in the chronological order that the work was
performed in. The writer feels that it would be of value, however, to briefly
discuss the general approach used during the study and briefly discuss how this
approach often changed the direction of the research.
The general objective of this research is to learn more about estuarine
systems and in particular the tidal flat systems, with particular emphasis
given to how man's activities can disrupt these systems to the eventual dis-
advantage of man. The numbers of components and relationships occurring
within these systems are so numerous and complex that a complete understanding
is essentially impossible. Research effort must be thus directed to study those
areas which appear to be of greatest importance. The difficulty is that our
knowledge of what we determine to be most important changes as we learn more
about the systems. This knowledge (obtained from all sources) hopefully
indicates new areas which may be of great importance. A research project must
be flexible enough to respond to new information yet stable enough to lead to
sound results. Lack of flexibility often results in the pursuit of those items
which are already known quite well, while lack of stability can lead to nothing.
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In the reported research project, an approach was used in which emphasis
was alternately given to mathematical models and experimental results. That
is, the experimental results improved the mathematical models while the models
in turn suggested where further experimental results might be most profitable.
As an example, the oxygen demands of tidal flat areas, particularly the benthal
deposits, was recognized from the start as an important consideration. First,
laboratory studies were made to study the benthal uptake. Next, a simple
mathematical model was developed. From this model, an in situ benthal respiro-
meter was designed, built and run. The results from the in situ benthal oxygen
uptake rate studies appeared to be effected by leakage from the respirometer.
A mathematical model of the respirometer system was developed and from that
model, correction for the leakage was made. The corrected benthal oxygen uptake
rates were then studied. The mathematical model results indicated that the
experimental results could be best explained if a substantial portion of the
measured oxygen uptake rate was due either to the dissolved oxygen, DO, diffusing
into the deposits or due to the release of a material which was oxidized rather
quickly (half life of several hours or less). Both of these processes suggested
that a portion of the benthal uptake (not including benthal plant respiration)
might be due to a quickly oxidizing material; the material being oxidized either
within the aerobic region of the deposit or within the overlying water. The
literature suggested that free sulfides might be such materials. Because free
sulfides (particularly hydrogen sulfide) are quite toxic, their presence within
the water might often be of greater significance than the low DO values which
result, in part, from the oxidation of the free sulfides. It was generally
felt at the time, however, that the rapid rate at which free sulfides are
oxidized would prevent its presence in waters containing measurable DO. Thus,
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it was felt, that the free sulfides which were released from the anaerobic
regions of the deposit would normally be completely (or near completely)
oxidized within the aerobic zone of the deposits. The literature also
reflected this notion. A mathematical model of the aerobic zone of the deposits
was developed. This model included the downward diffusion of DO, the upward
diffusion of free sulfides and the reaction between the two. The model results
indicated that under certain conditions, the free sulfide concentrations within
overlying waters could be significant. Experimental methods were then developed
and significant concentrations of free sulfides were measured in certain areas.
Field studies and an exhaustive literature review then led to a qualitative
description of what appears at this time to be the important processes leading
to both the oxygen uptake and the release of free sulfides. The description of
the benthic system exposed important processes for which very little experimental
data was available and expansion of the mathematical models without further
experimental results would have been unjustified and misleading. Thus the most
profitable results during the final phase of the project centered around field
and laboratory studies. A general benthic system model was developed, however,
it was determined that specific definition of a number of processes required
further experimental efforts. The author was reluctant to speculate on certain
specific descriptions within the model because such speculation might be too
easily accepted without further experimental results.
DETAIL AND PERSPECTIVE
The general research approach followed in this study not only involved a
feedback between mathematical model results and experimental results, but
also involved a feedback between different views or perspectives of estuarine
systems as described below.
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The real world appears to be organized into an integrated hiarchy of
organizational structures. On an extremely small scale, atoms are organized
to form molecules. On a large scale, the planets and sun are organized to form
the solar system. Large numbers of intermediate structures, some obvious and
some not, are of course present. The definitions of different structures and
the disagreements of these definitions will not be pursued. Rather, the point
to be made is that a given structure or entity is both made up of components and
is also a component of a higher structure.
In order to understand the natural world, man has found it necessary to
group what appear to be natural structures into larger groupings. As an example
of a functional grouping, individual organisms with similar functions have
been grouped into trophic levels. This grouping has enabled man to study the
relationships between large groups of organisms. Such a grouping thus
enables one to gain perspective, yet. because of this larger grouping, one loses
detail.
The same real world systems may be studied at a fine level of resolution;
(different degrees of grouping). A fine resolution leads to a gain in detail
with a sacrifice of perspective while a low resolution leads to a gain in
perspective with a sacrifice in detail (precision}.
Different investigators (and different professions} often look at the
same real world system from different levels of resolution. Such a varied
resolution approach often leads to information concerning detail and
perspective.*
This project, from the start, has studied the tidal flat system (a loosely
defined system having, however, some unique characteristics) from two general
* Author's Note: Some information may require both detail and perspective
simultaneously. Such information may well be essentially unattainable. At
this time, however, I will not pursue this rather philosophical question of
"ecological uncertainty."
9
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views (i.e. at two general levels of organization.] Related component parts
which make up the tidal flat system were studied. In addition, the larger
estuarine system, of which the tidal flats are components, was also studied.
Feedback occurred between the results gained from the different views;
that is, results gained at one level of organization influenced the direction
of work done at the other level of organization. As an example, the disper-
sion of saline water within an estuary has been a common subject of study.
In the early phases of this project, the examination of estuarine dispersion
coefficients was pursued. The intrusion of sea water results in conditions within
estuaries which are uniquely different from most fresh water streams. Most
f
mathematical models of estuaries, however, are very similar to those used in
fresh water streams with the most common difference being a temporarily
varying hydraulic regime. Saline water contains sulfate concentrations many
times higher than found in fresh waters (sea water contains 2655 mg/L of sulfates]
The significance of these higher sulfate concentrations is not apparent until
one increases the level of resolution of his view. These higher sulfate concen-
trations have significant effects upon the benthic systems which, in turn,
influence the water quality. One does not appreciate these effects until one
views the estuarine bottoms as systems of interacting components and processes
rather than boundary conditions to the larger "estuarine systems." Thus, a
significant result of the salinity dispersion examined at one level of resolu-
tion cannot be appreciated until a finer level of resolution is examined.
THE EVOLUTION OF THE STUDY
The general approach followed in this study can be most simply described
as an "evolutionary" process involving feedback between the four general
areas illustrated in Fig. 1.
10
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Math models
low level
resolution
high level
resolution
Experiment
FIG. 1 - AREAS OF FEEDBACK
The study, evolved with the direction of this evolution loosely guided,
through the feedbacks, by the three following questions:
1. Is it important?
2. Has it been done or is it being done?
3. Can you do it?
This evolutionary process serves to explain how a project which initially
emphasized dispersion coefficients and dissolved oxygen (DO) balance (two
popular subjects) shifted to a systems study of the sulfur cycle within estuarine
benthic systems (a neglected subject of potentially significant importance).
11
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SECTION IV
DESCRIPTION OF BENTHIC SYSTEM
INTRODUCTION
The following section will provide a general description of estuarine
benthic systems. This description has been developed during the course of
this research project through extensive literature searches and field and
laboratory investigations (1)(2)(3).
GENERAL BENTHIC SYSTEM
Any approach to an understanding of estuarine benthic systems must
involve the complex interactions of the biological, chemical, physical and
hydraulic processes. The basis for understanding such systems involves the
development of conceptual models. An investigator seeks to develop a
simplified model capable of satisfactorily describing certain important
aspects of an actual system whose complete complexity is beyond the capacity
of the investigator to perceive. In developing a model of a system, one
must trade between detail and perspective. Too great a detail makes it
difficult to define the relationships between the many components of the
model. Sacrifice of detail leads to a better perspective yet eliminates
useful information from the model. The level of resolution of the conceptual
model presented herein (Fig. 2) was selected to explain certain concepts
which are of importance to environmental quality. Omission of certain
processes, reactions or other influences should not imply that these omis-
sions are unimportant. Periodic reference to Fig. 2 will help to clarify
the following discussion.
12
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AIR
i
\ EXTERNAL
| SOURCES
\l
INSOLUBLE
INORGANICS
G
\ '
V~^ ^
k i
V"
, t
^-/CHEMOAUTOTROPHICTS
Jv. BACTERIA J
f t t
''/A'-
/ ^ ^ E
^fi PHYTOPLANKTON ) \ Ł
/ ^X \ ^
/ W' HETEROTROPHIC V * no,-...,
-X.--7- "*V BACTERIA J*-* ORGAN
" x ' S' '
XTERNAL
OURCES
/
cs WATER
ANAEROBIC
DEPOSIT
LINES f?EPftŁSENT PHYSICAL TRANSFER PROCESS
CHEMICAL REACTION NOTED BY •
* DENOTES AVAILABLE Fa (also Zn, Sn, Cd, Hg mil CuJ
FIG. 2 - GENERAL BENTHIC SYSTEM
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Inorganics and organics are deposited to estuarine benthic systems.
Inorganics, including sands, silts'and clays, are introduced into estuaries
from the ocean, upstream rivers and localized runoff. Organics originate
from sources outside the estuary, as well as from primary production within
the estuary. The system which results from such deposition is illustrated
in Fig. 2.
Decomposition of deposited organics is most often largely roicrobial
with bacteria predominating. (The influence of larger detrital and deposit
feeders is discussed in the following subsection.) The type of bacterial
decomposition occurring at any location is determined principally by the avail-
ability of hydrogen acceptors. When available, dissolved oxygen, DO, is used
as the hydrogen acceptor. In its absence, oxidized forms of sulfur, princi-
pally sulfate, become the principal hydrogen acceptors. Because nitrate
concentrations are nearly always far less than sulfate concentrations within
estuarine systems, nitrate reduction, which will occur before sulfate reduc-
tion, will not be discussed. The absence of suitable concentrations of both
oxygen and oxidized sulfur necessitates the use of endogenous hydrogen
acceptors. (For a discussion of endogenous and exogenous hydrogen acceptors
see Schroeder and Busch(4)).
The availability of exogenous hydrogen acceptors (DO and sulfates}
depends upon the mixing and advection within the deposits. Vertical mixing,
and thus the transport of exogenous hydrogen acceptors, is increased by
greater water velocities, high concentrations of dissolved oxygen and sulfate
within the overlaying water, high permeability of the deposits, and a high
rate of turnover by the larger organisms. This latter factor is, in part,
dependent on the interstitial water quality. Advection through the deposits
14
-------
depends on the permeability of the deposits and the direction and magnitude
of the hydraulic gradient.
The availability of hydrogen acceptors and organics determines the
nature and extent of bacterial decomposition which, in turn, largely deter-
mines the quality of the interstitial and interfacial waters. The avail-
ability of oxygen is an important factor affecting estuarine benthic systems.
Oxygen is added to the overlying water by reaeration, photosynthesis and
transport due to water movement. The interstitial DO concentration of
deposits is determined by a balance between DO transport from above (by
mixing and advection) and DO utilization (both chemical and biological)
within the deposits.
If the input of organics to deposits exceeds the transfer of DO,
aerobic decomposition will not be sufficient to decompose all of the
organics. Sulfate reduction will then proceed below the aerobic region.
The reduction of sulfates by heterotrophic sulfate reducing bacteria which
utilize the sulfate ion as a terminal hydrogen acceptor (5) results in the
release of hydrogen sulfide which is found in solution as part of the pH
dependent system
">S= (1)
In the present discussion, all components of the above relationship will be
defined as "free sulfide." At a pH of 6.5-7.0, the free sulfide is approxi-
mately evenly divided between H-S and HS with S~ being negligible (6). free
sulfides are also produced during anaerobic putrefi cation of sulfur contain-
ing amino acids, but this process is felt to be of lesser importance in the
marine envi ronment (7,8).
15
-------
Free sulfides form insoluble compounds with heavy metals, particularly
iron. Free sulfide quickly reacts with available iron within the deposits to
form ferrous sulfide, FeS, which gives benthic deposits their characteristic
black color (9). The input of this iron into the deposits results primarily
from the deposition of insoluble inorganics, which contain ferric oxides and
other insoluble forms of iron. Not all of this iron, however, is available to
react with the sulfides. Some additional reactive iron originates from the
decomposed organics. This latter source is usually less and thus the supply
of available iron within the deposits is often largely dependent on the nature
and extent of inorganic deposition (.10)- Other heavy metals such as zinc, tin,
cadmium, lead, copper and mercury all have solubility products significantly
below that of ferrous sulfide and thus, the presence of ferrous sulfide
indicates that ionic solutions of these metals within the interstitial
waters are not likely to be significant.
Free sulfide concentrations within benthic deposits will remain at low
levels (generally below 1 mg/1) when available iron is present. If avail-
able iron is sufficiently depleted, free sulfides within the anaerobic
regions of deposits will increase until their production at a given location
is balanced by the advective and diffusive transport, out of that location and
by the loss caused by reaction with any remaining available iron. Measured
free sulfide concentrations up to approximately 130 Jng/1 were found within
interstitial waters of tidal flat deposits though some loss may have occurred
during the analysis. Theoretical investigations indicate that Tnaximum concen-
trations might be several times higher if all available iron is depleted.
16
-------
If the aerobic layer of the sediment is thin enough to allow light
to penetrate to the anaerobic zone, populations of photosynthetic purple
and green sulfur bacteria may develop, utilizing the free sulfides as hydro-
gen donors, and producing free sulfur as a by-product. This may occur
below an algal mat and is possibly due to the lower compensation point for
bacterial photoreduction, and to the ability of the photosynthetic bacteria
to utilize longer wavelengths of light than can the algae (8,11).
If the rate of free sulfide production exceeds the rate at which it can
be converted to nondiffusible forms, such as ferrous sulfide or insoluble
free sulfur, the sulfide may diffuse upward into the aerobic regions of the
sediment or into the water column. Here it will be oxidized to sulfite,
thiosulfate, sulfate, or free sulfur (7,12,13).
The chemical reaction of free sulfides in aqueous solutions has been
studied by many investigators (7,12,13,14,15,16). Half lives of free sulfide,
in aqueous solutions, ranging from 15 minutes to 70 hours have been reported.
Several studies have described the oxidation of free sulfides to occur via
second order kinetics (6,7,12); however, such a description is a simplifi-
cation of an extremely complex chemical temperature, pH, and initial oxygen
and sulfide concentrations are all factors affecting the rate of oxidation
(7,12). The oxidation of free sulfide is catalyzed by the presence of
metallic ions, such as of Ni, Mn, Fe, Ca, and Mg, and is accelerated by some
organic substances such as formaldehyde, phenols, and urea. Thus, the oxida-
tion of free sulfides in estuarine and marine water may be much more rapid
than in distilled water due to the presence of catalysts. Within oxygenated
sea water the half life of sulfide has been reported to vary from 10 minutes
17
-------
to several hours (7,13,17). Studies indicate that estuarine waters stored
for a period of time after collection will display slower oxidation rates
of free sulfides than freshly collected waters (18). Since HS predominates
at the pH of sea water, it has been proposed that the oxidation proceeds by
the following reaction (19).
2HS" + 202^S2°3 + H2° ^2)
Following the above chemical oxidation, the thiosulfate ion is more slowly
oxidized to sulfate, probably with the intermediate production of other
oxidized forms. Sulfur oxidizing bacteria of the genus Thiobacillus
appear to be important in this final oxidation step (20,21).
If the benthic deposits are overturned or flushed with oxygenated water,
ferrous sulfide will be rapidly oxidized. Overturned sediments will normally
return to anaerobic conditions. A portion of the ferrous sulfide iron will
be returned to the sediment as available iron which can further react with
free sulfide to form more ferrous sulfide. Thus overturning and flushing
of sediment with oxygenated waters results in. a recycling of available iron.
When oxidation of sulfides occurs, either inorganically or by sulfur oxi-
dizing bacteria, some of the free sulfide and ferrous sulfide is oxidized to
elemental sulfur. In an anaerobic environment elemental sulfur then slowly
reacts with FeS to form pyrite. This latter reaction occurs on a time scale
of years and may lead to a more permanent depletion of available iron.
Free sulfide can be released to the overlying water even if these
waters do contain DO. Experimental and theoretical results presented in
a latter section of this report demonstrate that free sulfide concentrations
of approximately 1 mg/1 can persist in shallow tidal flat open waters as a
result of benthic sulfide release even when the DO of these waters is in
18
-------
the 4-6 mg/1 range. Hydrogen sulfide may also be released to the atmos-
phere particularly when the water depths are shallow or the benthic systems
are exposed.
High concentrations of free sulfides within the deposits and the
release of free sulfides to the overlying water and atmosphere can be
environmentally significant for a number of reasons; among these are the
following.
1. The release of free sulfides can increase the benthic oxygen
demand rate and thus lead to a decline in the aerobic zone of the
deposit and a lowering of the DO concentrations within the over-
lying waters, particularly with the interfacial regions. Though
these interfacial regions constitute a very small fraction of the
estuarine water mass, they are of high ecological importance.
2. Free sulfides, particularly hydrogen sulfide, are toxic at low
concentrations to fish, crustaceans, polychetes and a variety of
benthic microinvertebrates (8,16,20,22,23,24). Actual toxic
concentrations may be considerably lower than some reported in the
literature because of initial sulfide concentrations within batch
tests are often reported. Average concentrations throughout the test
period may be considerably lower due to chemical oxidation. In
tests which maintained nearly constant conditions, hydrogen sulfide
concentrations below 0.075 mg/1 (pH 7.6-8.0) were found to be
significantly harmful to rainbow trout, sucker, and walleye,
particularly to the eggs and fry of these fish (22).
3. The release of hydrogen sulfide to the atmosphere can cause an
air pollution problem. Not only does hydrogen sulfide have an
19
-------
undesirable odor but it is also toxic. Moreover, the release
of hydrogen sulfide from tidal flat areas may be a significant
input of atmospheric sulfur (25)(26)(27).
If the solubilization of organics at a given depth exceeds the downward
transport of DO and sulfates to that depth, decomposition of organics below
this depth must proceed through the use of endogenous hydrogen acceptors (not
shown in Fig. 2). Increased accumulation of organics can be expected, particu-
larly if the absence of available iron results in free sulfide concentrations
sufficiently high to inhibit endogenous decomposition. Methane fermentation
will occur below the region of sulfate reduction if conditions are suitable.
Formation of gases (principally methane) within these regions nay lead to
the disruption of the bottom, and the release of free sulfides to the over-
lying water.
LARGER PLANTS AND ANIMALS
The animal component of estuarine benthic ecosystems can be generally
divided into infauna (animals that live within the sediments) and epifauna
(animals that live on the sediment surface or just above it). Some infauna
make ephemeral pockets in the sediment which are filled as the animal moves
on; others make more permanent burrows and bring overlying water into the
sediment. Much of the infauna is microscopic, living among the sediment
particles.
Although the separation of benthic animals into infauna and epifauna
can be useful, the following discussion will rely more on feeding behaviors(2)•
Benthic animals are divided into three feeding types: selective particle
feeders, deposit feeders and filter feeders, (see Fig. 3)
20
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AIR
EXTRACELLULAR
ENZYMES
WATER
DEPOSIT
INSOLUBLE ORGANICS WITHIN WATER INCLUDE PHYTOPLANKTON, ZOOPLANKTON, DETRITUS, ETC.
FIG. 3 - CONCEPTUAL MODEL OF LARGER ANIMALS WITHIN BENTHIC SYSTEM
-------
Selective particle feeders may be herbivores, predators, or scavengers.
They may feed on whole organisms which they actively capture, or they may
feed on fragments of plants or animals. Crabs, some worms, most fishes,
and other more mobile species fall into this category. The food contains
little inorganic material and is generally broken down by mechanical processes
and then by chemical processes. The residues, inorganic materials, undigest-
ible organics, and resistant bacteria, are combined with mucous and coated
to form distinctive fecal pellets. Fecal pellets generally settle to the
bottom and may make up from 30% to 50% of the sediment, and in extreme cases,
where quiet bottom waters occur, they may account for up to 100% of the
sediment (28)(29). The fecal pellets of carnivores are generally loose pellets,
those of herbivores harder, and those of deposit feeders hardest. Many of
the pellets have characteristic shapes, size, and sculpturing, and are of
taxonomic importance. Some are quite fragile while others may persist for
.more than 100 years.
There are two general types of deposit feeders. Some move through the
sediment and take in the sediment as they go, digesting what they can of the
organic material and discarding as feces the undigestible organics and the
inorganic residues. These animals are mostly worms in estuaries and are not
generally in direct contact with the waters which overly the sediments. Other
deposit feeders bury themselves in the sediment but have siphons or other
extensions through which they "suck up" detritus that has recently fallen to
the sediment surface. Certain clams and worms feed in this way. Again these
species feed unselectively on the available food and are usually unable to
sort food very efficiently. Food of deposit feeders is broken down chemically,
and in some cases mechanically, and the residues are formed into fecal pellets
which contain much greater quantities of inorganic materials than do feces of
other feeding types.
22
-------
Filter feeders sieve water and remove particulate material. Mussels,
some clams, and some worms are examples of this category. Most filter feeders
use cilia to create currents of water over a mucous network which entangles
particles. These are called ciliary mucous feeders. Mussels are good
examples. Other species, particularly tube dwellers, may force water through
the tube with peristaltic body movements. Urechis caupo, the sausage worm,
is an example. The particle laden mucous is then taken into the digestive
system. Some clams "sort" the particles before they are taken into the
digestive system and discard the unusable sizes in mucous masses as pseudofeces,
The food that passes through the digestive tract does not usually require
mechanical maceration and is digested chemically. The feces of filter feeders
are primarily organic.
Not all animals fit neatly into these feeding categories. Some deposit
feeders may be somewhat selective, and some selective feeders may be quite
nonselective if food is scarce. Some animals like starfish that utilize extra-
corporeal digestion are true predators but do not otherwise fit neatly into
the selective feeding type.
Animals tend to break down larger particles through maceration and diges-
tion. The formation of fecal pellets places these particles on the bottom
rather than returning them to the water to increase the turbidity. The fecal
pellets are finally degraded by bacteria, but may pass through several deposit
feeders before final mineralization.
Certain animals, particularly ciliary-mucous feeders have a marked effect
on the turbidity of the overlying water (30). These organisms remove particu-
late matter from the water and compact much of it in the form of pseudofeces
which are larger than the suspended particles and therefore sink more rapidly
23
-------
to the bottom. This reduction in turbidity permits more light to reach the
benthic algae, enhancing the photosynthetic process and increasing the daylight
DO. At the same time, the removal of CO tends to raise the pH during the
daylight hours. Soluble organic wastes of all feeding types are discarded
into the water or interstitial water depending on the habitat of the particular
species.
Some benthic plants also tend to stabilize the benthic environment. Algal
mats can serve to reduce erosion of benthic deposits. When such mats become
extensive, they can significantly contribute to the formation of free sulfide
in the deposits below. Other plants, such as eelgrass, send roots into the
sediment, and many burrowing animals construct tubes that also reduce erosion.
These roots and tubes also provide shelter for infaunal species and may con-
tribute to their food as well.
Animals also influence vertical mixing within the sediment. There is a
great deal of mechanical mixing as burrowing species construct their tubes
or move through the sediment. Fecal pellets of infaunal species are frequently
brought to the surface and deposited there. Burrows may extend more than a
meter into the sediment and the constant reworking of sediment insures a
relative homogeneity of the sediment to that depth. Burrows also provide a
route for oxygen to reach into the sediment, and although sediments may be
anaerobic a few millimeters beneath the surface, there will usually be an
aerobic region immediately surrounding each burrow if the overlying water is
not devoid of oxygen. Conversely, burrows serve as a route through which
waste materials such as fecal pellets and dissolved organics can move out of
the sediment. Burrowing activities also serve to release inorganic nutrients
to the surface water where they may be utilized by photosynthetic plants.
Wave action over a beach filled with tubes may cause a pumping action through.
24
-------
the tubes, increasing aeration and, possibly, erosion of the wall of the tubes.
Such mixing can also contribute to the oxidation of combined sulfide (sudi
as FeS) and the "recycling" of iron to combine with produced sulfide and thus
prevent high levels of free sulfide.
The major role of green plants is to convert solar energy into a form
that can be used by plants and animals. Through the photosynthetic process
inorganic substances are converted into high-energy organic compounds. This
primary production is the ultimate source of food for all organisms. Phyto-
plankton, benthic algae and eelgrass are all important primary producers within
estuarine ecosystems. Organic materials produced by the plant components are
transferred through herbivores and several levels of carnivores to the sedi-
ments. Feedback occurs frequently so that the web concept is more descriptive
than the food chain. In the sediment, these materials are mineralized to
their inorganic end-products and then may again enter the cycle.
The role of particles in the marine water often is not appreciated. Heavy
metals may adsorb to these particles and if the particles remain in suspension
they may be carried far to sea before they settle. These metal-laden particles
may, however, be pelletized by various animals and deposited. The role of
animals in removing such particles may be very important.
Many of the effects that have been discussed deal with the transportation
of materials from the water to the sediment, but there is transport in the
other direction as well. Benthic species almost always have pelagic larvae.
Essentially all of these larvae must feed and develop within open water regions,
returning to the sediment for later life stages. Pelagic stages thus insure
wide dispersal of these species. Propagation of benthic animals thus depends
on the ability of pelagic life stages to leave the sediment.
25
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SECTION V
DESCRIPTION OF SITES USED FOR FIELD STUDIES
Five sites were used during various stages of this study for the
field studies. The four sites located within Taquina Estuary are shown
in Fig. 4. Site 1 was located on the south side of the estuary immediately
to the east of the Oregon State University Marine Science Center. This
site lay within approximately one mile of the estuary mouth, and was strongly
influenced by marine water. High salinities (33-35 parts per thousand), low
temperatures (45-50°F), and high, tidal current velocities (0.6 - 0.8 feet
per second) were characteristic here. This area appeared to be fairly
remote from, any major source of industrial pollution, and no excessive
domestic contamination was evident. The sediments here were heavily
colonized between +4 and +6 feet above mean low low water (MLLW) by large
populations of the mud shrimps Callinassa californiensis and Upogebia
pugettensis. They were very active in burrowing and mixing of the sedi-
ments at this elevation. There was a distinct lack of attached vegetation
in this range, but summer growth of the benthic alga Enteromorpha and of
Zostera was extensive below +4 feet MLLW. At approximately +7 feet MLLW
the sediment was covered by a thin, but very firm mat of unidentified
algae. A transect of total sulfides, redox potential and volatile solids
at site 1 is shown in Fig. 5.
Site 2 is located in the eastern portion of the Sally's Bend area of
Yaquina Bay. Water velocities w.ere high at times though usually slightly
lower than site 1. Burrowing by frenthic invertebrates was common in this
region. During summer periods, benthic algal growths: were noticeable.
Only limited studies were conducted at this site.
26
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Site 3 is approximately three miles upstream of site 1 on the east
side of the Yaquina estuary adjacent to Parker Slough. Waters here are
slightly less saline, have higher temperatures, and slightly lower tidal
velocities compared to site 2. Burrowing by benthic invertebrates is
extensive, as are mid-summer blooms of the benthic alga, Enteromorpha.
The only studies conducted at this site were measurements of benthic oxygen
uptake rates.
Site 4 was located about 14 miles upstream and 300 feet east of the
Yaquina River bridge at Toledo. Unlike site 1, this site lays in an indus-
trialized area characterized by extensive log rafting and wood processing
operations. The effect of fresh water was reflected in the lower summer
salinities (14-20 parts per thousand). Water temperature was higher than
at site 1, current velocities still fairly high, and the sediments were often
covered by large quantities of bark chips. Burrowing organisms and dense
growths of benthic algae were lacking.
Site 5 was located on the south side of Isthmus Slough in the Coos
Estuary. This site was located on a mud flat which was relatively
protected from the main channel currents by a dike and log storage area.
Tidal velocities were low, temperatures comparable to those at site 2, and
salinities intermediate between those of sites 1 and 2 (28-30 parts per
thousand during summer months). A number of sulfite process woodpulping
mills were located nearby. Extensive algal mats primarily of a salt water
species of Vaucheria were characteristic here, but burrowing organisms
were not evident. Organic content of the sediments was high. A general
purplish coloration to the water was very noticeable due to the photo-
synthetic purple sulfur bacteria.
27
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A comparison of particle size for the three major sites is given in
Table 1. Bacterial counts for sites 1, 2, 4 and 5 are given in Table 2.
PACIFIC
OCEAN
TOLEDO, OREGON
Statute miles
FIG. 4 - SITES LOCATED ON YAQUINA ESTUARY
28
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60O-
400-
200-
is
-200-1
o-
+ 200-
(A) TOTAL SULFIDES
7.2
40 3.1 1.6
TIDAL ELEVATION (feet above MLLW)
i r
-0.8 -2.3
-3.6
(B) REDOX POTENTIAL
i r
-0.8 -2.3
-3.6
TIDAL ELEVATION (feet above MLLW)
(C) VOLATILE SOLIDS
TIDAL ELEVATION (feet above MLLW)
FIG. 5 - TRANSECT AT SITE 1 - SUMMER 1970
29
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TABLE 1. PARTICLE SIZE FOR SAMPLING SITES
Depth
Site 1
Sand (a) Silt
(cm)
0 -
2 -
4 -
7 -
11 -
1
3
5
8
12
92.2
99.3
87.6
93.1
95.3
and clay (b)
7.8
0.7
12.4
6.9
4.7
Site
Sand
15.8
14.0
8.0
9.7
11.9
2
Silt
and clay
84.2
86.0
92.0
90.3
88.1
Site 3
Sand Silt
2.9
1.3
0.6
2.3
3.9
and clay
97.1
98.7
99.4
97.7
96.1
(a) Percent of particles larger than 63 microns.
(b) Percent of particles smaller than 63 microns.
-------
TABLE 2. BACTERIAL COUNTS PER GRAM OF WET SEDIMENT (SUMMER, 1970)
Location Depth
Site 1, 3.1 ft. 0- 1
MLLW
2- 3
4- 5
10-11
Site l,-2.3 ft. 0- 1
MLLW
2- 3
4- 5
Site 2 0-1
10-11
20-21
27-28
30-31
35-36
Site 4 0-1
10-11
20-21
30-31
Site 5 0- 1
2- 3
4- 5
7- 8
11-12
cmCc)
cm •*
«Cc>
cmCc)
cmCc:i
cm^
«) MPN using modified SIM medium or a modified medium for halophilic sulfate
reducing bacteria (30). (numbers per gram of wet sediment)
(c) Determined on 1 date
(d) Average on 2 dates
(e) Average on 3 dates 31
(f) Average on 4 dates
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SECTION VI
STUDIES OF BENTHAL OXYGEN UPTAKE
INITIAL LABORATORY STUDIES
During the early stages of this project, a series of laboratory tests
were conducted to determine the oxygen uptake rate of deposits removed from
site 4 (32) (33). Mud cores were obtained by inserting 3-inch plastic tubes
into the deposit. The tubes, with the deposit, were removed to the labora-
tory where tests were conducted within the same tubes to determine the
benthal oxygen uptake. Mixing within the overlying water was provided by a
plunger type mixing device.
These early laboratory results determined a benthic oxygen uptake rate
2
of approximately 1.9 gm/m /-day under conditions of low mixing and an uptake
2
rate of approximately 3.4 gm/m -day under conditions of higher mixing. It
was found that the depth of the deposits had no effect on the oxygen uptake
rate between the depths of 5.1 and 30.5 cm. When HgCl was added to the water,
the DO uptake rate decreased to one third of its value at both the lower and
higher mixing ranges. These early laboratory studies provided information for
the design of the field respirometer described below.
EXPERIMENTAL DESIGN AND PROCEDURE FOR FIELD STUDIES
Light and dark benthal respirometers developed during this research
were constructed from plexiglas half cylinders, 5.64 meters long by 0.152
meters wide (34) (35). The resulting long and narrow respirometer covered a
2
benthal area (0.813 m ) large enough so that small isolated inconsistencies in
bottom muds would not cause great variabilities in uptake rates. The
long, narrow shape was also required for simulation of actual mixing con-
ditions. In the respirometer designed during this project, velocities
32
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typical of a specific test site were generated over bottom muds by recircu-
lating water in the enclosed long respirometer. A flow development section
was constructed on the inflow end of the respirometer to distribute the
flow evenly over the respirometer cross section. With the volume to area
ratio used, reasonable oxygen uptake rates could be measured in four to
eight hours. Removable flanges were designed so that the flange portion of
the respirometer could be inserted into the bottom deposit some time before
the actual respirometer sections were attached. In this way, the bottom
deposit was allowed to come to equilibrium before attaching the respiro-
meter sections, pump and sampling hoses. The respirometer sections were
made in about 4-1/2 foot lengths and could be installed on the preset flanges
at low water. Rubber gasket material sealed all joints between flanges and
respirometer sections. Water was recirculated in a 1-1/4 inch PVC pipe
using a 1/4 hp submersible impeller-type pump. A 1-1/4 inch brass gate
valve regulated pump discharge and therefore the velocity within the
respirometer.
The respirometer was attached to the preset sealing flanges at low
tide and a typical sampling run was conducted during the time that the area
was covered with water within a tidal cycle. Samples, 20 ml in size, were
brought to the surface and analyzed in the field for dissolved oxygen using
a "Micro-Winkler" modification of the Standard Winkler-Azide method. DO,
temperature and salinity were measured both inside and outside of the respiro-
meter. The water removed for the samples was replaced with estuary water
through a one-way valve which allowed water to flow only into the respiro-
meter as samples were extracted. The displaced volume for all samples on the
longest run was less than 4% of the total respirometer volume. Temperatures
33
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were taken in place with thermistors inside and outside the respirometers.
Diagrams of the benthal respirometer have been published (34) (35).
To measure the amount of oxygen uptake due to free-floating organisms
in the bay-estuary water, planktonic respirometers were developed. Light
and dark planktonic respirometers were constructed of three inch diameter
plexiglas tubes, six feet in length. A small submersible impeller pump was
attached to one end and water was recirculated through the tubular respiro-
meter body at approximately 0.2 feet per second. Exterior hose arrange-
ments used to recirculate the enclosed water were constructed so that the hose
could be brought to the water surface for sampling. Methods of dissolved
oxygen sampling and analysis were the same as those used for the benthal
respirometers .
During the second grant year, a simpler benthal respirometer was also
developed and used. This simplified respirometer had attached flanges, was
4 feet long and had a small recirculation pump not capable of producing the
higher velocities possible with the large respirometer. Both light and dark
respirometers of this type were developed.
Tests indicated that only seven to eight percent of the visible light
was obstructed when passing through the 1/8-inch thick plexiglas from which
the respirometers were constructed.
MATHEMATICAL MODEL OF BENTHAL RESPIROMETER SYSTEM
To more thoroughly explain and analyze the results of the respirometer
study, a mathematical model of the benthal respirometer system was developed.
During the early benthal respirometer runs, excessive leakage was evident at
site 2. The leakage was found to be principally caused by extensive mud
shrimp burrow activity. A mathematical simulation and leakage correction
34
-------
model became a necessity for evaluating leakage as well as helping to determine
the importance of different mechanisms or processes of benthal DO uptake. The
model was applied to all test runs where salinity data were taken, and corrected
DO uptake rates were calculated where leakage existed. Salinity data were
used to evaluate leakage rates.
Definition of terms for the following mathematical model are shown in
Table 3. The mass balance concept is shown by the following expression.
Rate change of sub- Rate of substance Rate of sub-
stance mass within the = input into the - stance output
respirometer volume volume from the volume (3)
From Table 3 and equation (3), one obtains the following salinity balance:
Salinity Balance:
d(Sr) = (Q) (Sw) (QJ (Sr) (4)
dt Vr Vr
It was assumed that the only mechanism of salinity change was by direct leak-
age into and out of the system and that the volume of the respirometer was
constant and completely mixed. 0. and Q2 were assumed to be equal to Q.
Some initial changes in salinity might be caused by bottom scour, but such
changes that might arise from diffusion of material into or out of the bottom
deposit were also considered to be small compared to changes due to leakage;
therefore, terms containing Sb do not appear in equation (4).
A similar approach was used to model the changes in oxygen demand (BOD)
and dissolved oxygen. The following equations resulted.
Biochemical Oxygen Demand (BOD) Balance:
. _
35
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TABLE 3. DEFINITION OF TERMS FOR RESPIROMETER MODEL
Term
Sw
DOw
Lw
Sr
DOr
DOri
Ob
Units
ml/min
ml/min
o/oo
mg/liter
mg/liter
o/oo
mg/liter
mg/liter
Lr
Lri
Vr
K
Sb
Lb
Lbr
mg/liter
mg/liter
liter
1/min
o/oo
mg/liter
mg/min
mg/min
m
Description
Possible leakage into the system.
Possible leakage out of the system, Qi=Q2-
Salinity in the overlying water.
Dissolved oxygen concentration in the
overlying water.
Oxygen demand of the overlying water (BOD).
Salinity in the respirometer.
Dissolved oxygen concentration in the
respirometer.
Initial dissolved oxygen concentration
in the respirometer.
Oxygen demand of the respirometer water (BOD).
Initial oxygen demand of the respirometer
water.
Volume of the respirometer.
Decay coefficient of the oxygen demand.
Salinity of the bottom muds and interstitial
water.
Oxygen demand of the interstitial water (BOD),
Rate of input of oxygen demanding material
into the respirometer system from the
covered mud area (A).
Rate of oxygen diffusing into the covered
bottom mud area (A).
Mud surface area covered by the respirometer.
36
-------
Dissolved Oxygen Balance:
d(DOr) _ (QJ(DOw) (Q) (DOr) n f , Ob
~dt VrVr~~ ' CKJ CLrJ " Vr"
It was assumed that the rate of input of oxygen demanding material into
the respirometer (Lbr) was independent of changes in respirometer BOD (Lr).
No measurements were made to determine the interstitial BOD (Lb). Planktonic
respirometer studies indicated that Lw was negligible. The diffusion of
oxygen into the bottom deposit (Ob) was assumed to be at a constant rate.
CALCULATION OF LEAKAGE
The availability of the benthal respirometer model made possible
the calculation of leakage rates for the respirometer system using only
salinity data for water in the respirometer and outside of the respiro-
meter. By multiple regression analysis, curves could be fitted to
measured salinity data so that the vj-.- • , Sw, and Sr could be evaluated
at any time. Respirometer volume was a constant value of 60.4 liters.
Therefore, all necessary values in the salinity mass balance equation
were known, and estimates of leakage rates that occurred could be calcu-
lated using equation (4).
DO UPTAKE RATE CALCULATIONS
Equations (4) and (6) were solved using the Runge-Kutta method for
finite difference solutions. Fourth order solutions were obtained.
Curve fits were used to input required measured quantities such as
(DOr), (DOw), (Sr), and (Sw). By solving the salinity equation (4)
and the dissolved oxygen equation (6) simultaneously, total oxygen uptake
could be calculated by assuming (-K(Lr)-Ob/Vr) to be the unknown rate
of benthal oxygen demand exerted within the respirometer. When the
build up of BOD within the respirometer and the removal of BOD from the
37
-------
respirometer is small, the total benthal uptake exerted within the
respirometer may be divided by the bottom area of the respirometer to
obtain the benthal oxygen uptake. Neglect of BOD build up and release
will lead to underestimates of the benthal oxygen uptake. Studies
indicated, however, that this underestimate was quite small and could
be neglected for the reported runs. Corrections could then be made for
respirometer leakage. Therefore, even respirometer runs with high
leakage rates could be corrected to obtain estimates of the oxygen
uptake rates.
During most benthal respirometer runs, slightly more rapid rates
of oxygen demand were measured in the first hour of the run than for
any succeeding time period. Furthermore, small increases in salinity
usually occurred inside the benthal respirometers at start-up. Therefore,
during benthal respirometer start-up, water in the respirometer was
replaced by disconnecting the return flow pipe at the PVC union and
allowing the pump to introduce water into the respirometer. Even then,
rapid initial oxygen uptake often occurred. Only DO measurements taken
after the initial rapid uptake were utilized in calculating DO uptake
rates.
DO UPTAKE RESULTS
Light and dark planktonic respirometers failed to show any measureable
oxygen demand or production at either of sites 1,3 and 4. Large mats of
phytoplankton often found at site 3 were not included within these runs.
Large DO variations, often found in the areas in which these dense growths
appeared, indicated that production of dissolved oxygen and respiration by
these growths is significant.
38
-------
Parker Sloueh tests
Large amounts of leakage
O = Normal operation, little leakage
= Severe initial bottom scour
Q
FIG. 6 - SUMMARY OF CORRECTED BENTHAL OXYGEN UPTAKE RATES AT SITES 3 AND 4. (1969)
39
-------
Using the large dark respirometer, sufficient field data to calculate
leakage and uptake rates were collected during four runs at the Parker Slough
site and during seven runs at the Toledo test site during the first grant
year. The DO changes within the respirometer runs were corrected for leakage
as described above (34)(35).
At site 4, benthal uptake rates averaged over each run ranged from 1.4 gm
2 2
0 /m -day to 2.1 gm 0»m -day based on projected surface area. In all runs, the
DO uptake rate decreased with time as shown by the decreasing slope of the curves
shown in figure 6. Leakage rates and actual DO concentrations within the
respirometer also decreased with time. This resulted in apparent relationships
between the uptake rate and the DO concentration and water velocity as shown
in Fig. 7. Whether these relationships are true or whether they merely
reflect the simultaneous occurence of unrelated variations is not apparent at
this time.
Table 4 summarizes the average uptake values measured at site 4 for the
various mixing velocities used. Table 4 also shows the benthal uptake rates
for the same muds measured in the laboratory studies. As can be seen, the
in-lab mixed value is slightly higher than the values for benthal oxygen uptake
measured during this research. Uncontrolled mixing, sample disturbance,
differences in benthal plant respiration or a benthal deposit composition change
might all account for these differences in uptake rates.
Excessive benthal respirometer leakage occurred at the site 3 test site
as was seen by monitoring the salinity changes in the respirometer water.
Leakage rates, calculated by use of the mathematical model, as high as 60
liters per hour were not uncommon. Extreme care during benthal respirometer
40
-------
3-
T3
bo
I
I
c
2i40
1.92
1.44
0.96
0.48
i r
I
J_
012 345
Dissolved oxygen concentration, mg/1
• = 9-3-69, Velocity of 0. 4 fps
V = 9-5-69, Velocity of 0. 4 fps
O = 9-8-69, Velocity of 0.55 fps
A = 9-9-69, Velocity of 0.8 fps
FIG. 7 - VARIATION OF BENTHAL UPTAKE RATES AT SITE 4. (1969)
41
-------
placement did not result in a significant reduction of the leakage. Dye
injections into an operating benthal respirometer at site 3 showed that the
respirometer leaks were caused by the numerous mud shrimp burrows that
penetrated the area. The mud shrimp burrows apparently formed an interconnect-
ing maze of channels in the tidal flat deposit. On several occasions at low
tide, surges of water were observed coming from shrimp holes located upshore
from the low tidal water. The observed surges had the same period as small
waves breaking against the exposed tidal flat. This further substantiated the
inference that the shrimp burrows formed interconnecting networks and could form
passages for water circulation.
2
Computer corrected benthal uptake rates of from 4.8 to 8.5 gm 0 /m -day
were calculated for site 2. Before correction for leakage rates, comparison
of DO changes within the respirometers for the four runs at the Parker Slough
site showed little similarity. The rates of DO change were significantly
different and increases in the respirometer DO were occasionally measured.
After correcting for leakage, the corrected DO changes showed a definite pattern.
Three of the runs displayed quite similar uptake rates (see Fig. 6) while
the fourth run displayed a lower uptake rate. The leakage rates for this
fourth run were quite high and not reliable due to the small measured salinity
difference inside and outside of the respirometer. The most reliable results,
therefore, indicate that the Parker Slough area displayed an oxygen uptake rate
approximately 4 to 6 times greater than that of the Toledo area, or about
2
8 gm/m -day.
Simultaneous measurements using both the dark and light smaller respiro-
meters indicated that the total benthal oxygen uptake depends to a large
extent on the respiration of benthal plants. In regions of relatively high
42
-------
benthal photosynthetic oxygenation, as determined by a transparent benthal
2
respirometer, uptake rates of 3 to 6 gms/m -day were measured at site 1. At
2
this same site, a lower rate of 1.4 gms/m -day was measured in a region of low
benthal photosynthesis. These results, in which the smaller respirometers
were used are probably not as reliable as results obtained through the use of
the larger respirometer.
The larger uptake rates measured at site 3 compared to site 3 (both
measured by the large respirometer as previously discussed) appear to be due
to the large number of shrimp burrows at site 3 and the larger amount of benthal
plant respiration. Though no light respirometer runs were conducted at site
3, the extent of benthal photosynthesis was evident by measured DO concentrations
as high as 25mg/l. Also, the bottom at site 3 was observed to be covered with
fine bubbles (presumably oxygen) on several occasions. Several respirometer
runs had to be rejected because of the release of DO in the form of bubbles
within the respirometer. Collection of the gas was attempted with limited
success. Therefore, only those respirometer runs in which the DO was sufficiently
below saturation to prevent bubble formation were utilized.
A single small dark respirometer run at site 4 measured an uptake rate
2
of approximately 4 gms/m -day. A study of DO variation within the overlying
water indicated, however, that rates as high as 12 gms/m -day might be expected.
Both laboratory and field studies indicated that the DO uptake rate was
reduced by 30 to 40 percent when the system was poisoned with mercuric chloride.
These experiments, however, were quite limited.
A summary of the principal benthic oxygen uptake rates measured in this
study along with examples of rates provided from the literature are snmmarized
in Table 4.
43
-------
TABLE 4. SUMMARY OF BENTHIC OXYGEN UPTAKE RATE
Reference
This study (32) (33)
This study (32) (33)
This study (34) (35)
This study (34) (35)
This study (34) (35)
This study (34) (35)
Mekeown et al (36)
Mekeown et al (36)
Rolley and Owens (37)
Edwards and Rolley (38)
Edwards and Rolley C38)
DO Uptake Rates
gms/m^ - day
1.9
3.4
1.4
1.7
2.1
4.7-8.5
0.8
2.7
1.2
2.8-4.8
29
^ __
Comments
Laboratory estuarine sediments,
(site 4) , low mixing
Laboratory, estuarine sediments,
(site 4), high mixing
In situ (site 4), estuary,
velocity = 0.4 fps
In situ (site 4) , estuary,
velocity =0.55 fps
In situ (site 4) , estuary,
velocity = 0.8 fps
In situ (site 3) estuary burrow
holes present
Laboratory, artificial deposit,
no mixing
Laboratory, artificial deposit,
mixing
Laboratory, river muds, frequency
distributions provided
Laboratory, river muds, several
temperatures, magnetic stirring
Laboratory, river muds, magnetic
Hanes and Irving (39)
Stein and Denison (40)
Pamatmat and Banse (41)
3.2
5-6
0.6-1.2
stirring, bottom scour
Laboratory, magnetic stirring
Laboratory, Magnetic stirring,
some scour
In situ, Puget Sound, belljar,
stirring prop., deep water
44
-------
TABLE 4. SUMMARY OF BENTHIC OXYGEN UPTAKE RATE (cont.)
Reference
DO Uptake Rates Comments
, gms/m2 - day
Fair et al (42) 1.2-4.6
Baity (43) 1.8-5.4
Ogurrombi and Dobbins (44)
O'Connel and Weeks (45) 4.4
O'Connel and Weeks (45) 0.15-8.5
Laboratory, continuous water flow,
artificial deposit, long term
Laboratory, continuous water flow,
artificial deposits
Laboratory, artificial deposits,
magnetic stirring
Laboratory, artificial deposit,
partial scour
In situ, water current mixed
SENSITIVITY STUDY
To further understand the mechanisms that control benthal oxygen uptake,
sensitivity studies were made using the mathematical model previously
described. Numerical approximations to the solutions of these equations were
obtained by fourth order Runge-Kutta methods. By varying parameters in the
dissolved oxygen equation and the BOD equation, curves of oxygen uptake similar
to those measured were developed. It was found that, with the equations pre-
viously proposed, simulated DO variations corresponded to actual measured DO
variations only if a major portion of the oxygen demand was due to one of the
following:
A. the transport of oxygen from the water into the deposits,
B. demanding material released from the deposit into the water or
C. combination of A and B from above.
45
-------
First order decay coefficients,. K, needed in assumption B were 10 to 100
times greater than those normally encountered for BOD tests of polluted
waters. These results suggest that, under normal water flow conditions, the
principal oxygen uptake due to tidal flat deposits occurs in the immediate
vicinity of the deposits. Sulfides released from the anaerobic regions could
contribute toward such an uptake, though later research indicates that this
was doubtful at this location.
It should be recognized, however, that respirometer studies of approxi-
mately six hours may be too short to accurately measure the long term release
of BOD. High BOD values in tidal flat waters were not found, indicating that
such a mechanism was probably not substantial. Accurate measurements of BOD
(possibly COD) throughout a respirometer run might lead to better estimates
of such release. In general, it was found, however, that the BOD values were
too low for sufficient accuracy.
46
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SECTION VII
FREE SULFIDE IN OVERLYING WATER
GENERAL
During the second year of the project, it became apparent that the
sulfur cycle was of major importance. From an environmental quality
viewpoint, the benthic release of free sulfide was considered to be of
major interest for three reasons: (1) Free sulfides exert an oxygen
demand, (2) free sulfides are highly toxic to a wide variety of organisms
and (3) the release of hydrogen sulfide to the atmosphere could be a
major source of atmospheric sulfur.
In order to obtain free sulfides within estuarine benthic deposits
it is necessary to have a sufficient production of sulfides, primarily
through sulfate reduction, for a long enough duration of time to suf-
ficiently reduce the available iron. Conditions favorable for the presence
of benthic free sulfide include high salinity (and thus high SO.), slight
wave and current scour of bottom (to prevent oxidation of ferrous sulfide
and "recycling" of iron) and a high organic content within the deposits.
Site 5 appeared to best provide all of these conditions; moreover, the
bright purple color at portions of the mud surface and under algal mats
indicated the presence of photosynthetic purple sulfur bacteria and thus
the presence of free sulfides.
Even if free sulfides were present within the deposits, it was
questionable at that time if they would be found at levels sufficient for
measurement within the overlying water if the dissolved oxygen in this
47
-------
water were not depleted. This question was examined by a number of mathe-
matical models at the same time that analytical methods for measurement of
sulfides were examined. The assumptions made to develop these models
generally contribute to conservative (low ) estimates of free sulfide con-
centrations in oxygenated waters. The following two sections describe two
models used in this portion of the study.
MODEL OF FREE SULFIDE TRANSFER THROUGH AEROBIC ZONE
Consider a vertical column of deposit of cross-sectional area, A. While
oxygen demanded substances (free sulfides) diffuse upward, DO diffuses downward
from the surface. A second-order reaction between the DO and these substances
is assumed. The depth of the deposit, z, is taken as zero at the surface with
positive values increasing with depth. The chemical loss rate of DO per unit
volume of deposit, G, was taken as
G = yOCAndz (7)
in which
n = the volume of interstitial water within the slice
divided by the total volume of the slice,
0 = the DO concentration,
V = the second order reaction coefficient and
C = the concentration of oxygen-demanding material
The diffusive and advective flux, F, of oxygen within the deposit was
taken as
30
F =" D0Am IF + Aq° (8)
in which
m = the fraction of the surface area open to diffusion,
48
-------
q = the water transport rate per unit area, and
D = the effective diffusion coefficient for DO.
o
Performing a mass balance on a segment slice of depth dz and taking A,
n, D m, and q constant with distance, and A and n constant with time leads to
80 o 80 q 90
W=—~2nW-
oZ
As an approximation, both m and n may be taken as equal to the porosity.
In the following, advection will be taken as zero. Thus equation (9) reduces
to
- yOC (10)
Similarly, for the oxygen -demanding substances,
W ' Dc
is obtained in which D is the effective diffusion coefficient for C.
c
The benthal oxygen uptake, OD', caused by materials included in C,
is equal to the sum of the DO flux into the deposit and the flux of C out
of the deposit. Thus
an 3P
OD' = (Dm|Ł) - (D m^O (12)
o 8z Z=Q c 3z z=Q
Respiration of benthal algae and oxygen demand of materials not included
within C will contribute to the benthal oxygen uptake rate but are not included
within OD'.
49
-------
Approximations to the solutions of Equations 10 to 12 were obtained
by explicit finite difference methods. (46) Several runs were also
satisfactorily compared with an implicit finite difference scheme.
The relationship between the free sulfide concentration, S, and C will
depend on (a) the mass of free sulfides oxidized per mass of DO utilized
and (b) the fraction of C caused by free sulfides. As an approximation, S
may be taken as 50 to 100 percent of C. Let
P .^H-° C13)
W OD!
The model results indicated that P ranged from 0.4 to 0.8 with highest
values associated with low diffusion coefficients, low oxidation rates (low y)
low dissolved oxygen concentrations in the overlying waters and high benthic
oxygen uptake rates . Additional benthic oxygen demand not included in the
model would result in a further reduction of DO penetration into the deposit
and thus higher values of P . Thus, the computed values of P may well be
underestimates. The results indicate, therefore, that substantially more
than half of the upward diffusing free sulfide would be released to the
overlying water rather than being oxidized within the deposit.
MODEL OF FREE SULFIDE IN OVERLYING WATER
Consider a vertically mixed column of water of depth H and horizontal
area A. A deposit at the bottom of the column has a total oxygen uptake rate
given by OD. The upper surface of the water column is exposed to the air,
performing a mass balance of sulfides within the column and assuming a
steady state leads to
YP P OD
s =
50
-------
in which
S = the sulfide concentration,
Y = the mass of S oxidized per mass of DO utilized,
P = the fraction of OD resulting from free sulfides formed
within the deposit,
K = the first order decay coefficient for free sulfides
(equal to yO),
K. = the transfer coefficient of H?S across the air-water interface
and
f = the fraction of the free sulfides present as H?S.
Because of the simplifying assumptions of equation 14, large differences
between computed and observed value, should be expected. Estimates of
parameters for equation (14) which might apply to late summer conditions at site 5
are given in Table 5. It is emphasized that these parameters were rough
estimates based on a varied and limited amount of information.
The range of free sulfide concentration, S, in the overlying water
computed by equation (14) using the parameters shown in Table 5 is enclosed by
the solid lines of Figure 8. These results illustrate that significant
sulfide concentrations in the overlying water may occur under conditions
similar to those found at Site 5.
During the late summer and early fall of 1970, water directly above the
sediment of the four sites was monitored during a tidal cycle for dissolved
free sulfides and DO. Profiles of DO and free sulfides above the bottom were
determined by simultaneously drawing water into six 50-cc plastic syringes
attached by small-bore tygon tubing to six l.S-rnm inside diameter stainless
51
-------
Cn
ro
100
10
1.0
O.I
Approx. DO
(mg/L)
o
4.0
8.0
0.01
O.I 1.0 10
FREE SULFIDES (mg/L)
100
FIG. 8 - RANGE OF FREE SULFIDES IN OVERLYING WATER.
-------
TABLE 5. ESTIMATED PARAMETERS APPLICABLE TO SITE 5
DURING THE LATE SUMMER
AND EARLY FALL PERIOD
Parameter Range
Y 1.0
Pw 0.8
Ps 0.3-0.4
OD 4.0-12.0 ( gins /m2 -day)
K 8.0-70.0 (per day)
KA 0.5 (meters/day)
Ł 0.4-0.5
Reference
(12)
This work
(32) (33)
This work a
(7) (13) (16) b
(58) (59) (46) c
d
a - estimated from field studies.
b - obtained in sea and brackish water.
c - reference (59) shows similarity between DO and ^S transfer; reference
(58) gives DO transfers for low wind velocities.
d - based on pH of approximately 7 at site 5.
53
-------
steel tubes set horizontally at varying heights above the deposit. Determina-
tion of DO was made on 20-ml water samples collected in 30-cc plastic syringes
by a micro-Winkler method (32). Free sulfides were determined by immediately
fixing 20 ml of water sample inside the syringe with an equal volume of 50
percent antioxidant buffer solution. The standard solution, which contained
320 g, sodium salicylate, 72 g ascorbic acid, and 80 g sodium hydroxide and
made up to 11 in distilled water, prevented further oxidation of sulfide and
fixed the free sulfides as essentially all free sulfide ions (6). Sulfide
content was then determined by measuring potential on a pH meter equipped
with a sulfide membrane electrode and reading the sulfide concentrations from
a standard curve developed prior to each run (47) . This method was compared
with results obtained titrimetrically with iodine (48) and titrimetrically with
lead perchlorate using the probe as an end-point indicator (47). Results
showed close agreement among all three methods.
The variable ionic strength of estuarine waters was a source of error
in the sulfide measurements. Calibration of the probe was done with distilled
water with the antioxidant buffer solution and thus reported free sulfide
concentrations can be expected to be as much as 20% low. The method of free
sulfide determination was later changed to the subtraction method (47) to
avoid this problem. Free sulfide and some oxidation products of sulfide can
be expected to result in lower measurements of DO. Thus reported DO values
may also be underestimated particularly at low DO and high free sulfide values.
Free sulfide and DO profiles measured during daylight conditions at
Site 5 are shown in Fig. 9. Because of the benthal photosynthetic oxygenation,
DO values were highest near the mud-water interface. Significant free sulfide
concentrations were measured despite the relatively high DO values. Higher free
54
-------
DO (mg/liter)
5.0 10.0
15.0
0.5 1.0
FREE SULFIDE (mg/liter)
1.5
FIG. 9 - DISSOLVED OXYGEN AND FREE SULFIDE PROFILES AT SITE 5
55
-------
2400
I I I I I
I 1 1 1
2000
FIG. 10 - DISSOLVED OXYGEN AND FREE SULFIDES AT SITE 5
56
-------
sulfide concentrations were measured in the low mixed water region immediately
adj acent to the mud-water interface.
Results from four sampling periods at site 5 are shown in Fig. 10.
Samples for the runs shown in Fig. 10 were collected above the one centimeter
distance from the bottom. These results demonstrate the significant free
sulfide concentrations can occur in waters at site 5 even in the presence
of dissolved oxygen. The relatively stable free sulfide concentrations
shown in Fig. 10 are plotted with approximate water depth on Fig. 8.
These free sulfide concentrations were within the range roughly estimated by
equation 14 and the parameters of Table 5.
Free Sulfide Measurements at Other Sites: Attempts were made to measure
free sulfides in the overlying waters of sites 1 and 4. Free sulfides were
not detected at site 1. At site 4, free sulfides were detected at low levels
(1 mg/1 or less) only for a few samples collected when wave action appeared
to disturb the bottoms in the immediate vicinity of the sampling. These
occasional measurable levels may have been false readings due to the inter-
ference of iron on the sulfide probe. Such measurements at both sites,
however, were quite limited and were not sufficient to determine possible
seasonal releases of free sulfides.
During the late summer and fall, the Enteromorpha and Zostera
present at site 1 begin to decompose (49). During this same period, redox
potentials dropped dramatically in the upper few centimeters of the sediment,
microinvertebrates were observed to migrate from the sediments into the
overlying algal material, and the smell of hydrogen sulfide was noticeable (49)
It is likely that sulfide release may occur at site 1 during this period
where benthic algal growth is substantial, however, sampling was not
conducted during this period.
57
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SECTION VIII
CONDITIONS WITHIN BENTHIC SYSTEMS
FIELD MEASUREMENTS
While it is true that estuarine benthic systems can strongly influence
the overlying air and water quality, they must not be viewed only as a
boundary condition of the overlying regions. Benthic systems are of major
importance to the total estuarine systems . If they are to be understood, a
variety of sediment and water quality measurements must be taken within the
benthic systems.
During the final grant year of this project a field investigation of
benthic deposits within tidal flat regions was conducted. Sediment cores
were collected at sites (usually during low tide) using plexiglas coring
tubes. Cores from which interstitial water was obtained were extruded into
a plexiglas slicing trough and sliced into desired sections. Sections
were immediately placed into field presses so as to prevent chemical oxidation.
Interstitial water was extracted in the field within thirty minutes after
collection of deposit cores. Four field nylon presses (50) were constructed
for this purpose. Pressure not exceeding 150 psi was applied through the use
of nitrogen gas. Chemical determinations were either conducted immediately in
the field as done with free sulfides or were suitably fixed and determined
within several days. Soluble organic carbon (500) was measured on a Lira
Model 3000 total carbon analyzer. Sulfate was determined by the method of
Bertolacini and Barney (51). Free sulfides were determined with a sulfide
probe using the subtraction method (47). Chlorides were titrametrieally
determined (48).
58
-------
Redox potentials were determined in the field by fixing the cores in a
vertical position, placing a reference electrode on the sediment surface
and inserting a 1mm platinum wire into the core through small holes drilled
at appropriate intervals in the plexiglas core tubes. The reference electrode
used was a standard fiber junction reference electrode used in pH, measurements,
modified by fastening a fine-frit Gooch crucible about its tip, and filled
with saturated potassium chloride solution. Measurements were made following
standardization in a solution of known redox potential.
Total sulfides were determined by a modification of the standard titri-
metric method (48]. Volatile solids were determined by drying and combustion (48)
Samples used for total sulfides and volatile solids were kept in the plexiglas
cores and cooled or frozen until analysis.
Examples of sediment profiles are shown in Fig. 11. As expected, free
sulfides within the interstitial waters were highest at site 5.
A CLASSIFICATION OF ESTUARINE BENTHIC SYSTEMS
Under conditions of relatively constant inorganic and organic inputs,
five types of estuarine benthic systems described in Table 6 can develop.
These five types are determined by the exogenous hydrogen acceptors avail-
able to decompose the deposited organics and the amount of iron (and other
metals which form insoluble sulfides) largely from deposited inorganics,
available to react with free sulfides. Further subdivision of these five
types based, as an example, on the extent of methane fermentation or pyrite
formation are possible, but will not be discussed herein.
Fig. 12 qualitatively illustrates the general response of the estuarine
benthic systems described above to different continuous inorganic and organic
loading rates. The five regions in Fig. 12 correspond to the five types
59
-------
S04- MG/L
1000 2000
FREE SULFIDE-MG/L
30 60
REDOX POTENTIAL-MV
-200 -100 0 +100
IOO 2OO 3OO
SOC-MG/L
1000 2000
TOTAL SULFIDE-MG/KG
10 20 30
VOLITILE SOLIDS-%
FIG. 11 - EXAMPLES OF ESTUARINE BENTHIC TYPES
60
-------
TABLE 6. CLASSIFICATION OF ESTUARINE BENTHIC SYSTEMS
Deposit Aerobic
Type Decomposition
1 Dominant
2 SigniŁicanta
3 Significanta
4 Limited to a>b
Significant
5 Limited a'b
Sulfate
Reduction
Limited
(Organic limiting)
Significant
(Organic limiting)
Significant
(Organic limiting)
Significant
(Organic limiting)
Significant
(Sulfate limiting)
a - Dependent on DO in overlying water.
b - Aerobic zone in deposit limited by free sulfides.
c - Also increased accumulation of organics particularly
Intersitial
Free Sulfide
Low
Low
Low
High
High
if conditions are not
Methane
Fermentation0
Small
Small
Significant
Small
Significant"
favorable in
Methane Fermentation.
d - Possible inhibition by free sulfides.
-------
ORGANIC DEPOSITION RATE
(Expanded Scale)
FIG. 12 - QUALITATIVE DESCRIPTION OF ESTUARINE BENTHIC SYSTEM RESPONSE TO CONSTANT
DEPOSITION CONDITIONS.
62
-------
described in Table 6. The solid lines of Fig. 12 are positioned by the
availability of hydrogen acceptors while the dashed line is positioned by
the availability of iron. The precise quantitative definition of each of
these five regions depends upon a wide range of conditions including, but
not limited to, the amount of available iron in the inorganics, nature of
organics, hydraulic conditions, extent of biological turnover, and con-
centrations of DO and sulfate within the overlying waters. A quantitative
description of each of the regions in Figure 12 is not now possible. Fig. 12,
however, does illustrate the general response of the system shown in Fig. 2
to different constant deposition conditions. Qualitative changes in Fig. 12
due to different conditions can be pictured. Larger amounts of available
iron within the inorganics, as an example, would lead to clockwise rotation
of the dashed line defining regions 4 and 5. Smaller amounts of iron within
the inorganics would result in a counter-clockwise rotation of this dashed
line. Fig. 11 may not be applicable for very large deposition rates and does
not apply well when periodic scour occurs.
Past investigations of marine benthic deposits give some quantitative
descriptions of the different regions shown in Fig. 12 . Results of the
Puget Sound study (52) demonstrated a large decrease of benthic fauna (both
in numbers and types) and a strong hydrogen sulfide odor when volatile solids
became greater than ten percent of the total solids. This percentage might
roughly define, for these locations, the upper bound of region 5 in Fig. 12 .
Different benthic estuarine systems will support different sets of
plants and animals. The amount of DO available for respiration, the toxic
effect of free sulfides, and the amount and suitability of organics for
food supply will all serve to determine the nature of resident populations.
63
-------
Other factors, such as particle size distribution, salinity, and temperature,
will also determine the composition of benthic communities.
The regions described in Fig. 12 define the steady state benthic types
(shown in Table I) that would be approached if a given set of deposition
conditions persisted. The actual types present at a given time are a result
of past depositions. Current loading conditions define types which the
systems are then approaching. When loading conditions change, systems of one
type may shift toward a different type. Seasonal changes of benthic loadings
within Oregon estuaries appear to produce seasonal changes of types. Sufficient
data, however, are not now available to describe such changes adequately. The
data to the date of this report indicate that seasonal scour is an important
factor in estuarine benthic systems. The above classification system has the
shortcoming that a relatively constant deposition rate is assumed. Where
periodic scour occurs, some confusion can result from the use of this classification
method. A new classification method which accounts for scour is now being
developed.
The results presented in Fig. 11 can be used to illustrate three of the
five classification systems. Fig. HA demonstrates an example of a type 2
benthal system (though this system may be approaching type 4}. Below the 5 cm
depth, organics likely become limited. Free sulfides were detected at low levels
(less than 1 mg 1 ) only within the top 2 cm of the deposit. Burrows in this
region likely contribute toward the transport of oxygenated waters to deeper
portions of the deposit. Moreover, water flow through the deposit during low
tide likely laterally transported oxygenated water to the deposit. Such
transport leads to the oxidation of H2S and FeS (note positive redox potential
of the greater depths). This oxidation can result in the formation of elemental
64
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sulfur which may serve as an oxidizing agent leading to the formation of
pyrite (53) (note the decrease of total sulfides with depth). The formation
of pyrite has not been illustrated in Fig. 1, and the decrease of FeS may
have been due to a more complete oxidation.
Fig. 11B presents a type 3 benthal system in which sulfate becomes
limiting with depth. The positive redox potential within the first centi-
meter indicates that aerobic decomposition was likely significant. No
detectable free sulfides were measured.
Fig. 11C shows an example of a type 5 benthal system. Sulfate becomes
limiting and soluble organic carbon increases at greater depths. Free sulfides
within the interstitial waters were measured at levels above 60 mgl . The
concentration gradient of free sulfides indicates that free sulfides were
diffusing upward where oxidation occurred and downward to likely combine with
available iron. A negative redox potential was measured throughout the depth.
Aerobic decomposition was likely reduced by the upward diffusion of free
sulfides .
The highest percent of volatile solids was measured in the type 5 deposit
(Fig. 11C) indicating a high organic to inorganic deposition ration. The
lowest such ratio is suggested for the type 2 deposit (Fig. 11A) which had
the lowest percentage of volatile solids.
65
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SECTION IX
STUDY OF SULFATE REDUCTION USING EXTRACTS
GENERAL
At approximately the mid-point of this study, it became apparent that
the sulfate reduction within estuarine benthic systems was of major interest.
Sulfate reduction was examined by three different approaches: (A) Sulfate
reduction in organic extracts, (B) Sulfate reduction in mud slurries
employing S-35 tracer and (C) incubation and measurement of entire benthic
cores. Due to time constraints, the major investigative effort was given
to method A.
As previously discussed, the major producers of free sulfides in
marine and brackish water appear to be sulfate reducing bacteria. Those
organisms, belonging chiefly to the genus Desulfovibrio, are ecologically
quite versatile, and are ubiquitously distributed in nature (54). They
occupy habitats embracing a wide range of pH, salinity, Eh, temperature, and
osmotic and hydrostatic pressure.
Most cultures, however, appear to grow best between a pH of 6.2 and 7.9
and an Eh of -50 to -150 mv (54). Sulfate reduction itself tends to lower
Eh and raise pH of environments in which it occurs, the magnitude of such
effects depending upon the buffering capacity of the medium, and the end-products
of the oxidation-reduction process (54).
Based upon salinity tolerance, there appear to be two general, although
somewhat indistinct, physiological types. Those found within soil, sewage,
and fresh water are most active in solutions of less than one percent sodium
66
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chloride, and become inhibited at 1.5 - 3.0 percent concentrations. The
other group, occurring in marine and brackish waters, appears to require
sodium chloride solutions isotonic to sea water or sea water itself (54).
The trace mineral requirements of these organisms are but imperfectly
known and probably quite variable. Ferrous iron is essential, due to the
presence of a cytochrome system in species of Desulfovibrio (55).
Growth of sulfate reducers has been observed at temperatures ranging
from -11° to 104°C, but the majority occur in the ocean floor sediments at
temperatures below 5°C. They appear to grow best at 15 - 40°C(81).
The organic acids as a group (lactate, pyruvate, maleate, citrate,
propionate) appear to be the most readily available and preferred energy
source for sulfate reducers (54). In addition, fatty acids, simple alcohols,
and some mono and disaccahrides are suspect. Complex carbohydrates do not
appear to be directly utilizable, but the importance of other microorganisms
in the breakdown of these to utilizable forms has been noted (56).
Sulfate ions appear to be by far the most common hydrogen acceptor (55).
They are usually in abundant supply in sea water, and there is good evidence
that the process of sulfate reduction may have been dominant and had global
implications during past eras (57). There is some evidence suggesting that
sulfate reduction is not limited until the sulfate concentration drops below
10 mg/1, but it is probable that halophilic strains become limited at higher
concentrations. In estuaries, sulfate reduction in bottom deposits is
dependent on both a supply of sulfate and organic material within the inter-
stitial and overlying waters, and sulfate may become limiting at the head, and
organic material limiting at the mouth of estuaries (31).
67
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In addition to sulfates, the use of sulfite, thiosulfate, hydrosulfide
and several other sulfur oxides, as hydrogen acceptors, has been demonstrated
(54). These compounds, however, are not generally widely available in nature,
and are considered to be of considerably less significance. Since more
energy is derived from the more oxidized form, it is probable that it would
be preferentially utilized when available. The ability of sulfate reducers
to utilize elemental sulfur is questionable (55,54). Some autotrophic strains
are apparently able to utilize carbon dioxide and the bicarbonate ion as a
hydrogen acceptor (54).
The question of whether the activity of sulfate reducing bacteria in nature
and in synthetic media is affected by the hydrogen sulfide (or other free
sulfide) produced has been examined by several investigators (54,55,60).
It appears that the levels of free sulfide which may be tolerated are criti-
cally dependent upon pH, available sulfate, the nature of the energy source,
and the presence of cations which may form insoluble sulfides (54).
MEDIA PREPARATION
Sediment extract media were prepared as follows from sediment collected
from the upper five centimeters of deposits.
A. One liter of distilled water was added to one liter of sediment in
a large erlenmeyer flask, thoroughly mixed, and the resulting
slurry autoclaved for 45 minutes at 121°C. After removal from the
autoclave, the slurry was allowed to cool and the sediment removed
by centrifugation. The resulting clear extract was either utilized
directly as growth media, or lyophilized to produce a powder. In
some cases this powder was added to liquid medium to yield one of
higher organic concentration. Sulfate concentration were increased
by addition of sodium sulfate, or decreased by adding barium
68
-------
chloride. The pH was adjusted to near neutral by addition of sodium
hydroxide.
B. One liter of sea water was added to one liter of sediment and
treated as with the addition of distilled water as described in
method A.
C. One liter of either three or ten percent hydrochloric acid was
added to one liter of sediment and treated as with the addition
of distilled water as described in method A.
Algal Extract Media were prepared from the algal mat collected at site 5
by the following methods:
D. Two liters of algae and its associated water were placed into a
three liter erlenmeyer flask and autoclaved as with sediment
media. Following cooling, the algal material was squeezed using
a wooden fruit press, and the liquid collected and centrifuged.
Sulfate concentrations and pH of the media were adjusted as
previously described.
E. Approximately 250 ml of 10 percent hydrochloric acid was added to
500 ml of algae and treated similar to that in method D.
F. Small amounts of liquid were extracted from both sediment (collected
from the upper five centimeters at the sites') and algae (site 5)
by using a hydraulic press and specially designed squeezing
cylinder. In addition, approximately one liter of extract
was prepared by squeezing by hand algae which had been freshly
collected from site 5.
The results of preparing media by the various methods of organic
extraction are summarized in Table 7. Early attempts at preparing media
69
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TABLE 7. SUMMARY OF SOLUBLE ORGANIC EXTRACT PROCEDURES
Sample
Material
Site 1 -
sediment
Site 2 -
sediment
Site 3 -
sediment
Site 3
algal mat
Extraction Soluble organic(a) Sulfate(a)
procedure carbon (mg/1) (mg/1)
A(b) 150 - 250 nm. '
fbl
C 150 - 300 nm.
F (b) 50 3200
B 100 - 150 nm.
A 100 - 800 10 -
c(b) 3200 nm.
Ftb) 60 - 110 100
A^ 200 - 300 200
C^ -* 1600 - 3400 1400
F^ 140 - 310 1000 -
D 1500 - 3000 3400 -
ECb) 5400 nm.
p(b) 2010 4800
F(hand squeezed) 1125 2150
C<0
300
1500
4000
(a)
(b)
(c)
Approximate.
These extracts not used for media.
Not measured.
70
-------
from sediment (methods A and B) inevitably resulted in media having fairly
low organic carbon concentrations (100-800 mg/1). Extraction methods with
acid resulted in higher concentrations of organic carbon, but a large
percentage of this organic carbon precipitated out upon adjustment of the
pH with sodium hydroxide to neutral. Hence little was gained in the final
media by using acid extraction.
An analysis of sugars and organic acids was performed on a number of
extracts. The general results indicated that the pentoses present at site 4
likely reflected the input of wood products to this area. Sediment extracts
from site 1 showed small amounts of both pentoses and hexoses suggesting that
the organic material there may come from more diverse sources. The relatively
lower levels at site 1 reflect the lower organic concentrations measured
here. At site 5 high concentrations of mannose and glucose, both hexoses, as
well as high levels of propionate, butyrate, and acetate were present.
METHODS
A number of experimental approaches were attempted in this study. Only
the last approach, which took advantage of previous experience, is reported.
Approximately 600 ml of algal extract medium (prepared according to method
D) were placed into each of twelve 500 ml erlenmeyer flasks, organic concentra-
tions adjusted by dilution, and desired sulfate levels achieved as previously
described. Sodium chloride was added where necessary to adjust the chlorinity
of each culture to approximately 20 ppt. Two additional cultures were simi-
larly prepared using media prepared by hand squeezing algae from site three.
Oxygen was initially removed by sparging for 15 minutes with carbon dioxide-free
f
nitrogen. The pH was then adjusted to 7.5 - 8.0 by addition of 3.0 N sodium
71
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hydroxide, and Eh lowered to approximately -100 mv by addition of a small
quantity of sodium sulfide. Each flask was inoculated with two ml of mixed
culture and immediately capped by a rubber stopper fitted with a glass tube
and serum cap. To prevent leakage, modeling clay was liberally applied
around the edges, and the stopper further fastened down by masking tape.
Each flask was shaken to disperse the inoculum, and initial samples were
taken for analysis. The flasks were then incubated in the dark at 20°C. A
control was set up by filling several tubes with sterile extract, and capping
tightly. No significant changes were measured in the control series.
Samples were withdrawn from the flasks at appropriate intervals with a
syringe which had been flushed and prefilled with nitrogen. By exchanging
the gas for the sample, the flask was maintained anaerobic and development
of a negative pressure due to extraction of the samples was avoided. Flasks
were shaken thoroughly prior to sampling in order to produce a fairly homo-
geneous medium, giving a more representative sample. Total sulfide was
measured by adding 10 ml of sample to a known volume of acidified 0.025 N
iodine solution and back-titrating with 0.025 sodium thiosulfate.
Sulfate was determined by a colorimetric procedure using barium chloranilate
(51). Samples were passed through a Dowex 50w-x8 20-50 mesh H cation exchange
column to remove interferring ions, diluted to 40 ml if necessary, and added
to 50 ml of 95 percent ethanol and 10 ml potassium phthalate buffer. Approxi-
mately three grams of barium chloranilate were added to precipitate the sulfate.
After shaking for ten minutes, the solution was filtered, and the optical dens-
ity determined on the filtrate with a spectrophotometer. Sulfate concentrations
were read from a standard curve.
Soluble organic carbon was used as a measure of the soluble organic
material present. Five ml of centrifuged sample were placed into 20 ml
72
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screw cap test tubes in an ice bath, and carbonate carbon removed by acidifying
to pH 2.0 - 3.0 with three percent phosphoric acid and sparging with carbon
dioxide-free nitrogen for 10 minutes. Determination of the remaining soluble
organic material was made using a Lira Infrared Analyzer Model 3000. Eh was
measured with a platinum wire electrode, and pH with indicator paper.
Fourteen cultures were run.
Cultures 1 through 12 contained media prepared from autoclaved algae
(method E), whereas cultures 13 and 14 contained media prepared from squeezed
unautoclaved algae (method F). During the initial portions of each run,
sodium hydroxide was added to maintain a satisfactory pH. After several
days, stable conditions of pH and Eh were achieved, further addition of base
became unnecessary, and active sulfate reduction began to occur. Day 0 (zero)
of the experiment was set when the pH and Eh became stabilized. This occurred
three to five days following inoculation.
RELATIONSHIP BETWEEN SULFIDE, SULFATE, AND ORGANIC CARBON
Dissimilatory sulfate reduction requires an organic source for energy, and
sulfate as a hydrogen acceptor (20,55). The stoichiometric relationship for sul-
fate reduction is given as (20)
SO^ + 2C . -*• S= + 2C00
4 organic 2
-(12) -(3) +(4) (15)
where the numbers below the chemical formulas indicate the reaction on a
weight basis. Thus during sulfate reduction, a production of 1.0 mg of
sulfide will theoretically require 3.0 mg of sulfate and 0.75 mg of organic
carbon. These 'yield ratios' have been calculated for the mixed cultures
73
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TABLE 8. EXPERIMENTAL YIELD RATIOS AND MAXIMUM RATES
OF SULFIDE PRODUCTION FOR EACH CULTURED)
Culture No. Y (b)
1 3.2
2 3.0
3 3.0
4 3.2
5 3.0
6 3.4
7 3.1
8 3.2
9 3.0
10 3.2
11 3.0
12 3.1
13 3.1
14 2.9
Theoretical 3.0
Y (c)
0.82
0.89
1.60
2.00
2.30
6.05
0.86
0.86
0.91
0.80
0.88
1.10
0.94
0.84
0.75
V (d)
max
54
60
54
48
35
10
70
62
51
30
20
10
54
34
—
Sulfate
Limitation
No
Slight
Slight
Large
Large
Large
No
No
No
No
No
No
No
No
(a) Yield ratios (Y and Y ) calculated over the 14 days of
the experiment. soc
O) (mg/1 of sulfate utilized) / (mg/1 of sulfide produced).
(c) (mg/1 of soluble organic carbon utilized) /(mg/1 of sulfide produced)
(d) The maximum rate of sulfide production (mg/1-day) as measured
over three day intervals, during the 14 days of the experiment.
74
-------
over the 14 days of experiments, and are presented in Tables 8 and 9 where
_ mg/1 of sulfate consumed
L ~ mg/1 of sulfide produced (16)
and
_ mg/1 of soluble organic carbon consumed
C mg/1 of sulfide produced (17)
It is important to recognize that these yield ratios reflect the activity of
all the bacteria within the mixed cultures of the experiment and not just the
sulfate reducers.
TABLE 9. SUMMARY OF YIELD RATIOS
SulfateV sulfide Carbon/sulfide
Y Y
TL *C
Theoretical 3.0 0.75
Average overall runs 3.1 1.49
Average overall runs
with no sulfate limitation 3.09 0.89
Agreement between the theoretical value of Y for sulfate reduction
and the values of Y determined from the experiment was good. The slightly
L
higher experimental values may be due in part to assimilatory reduction of
sulfate during initial growth of all of the bacteria present in the cultures.
Y_ was more variable than Y and exceeded the theoretical value of
C ij
0.75 in every culture. Since these mixed cultures contain bacteria other
than sulfate reducers, which are capable of ^.utilizing the organic carbon,
this result is not surprising. That these other bacteria are capable of
75
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such utilization was apparent from the drop in soluble organic carbon in
the absence of measurable sulfate reduction prior to day 0. In general,
where sulfide production became curtailed by deficiency of sulfate, the value
of Y increased well above the theoretical, as would be expected.
LJ
RATES OF SULFATE REDUCTION
The agreement between the theoretical and experimental yield ratios shown
in Tables 8 and 9 indicate that sulfate reduction was responsible for the sulfide
production measured. The maximum rate at which this sulfate reduction occurred
(expressed as mg. of sulfide/1-day) in the extract experiments is shown (V )
nicix
in Table 8. Rates have been estimated over a three day interval in each culture
in which the maximum rate appeared to be occurring.
Results obtained in cultures 9 and 10 correspond closely to those of
culture 13 and 14 respectively. (see Figs. 21, 22, 25, 26, and Table 8).
This close agreement indicates that the effect upon sulfide production of
autoclaving the algae used in the preparation of cultures 1 through 12 was
not significant. The media for cultures 13 and 14 were obtained by hand
squeezing the algal mat found at site 5.
To further explain the results of the extract experiments a mathematical
model was used which relates the production of sulfide to the utilization of
sulfate and soluble organic carbon. Results indicated that the sulfide
production rate increased as the concentration of sulfate and/or organic
carbon increased. This effect, however, was most pronounced at lower levels
of sulfate and carbon, and became relatively small at higher concentrations.
An equation which has been used to describe such a saturation effect at high
substrate concentrations is the common Michaelis-Menton equation shown below:
76
-------
= Rmax CIS)
in which P is the concentration of the product. R the maximum rate of
max
product formation, N the substrate concentration, and K the substrate
concentration at which dP/dT = 1/2 R
max
With sulfate and soluble organic carbon serving as substrates, the
rate of sulfide production may be expressed as
f = Rmax tK~rT> %TTl:> C19D
where L is the sulfate concentration, C the soluble organic carbon
concentration, and K and K the Michaelis coefficients for L and
J_i C
C respectively. Bacterial populations are assumed to be relatively high
and stable.
Since sulfate and soluble organic carbon are being consumed in the
production of sulfide, their rate of utilization may be expressed as
5F
and
dt - ~'c W
The value for Y for use in the model was obtained by averaging the
L
measurements of this yield coefficient for cultures 1 through 12 (Table 8)
Y was similarly obtained, but by weighting most heavily those values
of Y from cultures which were not markedly sulfate limited during the
Li
77
-------
experiment. Measurements from cultures 13 and 14 were not included in the
average since these cultures were prepared with unautoclaved media.
Initial comparison of experimental results with calculated results of
equations 19, 20 and 21 demonstrated that production of sulfides responded
more sharply to changes in sulfate concentrations at low sulfate concentra-
tions than is described by the traditional Michaelis-Menton equation. This
sharper response could be expressed within the model by raising the sulfate
concentration in equation 8 to higher powers. Based on simulation of cultures
3 through 6, it was decided to raise the sulfate concentration to the 1.3
power, thus replacing equation 19 by
= Rmax
max
L
Estimates of R , K! and Kr were obtained using multiple nonlinear
max L L
regression analysis to fit the combined data of cultures 1 through 12 to
equation 22. Since the model assumed no lag in sulfide production, data
for the first one to two days of those cultures in which an obvious lag
occurred was not included in the analysis .
The result of the regression indicated a maximum rate of sulfide
production (R ) of 77 mg/1 day. The values of K' and K., were 320 mg/1 and
650 mg/1 respectively. A value of 3.1 was used for Y and 1.0 for Y . The
L 0
resulting equations with the fitted parameter estimates are shown below:
1. 3 _
> C23)
78
-------
- -1'0 <> C25)
Approximations to the simultaneous solutions of equation 23, 24 and
25 were obtained by digital computer using a fourth order Runge-Kutte finite
difference method. The simulations of cultures 1 through 12 are shown in
Figs. 13 through 24. Solid lines represent the values of sulfide, sulfate,
and soluble organic carbon calculated by the mathematical model. Circles are
the data taken from the respective cultures of the experiment. The values
of pH and Eh indicated are from the experiment.
Agreement between simulated and measured results are especially good for
those cultures in which the rates of sulfide production were highest. Agree-
ment was in general best for sulfide and sulfate, while the calculated con-
centrations of soluble organic carbon deviated more from the experimental
measurements. The selection of a Yr of 1.0 for the simulation model is
reflected in these results. The average value obtained for those experiments
in which sulfate was not limiting, 0.89, would have resulted in closer agree-
ment except where sulfate became limiting. Part of the difference between
the simulation and experimental results can be attributed to the lag period
which occurred in many of the cultures from day zero to one. It will be re-
called that the model assumes no such lag, and that data for this lag was not
included in the regression analysis. Excluding the lag period from Figs. 13
through 26 by considering day one the beginning of the experiment, would result
in a much closer agreement of simulated with measured results in most cultures.
79
-------
Equations 23, 24, and 25 were used to simulate production of sulfide
and consumption of sulfate and soluble organic carbon in cultures 13 and 14
(see Figs. 25 and 26. Simulated sulfide production was within 10 percent of
the actual production. Soluble organic carbon consumption was under-estimated
as with simulations of all cultures having a Y~ less than 1.0. A value of
\j
0.89 would have provided closer agreement.
80
-------
o>
E
1600
O i
K 1200
(9
(T
O
LU
-I
E
ui
t-
to
>
E
2 4 6 8 10 12
TIME(doys)
FIG. 13 - EXPERIMENTAL RESULTS OF CULTURE 1 COMPARED WITH EQUATIONS 23, 24, and 25.
81
-------
N
0>
E
1600-
S 1200
(E
O
LJ
_l
CO
800-
400 -
0>
Ł
LJ
O
800-
600-
400 -
200 -
0 2 4 6 8 10 12 2 4 6 8 10 12
TIME (days)
FIG. 14 - EXPERIMENTAL RESULTS OF CULTURE 2 COMPARED WITH EQUATIONS 23, 24, and 25.
82
-------
-1600
- 1200 ~
•s.
O>
E
- 800 <
v>
>
E
- 7
0 2 4 6 8 10 12
2 4 6 8 10 12
TIME(doys)
FIG. 15 - EXPERIMENTAL RESULTS OF CULTURE 3 COMPARED WITH EQUATIONS 23, 24, and 25.
83
-------
10 12
TIME(days)
FIG. 16 - EXPERIMENTAL RESULTS OF CULTURE 4 COMPARED WITH EQUATIONS 23, 24, and 25,
84
-------
1600
o
9 1200
o
o
z
Ill
m
o
w
800
400
E
LJ
150-
100
50-
1 1 1 1 T
1 1 1 1 h
o o
J L
I I L
024 6 8 10 12
1 1 1 1 T
-0 -O-
H—I—I—I—h
i i ' ' ' L
2 4 6 8 10 12
400
300 -
o>
E
LJ
200
E
TIME( days)
FIG. 17 - EXPERIMENTAL RESULTS OF CULTURE 5 COMPARED WITH EQUATIONS 23, 24, and 25.
85
-------
E 1500
o
m
o
z
o
o
m
o
w
1000 -
500-
E
111
o
ii_
en
0 2 4 6 8 10 12
8 10 12
TIME(days)
FIG. 18 - EXPERIMENTAL RESULTS OF CULTURE 6 COMPARED WITH EQUATIONS 23, 24, and 25.
86
-------
§
0:
<
O
O
(E
O
UJ
_l
CD
1600-
1200-
800-
400-
01
E
UJ
o
CO
800-
600-
400 -
200-
024
8 10 10
TIME (days)
FIG. 19 - EXPERIMENTAL RESULTS OF CULTURE 7 COMPARED WITH EQUATIONS 23, 24, and 25.
87
-------
O
ffl
CC
O
O
z
O
CC
O
UJ
CD
O
CO
1600
1200
800
400
N.
0>
8
UJ
O
en
800
600
400
200
-i 1 1 1 r
I i I
0 24 6 8 10 12
H 1 h
H 1 1 1 1 h
i i i
4000
3000 -
v
en
Ł
ui
2000 <
1000
100
0
-100
-200
>
E
UJ
2 4 6 8 10 12
TIME(mg/l )
FIG. 20 - EXPERIMENTAL RESULTS OF CULTURE 8 COMPARED WITH EQUATIONS 23, 24, and 25.
88
-------
o>
e
z
o
m
cc
o
cc
o
UJ
_J
m
o
V)
800-
600-
400-
200-
01
E
LJ
800-
600-
400-
200-
8 10 12
TIME(days)
FIG. 21 - EXPERIMENTAL RESULTS OF CULTURE 9 COMPARED WITH EQUATIONS 23, 24, and 25.
89
-------
400
1 '
CD
(E
< 300
o
o
z
<
O 200
cc.
o
00
o
en
100
o>
E
UJ
Q
(O
400
300
200
100
~i r
O
o
H 1 1 h
8 10 12
4000
3000 ~
ill
2000 <
u.
V)
1000
>
E
UJ
i
Q.
8 10 12
TIME(days)
FIG. 22 - EXPERIMENTAL RESULTS OF CULTURE 10 COMPARED WITH EQUATIONS 23, 24, and 25.
90
-------
- 200
O>
O
CD
L>
u
cc
O
UJ
_l
CD
O
CO
0>
E
ui
Q
CO
150
100
50
200
150
100
-\—I—I—I—I—h
I I I
0 2 4 6 8 10 12
H \ 1 \
2000
1500
1000
500
0>
UJ
CO
>
E
LJ
24 6 8 10 12
TIME(days)
FIG. 23 - EXPERIMENTAL RESULTS OF CULTURE 11 COMPARED WITH EQUATIONS 23, 24, and 25.
91
-------
0>
E
O
CD
K.
<
O
O
tr
o
bJ
_l
DQ
LU
Q
100 -
02 4 6 8 10 12
6 8 10 12
TIME(days)
FIG. 24 - EXPERIMENTAL RESULTS OF CULTURE 12 COMPARED WITH EQUATIONS 23, 24, and 25,
92
-------
o
m
a:
<
o
o
o
cc
o
CD
o
en
800
600
400
200
800
600
UJ
Q
;r 400
V)
200
1 T
H 1 h
1 1 1 1 T
-200
4000
3000 N
01
2000
1000
>
E
LJ
I
CL
2 4 6 8 10 12
TIME (days)
FIG. 25 - EXPERIMENTAL RESULTS OF CULTURE 13 COMPARED WITH EQUATIONS 23, 24, and 25.
93
-------
X
o>
E
8
o
o
a:
o
m
400
300
200
100
400
300
- 200
0)
100
1 1 r
-1 1 1 L
024 6 8 10 12
o
-i ' '
J L
2000
1500 -
o<
E
LJ
1000 <
500
100
0
-100
-200
UJ
X
a.
2 4 6 8 10 12
TIME( days)
FIG. 26 - EXPERIMENTAL RESULTS OF CULTURE 14 COMPARED WITH EQUATIONS 23, 24, and 25.
94
-------
SECTION X
SULFATE REDUCTION STUDY USING S-35
GENERAL APPROACH
In order- to better quantify the extent of sulfate reduction that
occurs within estuarine sediments a laboratory study employing S-35 was
conducted (61). The study involved the placement of a known quantity of
Na S 0 into homogenized mud samples. Samples were then incubated. At
selected time intervals, hydrogen sulfide was driven off as a gas and col-
lected. The amount of S-35 within the collected sulfide was then determined.
Thus, the percent of S-35 added as sulfate which was converted to sulfide
over a given time interval could be calculated . With the known initial sulfate
concentration of the sample, the rate of sulfate reduction could then be
computed. Benthic samples were collected at site 5 with 2-inch I.D. plexiglas
cylinders. Samples were transferred to the radiation laboratory at Corvallis,
Oregon in a styrofoam cooler.
SAMPLE PREPARATION
Immediately after the samples had been taken to the laboratory, they
were transferred to a portable plastic glove box containing a top loading
single pan balance; a large glass trough; conical glass incubators, each of
which has a side branch sealed with a tight-fitting rubber serum cap; and
other necessary apparatus for sample transference. The glove box was then
sealed and filled with pre-purified nitrogen which was being used to minimize
excessive air entrainment into the soil samples during the transferring processes,
The top 10 cm of the cores were extruded and homogenized in the glass trough.
95
-------
Five (5) gm of the homogeneous soil sample and two (2) ml of glass-distilled
water were carefully transferred into each tared incubator which was then
stoppered and removed from the glove box.
Since biological sulfate reduction requires strict anaerobiosis, care
was taken to remove all oxygen from the sample and the incubators used in
the study. The gas in each stoppered incubator was replaced with pre-purified
nitrogen by puncturing the rubber serum cap with a hypodermic needle connected
to a glass manifold. This manifold consisted of 6 hypodermic needles, a three-way
valve, and a U-tube mercury manometer, and was connected through the 3-way valve
to a vacuum line and to a supply of pre-purified nitrogen. The incubators
were alternately evacuated and filled with purified nitrogen. During this
process, a slightly reduced pressure was maintained in the incubators and the
glass manifold, as was indicated by the mercury manometer. Using this procedure,
intense reducing conditions were achieved in a relatively short period of
time (15).
A known amount of radioactivity, approximately 1 microcurie, in the
35
form of S-35 labeled sulfate solution (Na S 0.), was injected through the
serum cap into the sample. The contents of each incubator were mixed by
shaking vigorously for half an hour. The incubators were covered with
aluminum foil to reduce excess growth of photosynthetic bacteria and algae,
and were incubated at 18°C in a constant temperature room or water bath.
Cylindrical plexiglas vials of six inch long and four inch diameter were
initially used for incubation. Because these vials leaked and many other
difficulties were encountered during the addition of the radioactive sulfate
solution, modified wide-mouth erlenmeyer flasks fitted with ground glass
stoppers were used as incubators.
96
-------
INITIAL CONDITIONS
The initial sellable Sulfate concentration of the homogenized soil (prior
to the addition of 2 mis distilled water to 5 gms of sediment) was determined
by a colorimetric procedure using barium chloranilate (51). Approximately
30 gm of the soil sample were placed in an improved interstitial water sampler
(50). Interstitial water was squeezed out when compressed at 80-100 psi and
collected in a plastic hypodermic syringe. Then the water sample was filtered
to remove most of the suspended matter. After filtration, the sample was passed
through a Dowex 50 Wx-8 20-50 mesh H cation exchange column to remove inter-
ferring ions. Ten ml of this extract was mixed with 2.4 ml of barium chlorani-
late solution and 1.6 ml of acetate buffer solution. The unused barium
chloranilate and precipitated barium sulfate were filtered and 6.8 ml of the
pink filtrate were mixed with 0.46 ml of EDTA - NaOH solution. The transmittance
of the filtrate was determined at 520 my with a Beckman DB spectrophotometer.
Sulfate concentration was read from a standard calibration curve.
The soluble organic carbon concentration in the interstitial water of
the homogenized soil sample was determined. Five ml of the squeezed water
were placed into a 20 ml screw cap test tube in an ice bath. Samples were
tested in the Oregon State University Microbiology Department by using a
Lira Infrared Analyzer Model 3000.
Water content of the homogenized soil sample was determined by drying
the sample at 105°C.
COLLECTION OF SULFIDE
After incubation at 18°C for desired time periods, an incubator was
removed from the constant temperature room, and carefully connected to the
gas collection apparatus. The incubator was then lowered into a heated water
97
-------
bath. A small amount, several mg, of elemental mercury was added as a
catalyst (62) into the sample before acid digestion.
The digesting agent, initially concentrated hydrochloric acid, was added
into the incubator through a 250 ml pressurized separatory funnel connecting
to a supply of pre-purified nitrogen used as a carrier gas for evolved
hydrogen sulfide. Concentrated hydrochloric acid was later substituted by
hydriodic acid because concentrated hydrochloric acid was found to react
very slowly with several mineral sulfides, such as pyrite and chalcopyrite
(63).
The released hydrogen sulfide was carried by a stream of pre-purified
nitrogen through a train of gas wash bottles, each of which contained 100 ml
of 5N sodium hydroxide solution. The gases being emitted from the sampling
train were tested for any trace of hydrogen sulfide by using lead acetate
solution or a Kitagawa low concentration hydrogen sulfide detection tube.
After the complete decomposition of the soil sample, the apparatus was
continuously purged with pre-purified nitrogen for about an hour.
After acid digestion and the purging of the soil sample with nitrogen, the
pressure within each gas wash gottle was released by venting to the atmosphere
through the 3-way valves connected between the three gas wash bottles. The
valves were turned off in a reverse sequence; that is, the pressure in gas
bottle 2 was released before that of the first. Then the nitrogen supply
was turned off.
One ml of the sodium hydroxide-sodium sulfide mixture in each wash
bottle was injected into a liquid scintillation counter vial containing 18 ml
of "Aquasol" and 2 ml of distilled water. The activity of the sodium hydroxide-
sodium sulfide mixture was measured by a Tri-Carb Scintillation Counter with
98
-------
a photomultiplier voltage gain of 12.2% and window settings of 40 and 1000
respectively.
During the acid digestion process, condensate was formed and accumulated
in the connection between the incubator and the first gas wash bottle. A large
Y-connector had been used to join the incubator and the gas wash bottle. The
sample of the condensate was injected into a liquid scintillation counting
vial containing 20 ml of Aquasol. The condensate was found to contain some
radioactivity. Thus, not all of the S-35 released as hydrogen sulfide was
collected in the gas collection bottles. It was felt that this amount was
quite small, however, future studies should account for the amount.
It was observed that there was a possibility of contamination of the
sodium hydroxide solution in the first gas bottle. Therefore, it was
decided that any succeeding studies would employ a condenser connected at
the top of the incubator so that only relatively dried and cooled gases could
pass into the gas wash bottles.
Radioactive tracer assays, based on the liquid scintillation counting
method, depend on the optimization of the counting efficiency, which is
mainly influenced by two main factors:
(1) The figure of merit (64).
(2) The counting efficiency of the counter for S-35 isotope.
2
The figure of merit, (S /B) is the highest possible counting rate of
the specific isotope and the lowest possible counting rate of the background.
It was determined by continuous measurement of the net count rate of the
activity of a known standard S-35 solution(S) and of the background (B) while
periodically altering voltage gain of the photomultiplier. The optimum
condition was achieved when the voltage gain setting established a maximum
2
value for the expression S /B.
99
-------
In order to determine the efficiency of the counter for the activity
of S-35 isotope, the quenching effect had to be considered. The quenching
effect is any reduction of efficiency in the energy transfer process in a
given liquid scintillation counting sample, said reduction being caused by
color quenchers, chemical quenchers, and diluters. For the Tri-Carb
Scintillation counter, the quenching effect for each sample is determined
by an external standard, and the result is expressed.as Automatic External
Standard (AES) Ratio (64).
The counting efficiency, the ratio of the net photon count rate in
cpm to the disintegration rate (decay rate) in dpm, associated with each
quenching condition was determined by adding variable amounts of chloroforni,
a chemical quencher, to a series of samples containing 20 ml of Aquasol and
a known amount of radioactivity of S-35 (65). A quenching calibration
curve of the AES ratio versus the percentage counting efficiency was
plotted.
For each sample that was used in the study of microbial sulfate
reduction, the disintegration rate was determined with the quenching
calibration curve and an average value of the net count rate. The
disintegration rate was then converted to radioactivity in micro-curies
by dividing the dpm value with 2.22 X 10 . Since S-35 isotope has a half
life of 87 days, the experimentally determined radioactivity had to be
corrected using the equation for self disintegration:
N = N eyT (26)
where N = the measured activity in yc
N = the actual activity in yc if no
self disintegration exists
100
-------
X = decay constant = 0.00794/day
T = time in days since the addition
of S-35 activity to the soil sample.
The total amount of labeled sulfide generated microbially was determined
by multiplying the summation of the total actual activity by the total volume
of sodium hydroxide solution used to capture the hydrogen sulfide gas.
RATES OF SULFATE REDUCTION
The results of two concurrent experiments are summarized in Table 10.
The results show that sulfate reduction proceeded without a significant lag
time. The maximum rate of sulfate reduction was approximately 71 mg of
sulfide (S~) produced per day per liter of interstitial water of the original
sample. Recall, however, that 2 mis of distilled water were added to 5 gms
of wet sediment. Within this mixed slurry, the maximum rate would be approximately
46 mg/L-day. These results are remarkably close to the results obtained from
the extract experiments previously described. If a lag time did occur within
the first three day period, however, the maximum rate would have been higher.
The results indicate that the experiments were approaching ninety percent
recovery of the original S-35. The apparent lack of total recovery with
time may have been due to the experimental proceedure. As explained, S-35
did collect in the lines to the gas collection flasks. Physical adsorption
and biological incorporation may also have resulted in a delay in the
release of the remaining ten percent as hydrogen sulfide.
In addition to maximum rates of sulfate reduction comparable to the
extract experiments, the S-35 experiments also suggest that the rate of
sulfate reduction does not become sharply limited by the sulfate concentration
until sulfate concentrations fall below 200-300 mg/L. This observation is in
agreement with the results of the extract experiments previously described.
101
-------
TABLE 10. SUMMARY OF S-35 RECOVERY
Incubation
Time
(days]
3
6
9
14
21
28
Recovered
Run 1
37.86
70.02
77.8
79.36
85.06
85.29
S-35 (percent)
Run 2
40.98
68.46
76.24
80.39
83.5
83.71
Study conducted during summer 1971.
102
-------
SECTION XI
OTHER ESTIMATES OF SULFATE REDUCTION
DIFFUSION OF SULFATE
It is possible to obtain an estimate of the rate of sulfate reduction
within a deposit by considering the downward flux of SO . If the deposit
has remained undisturbed for a sufficient time, a steady state will be
approached which may be approximated by the equation
2
DL~Hr = A (2y)
in which D is the diffusion coefficient for sulfate (assumed to be constant
with depth), L is the sulfate concentration, z is the depth from the sedi-
ment surface and A is the mass of sulfate removed per unit time per unit
volume of interstitial water. Assume that A is constant with depth until L
becomes limiting at concentration L1. Below the limiting depth (the depth
at which L=L'), A will be equal to zero. The following equation is then
obtained from equation (27)
2DT(L -L1)
A = L ° (28)
in which L is the concentration of sulfate of the sediment surface and z'
o
is the limiting depth at which L=L'. The maximum value of LQ will be
2655 mg/L (the sulfate concentration in sea water). The solution of equation
(28) for values which might be common with estuarine sediments is graphically
-5 2
shown in Fig. 27. A reasonable (9) value for DL(.5 x 10 cm /sec) is shown
within a range (10"6cm2/sec to 10"5 cm /sec) that might be expected.
103
-------
>.
-8
3
300
— 200
LJ
o:
o
Q
LJ
o:
UJ
C/)
100
D-IO'5cmz/sec
D=0.5 xlO'5cm2/sec
D=/0-6cm2/sec
0 12345678
DEPTH TO SULFATE LIMITATION (z')-cm
FIG. 27 - RELATIONSHIP OF SULFATE REDUCTION RATE TO SULFATE DIFFUSION ASSUMING STEADY STATi
104
-------
A number of sulfate profiles within the sediments of tidal flat areas
were measured and more than twenty such profiles were obtained. Most did
not appear to be approaching the steady state described above. That is, the
variation of L with depth similar to the solution of equation (27) was not
clearly defined. For most profiles, scour and partial turnover appeared to
have influenced sulfate profiles. The profile shown in Fig. 10 does indicate
an z1 value of approximately 3 cm. It is difficult to define the sulfate
profile within a depth of 3 cm. Moreover, the uneveness of the mud surface,
the looseness of the deposit surface and the influence of partial scour should
be recognized. Thus the use of equation (28) can only be expected to provide
an order of magnitude estimate of A. Using the values, z'=3 cm, L -L' =
-5 2
1500 mg/L and DT = 0.5x10 cm /sec, one obtains an approximate value of A
Li
equal to 150 mg/L-day. This corresponds to a rate of sulfide production
of 50 mg/L-day (as S) which falls within the range of the experimental results
previously discussed.
INCUBATION OF BENTHIC CORES
Estimates of the rate of sulfate reduction were obtained by collecting a
number of sediment cores at a given site in 2 inch I.D. plexiglass tubing.
Cores were incubated at room temperature (approximately 25°C) and profiles of
sulfate concentrations within the interstitial water of cores were determined
at time intervals of 1 to 6 days. Rates of sulfate reduction were calculated
by determining the differences of sulfate concentrations between cores
incubated for different time periods. No corrections for the diffusion of
sulfate were made. Effort was made to collect cores from areas where the
deposits appeared to be horizontally uniform. Variations between cores from
which the rates were determined, however, did result in considerable scatter
of results. A summary of the rates so determined are given in Table 11.
105
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TABLE 11. SULFATE REDUCTION IN CORES AT ROOM TEMPERATURE C25°C)
Starting
Date
10/ 1/71
10/18/71
10/18/71
10/ 2/71
10/ 7/71
(a) (b)
(a)
Ca)
(c)
(c) Cd)
Site
4
4
4
5
5
Sediment
Temp.
(•c)
--
13
13
18
15
Volatile Solid
(%-dry weight)
13-16
13-20
15
Initial SO.
(rog/D
1000
1100
700
1000
800
SO. Reduction
A
(rag/ 1 -day)
50
40
30
100
70
S Production
(mg/l-day)
17
13
10
33
23
(a) Values averaged over top 4 cm of core
(b) Soluble organic carbon = 15-130 mg/1
(c) Values averaged over top 3 cm of core
Cd) Soluble organic carbon = 33-215 mg/1
-------
SUMMARY OF SULFATE REDUCTION RATES
The rates of sulfate reduction measured during this study are compared
to those reported by other investigators (Table 12). Maximum rates in the
study by Edwards (60) were higher than those measured in this study. The
higher incubation temperature (30°C) , and use of sufficient lactate (a
completely utilizable carbon source) to produce soluble organic concentrations
above those in the extract's media (Section IX) would likely account for these
higher rates. The maximum rates measured by Nakai and Jensen (66) were within the
range of those reported measured in this study. Measurements by Ivanov (20)
and Sorokin (21) of sulfate reduction in lake muds were determined through the
35
use of labeled sodium sulfate (NA S 0.). The difference in units used to
report their rates makes comparison difficult. If it is assumed that their
mud samples were roughly 50 to 75 percent water, then the reported rates
would range up to approximately 40 mg sulfide per liter per day. In their
study, 10 cm mud samples were utilized and the rates reported based on pro-
duction of sulfide over this depth. If the production were occurring within
only the top few cm, however, then the rates reported might underestimate the
actual production occurring within this active upper region. Accounting for
these factors would produce approximate agreement with the range of those
measured in this study. None of the studies shown in Table 12 provide
estimates of the rate of sulfate reduction that might occur immediately below
the mud surface. Within the upper anaerobic regions of high organic deposits,
rates higher than those shown in Table 12 might be possible particularly
within the regions immediately under dense algal mats.
107
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TABLE 12. COMPARISON OF THE RATES OF SULFIDE PRODUCTION MEASURED
IN THIS STUDY WITH THOSE OF PREVIOUS INVESTIGATORS
Investigator
Sulfide Production
rates (a)
Comments
This study
This study
This study
This study
10-701
46
50(
10-23(
00 C<0
Sediment Extracts, See Section IX
Mud Slurry, See Section X
In situ sulfate profiles, order of
magnitude only, see Section XI
Incubation of cores, limited study,
see Section XI
Edwards (65)
Edwards (65)
Ivanov (20)
Ivanov (20)
Nakai and Jensen (66)
Sorokin (21)
Sorokin (21)
200-250 ^ J
100-150^
0.5-1.5^
12-19^
10-45^
0.1-0.2Cb)
10-15^
In lab, pure batch cultures of D.
desulfuricans on Macpherson's
medium, stable populations
In lab, pure batch cultures of D.
desulfuricans on Macpherson's
medium, stable populations
Field measurements in 10 cm mud
cores from deepest part of lake,
determined with 3^5
Field measurements in 10 cm mud
cores from slope of lake,
determined with S-^
In lab, mixed cultures containing
sulfate reducing bacteria, cultures
consisted of 30 ml sea water and
65 ml wet sediment
Field measurements in lake water
using S-55
Field measurements in muds collected
from slope of lake near river mouth,
S35 used
(a) Approximate range
Cb) mg(S)/l- day
(c) Maximum rate
(d) mg(S)/ Kg wet sediment-day
108
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SECTION XII
BENTHIC SULFIDE RELEASE
GENERAL
Attempts were made to measure the benthic sulfide release at site 5
through the use of a benthic respirometer. A single successful respirometer
run was completed during the summer of 1970. A sulfide release rate of 1.6 gm
2
gm/m -day was measured in a 2-1/2 hour run without correction for oxidation
2
of free sulfide. An oxygen depletion rate of 3.2 gm/m -day was also measured.
If approximately half of this oxygen uptake rate was due to the oxidation
of free sulfides, then the total sulfide release rate may have approached
2
3.2 gm/m -day. The respirometer was later modified to permit nitrogen gas
sparging of DO from the installed respirometer. Three additional respirometer
runs were repeated at site 5 during the summer of 1971. (67)
MODIFIED BENTHIC RESPIROMETER
The modified benthal respirometer (similar to that described in Section VI)
consisted of a black plexiglas tunnel, a small submersible pump, an expansion
chamber, a system of hoses, a flow indicator, a sampling port, and a sparging
unit. The tunnel was 2-7/8 inches in diameter, 49 inches long, with 2-inch
flanges attached to the base, and covered an area of 282 square inches (0.182
square meters). It was painted black to prevent photosynthetic oxygen genera-
tion by algae covering the benthal deposit. The tunnel was placed on the mud
at low tide and filled by the rising tide through holes in the top which are
then sealed by rubber stoppers. Water was circulated at a rate of about 4 gallons
per minute through the device in order to thoroughly mix the contents and
109
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facilitate sampling at the estuary's water surface. Water was circulated
through a 5/8-inch diameter garden hose. A strip of plastic cloth was
installed to flutter inside a three-inch length of clear plastic tubing to
indicate that flow was being maintained. Sampling was done with a 30 ml syringe
through a septum installed in a plastic "tee" located on the circulation hose.
Duplicate samples were taken every twenty minutes for DO and free sulfide
measurement by methods described in the previous section.
At the onset of the 1971 runs, nitrogen gas was sparged into the pressure
line of the pump, collected in the expansion chamber, and relieved through a
valve in that chamber. This chamber, made from a 27-liter plastic carboy, had
inlet and outlet fittings and an adjustable relief valve. Nitrogen sparging
was intended to strip the DO from the water, and thus permit the buildup of
free sulfides without oxidation.
Care had to be used during sparging to maintain the same volume of water
within the device by adjusting the release rate to equal the sparging rate.
Constant volume was necessary to facilitate computation of the mass of sulfide
released from the observed rise of sulfide concentration. This was done with
the aid of a line on the expansion chamber to indicate a constant gas volume.
The total volume of water within the device, with one liter of gas in the
expansion chamber, was 38 liters. Excess pressure had to be avoided to pre-
vent pushing the tunnel off the mud, as happened in the earlier runs.
RESULTS OF SULFIDE RELEASE MEASUREMENTS
The results of the 1971 runs are presented in Figs. 28, 29 and 30. The
release rate of sulfide in each experiment was expressed by a high and low
rate shown by the lines in these figures and in Table 13. When DO was present,
the sulfide release rate increased with time. When nitrogen sparging removed
the DO, the sulfide release rate decreased with time. A number of explanations
110
-------
1200 1300 1400
TIME-hrs
• = free sulfide
o = dissolved oxygen
FIG. 28 - BENTHIC SULFIDE RELEASE WITHIN RESPIROMETER - SITE 5 - 8/3/71.
1500 1600 1700
TIME-hrs
1800
E
I
•z.
LJ
CD
O
O
UJ
CO
CO
0
• = free sulfide
o = dissolved oxygen
FIG. 29 - BENTHIC SULFIDE RELEASE WITHIN RESPIROMETER
Nitrogen sparging shown by arrow).
- SITE 5 - 8/9/71. (Period of
111
-------
4.0
E 3.0
i
in
LJ
Q
^j 2.0
r)
1/5
1.0
0
1300 1400 1500 1600
TIME-hrs
• = free sulfide
o = dissolved oxygen
FIG. 30 - BENTHIC SULFIDE RELEASE WITHIN RESPIROMETER - SITE 5 - 8/20/71. (Period of
Nitrogen sparging shown by arrow.
112
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TABLE 13. ESTIMATES OF SULFIDE RELEASE RATES AND OXYGEN UPTAKE
RATES AS MEASURED IN RESPIROMETER EXPERIMENTS AT SITE 5
Date
9/ 3/70
9/ 3/71
9/19/71
9/20/71
2
S Release Rate (gm/m -day)
1.6- 3.2Ca>
1.0- 9.2
0.8- 7.2
1.0-15.0
2
DO Uptake Rate (gm/m -day)
3.2
-
15-39^
-
(a) Upper estimate based on assumption one half DO uptake due to
sulfide oxidation with one to one sulfide oxygen mass ratio.
(b) High rates may be due to interference of DO measurements.
are possible and the data are not adequate to select from such explanations.
The initially high rate during the 8/9/71 and 8/20/71 runs may have resulted
from the disturbance of the bottom by the respirometer. In these same runs,
the decline in the sulfide release rate may have resulted from the input of
Fe and the formation of FeS. The termination of nitrogen sparging may also
have contributed to the decline in sulfide release. The high DO decrease in
the 8/3/71 run may be due to sulfide interference of the micro-winkler test.
The difficulties in performing these respirometer runs should not be
underestimated and the results must be interpreted with these difficulties in
mind. The results including the 1970 run do indicate that the sulfide release
rate in this region during late summer likely falls in the approximate range
of 1.0 to 10.0 mg S~/m -day. Any finer evaluation of the results is likely
unwarranted. More reliable sulfide release rates would likely be obtained
through the use of laboratory flow-through systems. The benthic sulfide
113
-------
release rates which are possible within a range of measured sulfate reduc-
tion rates are shown in Table 14. These rates are maximum values as it is
assumed that all sulfide produced is released.
PROFILES OF FREE SULFIDE
Free sulfide concentrations were measured within the interstitial waters
of sediment cores. In general, gradients of free sulfides seldom exceeded
40-50 mg/l-cm (see Fig. 11) and this value was approached only at site 5.
If the release of free sulfides depended on molecular diffusion only, such
2
gradients would result in a sulfide release rate of approximately 0.2 gm/m -day.
Maximum free sulfide gradients within the immediate surface regions of the
deposits, however, were likely higher than those measured due to the difficulty
of sampling interstitial waters within small distances. Moreover, partial
turnover of the immediate surface regions can be expected and thus the upward
2
flux of free sulfide would be increased. Thus the value of 0.2 gm/m -day is
likely a low estimate of the higher benthic sulfide release rates attained
at site 5.
SUMMARY OF BENTHIC SULFIDE RELEASE EXPERIMENTS
Measurement of benthic sulfide release likely occurred for any prolonged
period only at site 5. This site had all of the conditions which would contribute
to maximum sulfide releases. Experimental results were extremely variable and
2
one can only conclude that sulfide release rates approximating 1.0 gm/m -day
should not be considered unreasonable for conditions similar to those of site 5.
This estimate is extremely crude, however, given the very limited and variable
experimental results. Moreover, site 5 is not typical of estuarine tidal flats
and the average sulfide release rates for typical tidal flats can be expected to
be considerably lower than those experienced at site 5.
114
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TABLE 14. POTENTIAL BENTHIC SULFIDE RELEASE RATES
(GMS/M2-DAY) (a)(b)
W a?
4H 13
O 1
rH
•P M
5 e
B
O O
M-rt
Oi +J
fH O
(D 3
> TJ
< o
(X
20
40
60
80
100
2
0.2
0.4
0.6
0.8
1.0
Depth of
3
0.3
0.6
0.9
1.2
1.5
Sulfate Reduction-cm
4
0.4
0.8
1.2
1.6
2.0
5 6
0.5 0.6
1.0 1.2
1.5 1.8
2.0 2.4
2.5 3.0
(a) Assume 50% water content by volume
(b) Assume all sulfide produced is released
115
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SECTION XIII
MIXING WITHIN DEPOSITS
GENERAL
The vertical mixing that occurs within estuarine benthic systems is an
important factor in determining the conditions within the benthic system and
the influences of the benthic systems upon the overlying water and air.
Vertical mixing within deposits depends on a variety of factors. Hydraulic
factors, such as tidal changes in water depth, water velocities, wave action
and low tide drainage patterns all contribute toward vertical mixing (68)(69).
The burrowing and movement of organisms within deposits leads to greater vertical
mixing (38)(70)(71). Dye tests at site 2 indicated a large exchange of water
through mud-shrimp burrows (34). The presence of fine particles within the
deposits tends to reduce vertical mixing (68)(69). Ebbing waters were found
to drain freely through the deposits at site 1 while a lack of such drainage
was noted at station 5. Water velocities measured at sites 3 and 4 with a price
current meter were generally below 0.6 fps.
The presence of biological growth may retard the passage of water through
the deposits and thus reduce the hydraulic tidal mixing. At site 1, permeabili-
ties of 0.04-0.03 cm/min were measured at locations where biological growths
were not noted while, in this same general region, the permeability in the
top twelve centimeters was reduced to 0.0008 cm/min in areas of noticeable
biological growth on the deposit surface.
In regions where vertical mixing is reduced to that of molecular diffusion,
the departure from straight-line diffusion, due to the deposit particles,
reduces molecular diffusion coefficients by approximately 30 percent from that
116
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of pure water (72)(73). Electric interaction between particles and inter-
stitial water can also lead to a reduction of molecular diffusion (74).
Diffusion within biological slimes can also be significantly less than diffusion
through pure water (75) . Vertical mixing can thus be expected to vary from
relatively large hydraulic exchanges, often facilitated by burrowing organisms
and large particle sizes (site 1), to values less than that of molecular
diffusion in pure water. Vertical mixing can thus be expected to vary from
relatively large hydraulic exchanges, often facilitated by burrowing organisms
and large particle sizes (site 1), to values less than that of molecular
diffusion in pure water (possibly the case at sites 4 and 5).
Care must be exercised in the use of reported molecular diffusion
coefficients of substances within interstitial water, as the concentration
of sorbed material which does not diffuse is often included with the material
in solution (74). Concentration gradients based on the sum of sorbed and
nonsorbed materials will lead to lower calculated diffusion coefficients than
if only the material in solution were used to determine the gradients.
TIDAL MIXTURE
During August and September of 1970, a series of experiments were con-
ducted at sites 1 and 4 in order to study the discharge and recharge process
of tidal flat pore water during a tidal cycle (76).
During the first series of experiments at site 1, measurements of the
hydraulic surface within the flats were obtained by means of pipes driven
into the deposits. It was found, however, that the hydraulic response of the
system was too slow. During later investigations, the hydraulic surface was
measured through dug holes with the surface referenced to stakes of known elevation.
117
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Changes of the hydraulic surface for three stations located at site 1 are
shown in Fig. 31. The station elevations are shown on Fig. 32 and the free
water surface elevations, which were measured during the sampling period
shown on Fig. 33.
These results suggest that during the ebbing of the tide, pore water
drains from those areas exposed by the tide. Drainage is along the beach,
passing through the surface at the region above the free water surface. The
slope of the hydraulic surface at approximately 2 hours after low tide is shown
on Fig. 31. As seen in this diagram, the hydraulic surface meets the ground
surface between stations C and D. Similar hydraulic surface profiles can be
drawn for different times within the tidal cycle from the data given in Fig. 30.
Thus, the data show that the flow of pore water during the ebbing tide and during
the low tide period results in a surface flow component at different elevations
at different periods of the tidal cycle. The magnitude of this surface flow
component depends on the permeabilities found at each location (see Table 15).
During the tidal cycle, the hydraulic surface within the flats dropped by
approximately 1.5 to 0.5 feet, the larger drops being measured at the higher
elevations. Above the hydraulic surface, the water content was reduced to
approximately 90 percent saturation. These percents, shown in Fig. 31, should
only be considered as rough approximations due to the difficulty of obtaining
undisturbed one-inch thick samples.
Using the mixing length description of the diffusion coefficient, the
data presented above can be utilized to compute rough estimates of effective
diffusion coefficients due to the tidal pore drainage described above. These
effective diffusion coefficients are approximately 10 to 50 times greater than
molecular diffusion coefficients (for oxygen) in water. In addition, oxygen may
be introduced in large amounts to the unsaturated region.
118
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I2
5
o
\b Sept (970
Time. - Clock Hours
17 Sept
FIG. 31 - HYDRAULIC SURFACE DURING EXPERIMENTS AT SITE 1,
-2
d: cfeptt ofsffaf/e- cm. i -> --><"
_ n: Wuaie. i/o/as / /o/a/ volume.
ZOO
Distance. • feet
30O
40O
FIG. 32 - CONDITIONS AT SITE 1 DURING DRAINAGE EXPERIMENT.
119
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Ground Surface. - S fat/Off A"
Ground Surface. -Station &
Hydraulic Surface.
Station
Ground Surface • station 'C'
16 Sept 1970
Time. • Clock Hoars
17 Sept
FIG. 33 - FREE WATER ELEVATIONS DURING EXPERIMENTS AT SITE 1.
120
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TABLE 15. PERMEABILITY AND VOID RATIO AT SITE 1
Elevation
[ft above MLLW)
8.3
6.8
6.0
6.0
3.8
1.7
-1.0
Permeability
Ccm/min)
0.04
0.0008^
—
—
—
—
0.03
Vol. Voids/Total Vol.
0.38
—
0.36
0.38
0.44
0.44
0.42
(a) Plant growth on surface
At site 4, no similar tidal drainage could be measured. Water would
remain within pools (with some evaporation) until the flood tide covered
flats. Vertical mixing was thus reduced to molecular diffusion and the turnover
by organisms within the deposit. This later method did not appear to be sub-
stantial at site 4. Partial scour of the region was also noted, due to wave
action as the tide rose and fell.
The characteristics of the deposits at sites 1 and 4 were quite different.
The deposits at site 1 contained a reasonably uniform sand while large amounts
of silts and clays were present at site 4. It appears that the difference in
particle size was chiefly responsible for the differences in drainage character-
istics and thus the differences in vertical mixing.
121
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During the flood tide at site 1, the hydraulic surface began to rise
slowly as the free surface approached a given elevation (see Fig. 31). This
rise became abrupt at each location as the flooding waters covered that
location. A more detailed description of this study has been presented (76).
PERIODIC SCOUR
Though time was not available to maintain a sufficient sampling program
to define seasonal changes of conditions with the benthic systems of tidal flats
the data available do suggest that periodic scour of the deposits, generally to
a depth of 5-10 cm, is an important factor in determining these conditions.
The scour due to wave action, high water velocities and extreme tides can
result in a periodic oxidation of ferrous sulfide and a "recycling" of the
iron. Such physical disruptions tend to prevent the depletion of available
iron and the subsequent buildup and release of free sulfide. The influence
of oxidizing conditions upon the concentrations of free sulfides is illustrated
in Fig. 34 which is a composite of results from sites 1,2,4 and 5. The
sulfates brought into the deposit as a result of a physical turnover may not
provide for sufficient sulfate reduction to deplete the available iron.
Depletion of available iron may thus require the additional diffusive transport
of sulfate from the surface either through molecular diffusion or through partial
(or limited) scour. Thus, the buildup of free sulfide is most likely within the
top 5-10 cm. Periodic turnover of this region sufficiently to oxidize the
ferrous sulfide, however, would prevent this buildup. The influence of such
physical turnover is indicated by the progressive change in the total (acid soluble]
sulfides at site 4 as shown in Fig. 35 (free sulfide concentrations were below
detectable levels). The change appears to be most significant above 10 cm.
The decline of total sulfides as winter is approached reflects the declining
122
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3000-
I
Ld
Q
CO
2000
1000
• • /
*.rv .''.^i
-200 -100 0 +100 +200 +300
REDOX POTENTIAL - mV
FIG. 34 - RELATIONSHIP BETWEEN TOTAL SULFIDES (ACID SOLUBLE) AND REDOX POTENTIAL WITHIN
DEPOSITS AT SITES 1, 2, 4, and 5.
500 1000 1500
TOTAL SULFIDE-mg/kg
FIG. 35 - TEMPORAL CHANGE OF TOTAL SULFIDE (ACID SOLUBLE) PROFILES AT SITE 4.
123
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salinity (and thus sulfates) during this period at this site. The trend may
also reflect and increase wave action due to storms during this period.
A limited or partial scour or turnover of the upper layers of deposits
may not be sufficient to result in any substantial oxidation of the ferrous
sulfide. However, such limited scour can serve to transport sulfate into the
deposit. A partial turnover which entrained overlying water into the sediments
would normally contain 4-10 mg/1 of dissolved oxygen yet the sulfate concen-
tration could be 2000 mg/1 or greater. Thus limited turnover could contribute
to a depletion of available iron and a buildup of free sulfide (generally withir
the top 10 cm) if sufficient organics were available. Such limited scour in
the absence of a more complete scour would be most likely to occur within
protected areas (sloughs, diked areas, etc.) with relatively unconsolidated
sediments of high organic content.
124
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SECTION XIV
GENERAL BENTHIC DEPOSIT MODEL
GENERAL
As discussed in the introduction of this report, the basic approach
employed in this study involved two distinctive features. These features
were:
1. A complementary feedback occurred between the mathematical model
studies and the experimental portions of the project.
2. The study was approached at several levels of resolution which
complimented each other. That is sub-systems were viewed as
integrated systems with component parts and these same sub-systems
were also viewed as component parts of a higher level system.
The objective of this section of the report is to provide a benthic system
model in the language of mathematics. The model will serve to integrate the
principal processes shown in Fig. 2 and provide a framework for integrating the
results previously discussed into the general benthic system. The specific
mathematical description of each of these component processes will not, however,
be given. Though the previous sections of this report do give insight into
some of these component processes, particularly sulfate reduction, further
definition must be based on continued feedback with experimental results.
The model will be developed by first stating the principal ass-umptions.
Then, a description of the general distribution in space and time of soluble
and insoluble materials within benthic systems will be provided. The major
biochemical reactions which occur within benthic systems will then be presented
in two slightly different models. The intent of presenting the second model,
125
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which is a simplified version of the first biochemical model, is to show how
some simplifications can be made without necessarily sacrificing too much from
the capabilities of the model. Finally, the distribution equations will be
combined with the biochemical models to obtain two slightly different models
which are relevant to environmental studies of benthic systems. The models
are developed only for the benthic systems and do not here include the over-
lying air and water. The system described, however, is nearly applicable to
the overlying water.
PRINCIPAL ASSUMPTIONS
Several principal assumptions and limitations of the model are listed
below:
1. The model will be one-dimensional; considering only variations in
the vertical direction.
2. The model will assume an equilibrium between sorbed and nonsorbed
materials.
3. Compaction of deposits will not be included in the model.
4. It will be assumed that the biological reactions are not limited by
the concentrations of the microorganisms.
5. Deposition and scour will not be explicitly described in the model.
However, the model description is such that numerical simulations
based on the model will be able to accommodate a wide variety of
deposition and scour patterns.
6. As in all models, many different substances will be grouped into
large categories. As an example, degradable organics will be
grouped into two categories (insoluble and soluble). If necessary,
these groupings may be broken down further.
GENERAL DESCRIPTION OF SOLUBLE MATERIALS
Consider a vertical section of deposit of cross-sectional area A with a
fixed reference (z=0) at some depth below the surface of the deposit. Positive
126
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depth, z, is measured from this reference toward the surface. A positive
interstitial water velocity is in the upward direction and a positive con-
centration gradient indicates an increasing concentration from the reference
to the surface. Taking a mass balance on a slice of depth dz within the deposit
leads to
AdzZH (29)
^
3t
in which n is the fraction of the deposit filled with water, S' is the
concentration of a soluble material, within the interstitial water, F. is
the total flux of this substance into the slice, F is the flux of the
substance out of the slice, ZG is the sum of the sources and sinks of the
soluble material from within the interstitial water and EH is the sum of the sources
and sinks of the soluble substance from insoluble states. Hereafter the
prefix, G, for biochemical reactions will denote a rate of mass change or trans-
fer per unit volume of interstitial water. The prefix, H, will denote a
rate of mass change or transfer per unit volume of wet sediment.
The flux of S' due to advection and diffusion is taken as
P\C T
F = -D ,Ant2_ + UAmS' (30)
S o Z
in which D , is the vertical diffusion coefficient for S', U is the
vertical velocity (positive upward) and m is the fraction of A open to dif-
fusion and advection. Let
F = F. + dz (31)
o i 9z
127
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Substituting equations (30) and (31) into equation (29) leads to
, ^c^
3t 3z 3Z
Assuming A to be constant with distance and time and n to be constant with
time reduces equation (32) to
3(D ,m(3S'/3z) ,
I - s' -- 1 + ZG + i (33)
n 3z n 8z n
If D , , U, n, and m are further considered as constant with depth and
if n is set equal to m, equation (33) reduces to
For simplicity, equation (34) will be expanded to include the biochemical
reactions rather than the more general equation (33). The biochemical reactions
presented herein can be easily be accommodated into equation (33) if desired.
GENERAL EQUATION OF INSOLUBLE MATERIALS
Following a similar mass balance as above, one obtains the general
equation for insoluble materials shown below
|^- = EH H- nEG C35)
in which I' is the concentration (mass per unit volume of wet sediment) of
the insoluble materials.
BIOCHEMICAL MODEL I
The primary biochemical reactions occurring within the estuarine benthic
systems (see Fig. 2 and Section IV) are shown in the simplified mass transfer
diagram of Fig. 36. Definition of terras is given in Table 16.
128
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6F
YFBHFS.
INSOLUBLE
AVAILABLE
IRON
(IF)
i
HIF
r
1
-E
_E
6S^
INSOLUBLE
SULFIDE
(FS)
Nt«,
/SOLUBLE
AVAILABLE]
IRON
HO
ELEMENTAL
SULFUR
(SU)
FIG. 36 - BIOCHEMICAL MODEL I FOR ESTUARINE BENTHIC SYSTEM.
(Available Iron also includes other metals which
form insoluble sulfides; Zn, Sn, Cd, Hg and Cu.)
129
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TABLE 16. DEFINITION OF TERMS FOR
BIOCHEMICAL MODEL I OF ESTUARINE BENTHIC SYSTEM
GC, = the aerobic biochemical degradation of soluble organics,
GC = the degradation of soluble organics by sulfate reduction,
GC = the non sulfate reduction anaerobic biochemical degradation of soluble organics,
O
GF = the oxidation of soluble available iron to insoluble available iron,
GS = the oxidation of free sulfides (primarily, chemical reaction with dissolved
1 oxygen but also due to aerobic autotrophic bacteria),
GS = the reaction of free sulfides with soluble available iron (and other materials
which form insoluble sulfides),
GS = the loss of free sulfide through photosynthetic anaerobic sulfur bacteria,
HFS = the oxidation of insoluble sulfides (primarily ferrous sulfide) with dissolved
oxygen,
HFS = the oxidation of insoluble ferrous sulfides with elemental sulfur,
HIC, = the solubilization of organics under anaerobic conditions,
HIC = the solubilization of organics under aerobic conditions,
HIF = the solubilization of available iron,
HO = the dissolved oxygen demand due to other causes (e.g., oxidation of ammonia,
benthic plant respiration),
HP = the oxidation of pyrite under aerobic conditions,
YCI = mass of soluble organic carbon per mass of insoluble organic carbon solubilized
under aerobic conditions,
mass of sulfide produced per mass of organic carbon utilized in non-sulfate
reduction anaerobic decomposition,
Ypg = mass of insoluble available iron per mass of insoluble sulfide oxidized,
YFS = mass °f available iron Per mass of free sulfide reacted to form insoluble sulfides,
YFSU = mass of elemental sulfur used per mass of insoluble sulfide transformed to pyrite,
YIC = mass oŁ soluble organic carbon per mass of insoluble organic carbon solubilized
under anaerobic conditions,
\C = maSS °f sulfate Per mass of soluble organic carbon utilized in sulfate reduction,
YOC = mass of oxygen Per mass of soluble organic carbon oxidized,
130
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TABLE 16. Cont.
Y = mass of oxygen per mass of soluble iron oxidized,
OF
Y = mass of oxygen per mass of insoluble sulfide oxidized,
OFS
Y = mass of oxygen utilized per mass of insoluble organic carbon solubilized,
01
Y = mass of oxygen per mass of free sulfide oxidized,
Y = mass of pyrite formed per mass of insoluble sulfide reacted with elemental sulfur,
P
Y = mass of sulfate produced per mass of soluble organic carbon utilized in sulfate
SC reduction,
Y = mass of elemental sulfur per mass of insoluble sulfided aerobically oxidized,
O\J
Y = mass of elemental sulfur per mass of free sulfide oxidized.
131
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Insoluble materials are shown in the rectangles while soluble materials
are shown in the circles of Fig. 36. Symbols used to designate concentrations
of materials are given in Fig. 36. Concentrations of soluble materials are
given in mass per unit volume of interstitial water while concentrations of
insoluble materials are given in mass per unit volume of wet sediment. A
principal assumption which led from Fig. 2 to Fig. 36 is that the biochemical
reactions are not limited by the concentrations of micro-organisms.
The relative significance of the different reactions shown in Fig. 36 and
Table 16 will depend on the conditions within the benthic system (See Chapter IV).
Three benthic conditions and the relative importance of the reactions within
these conditions are given in Table 17. The aerobic regions, (portions of
the benthic system containing dissolved oxygen) are generally located within
the top few millimeters of the deposit. Such regions also frequently line
burrows within deposits. Aerobic conditions are also found for short periods
of time after the deposit has been physically overturned. If organic material
is sufficient, an anaerobic region within which sulfate reduction takes place
is normally located below the aerobic region. If organics are high, sulfate
may become limiting within several centimeters of the surface.
Microbial decomposition may occur below the depth of the sulfate limitation.
Free sulfides may diffuse from above and react with available iron in this
region. In addition, organics and iron may be solubilized in this region
and diffuse upward into the region of sulfate reduction. Methane fermentation
within this deep region may result in a physical disruption of the deposit
above.
The biochemical reactions shown in Fig. 36 and Table 16 may be combined
with equations (34) and (35") to provide the following mathematical description
of the benthic system.
132
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TABLE 17. DEPENDENCE OF REACTIONS OF BIOCHEMICAL MODEL I
ON INTERNAL BENTHIC CONDITIONS
Reaction
Aerobic Region
Anaerobic-Sulfate
Reduction Region
Anaerobic, No Sulfate
Reduction Region
HICj
HIC2
GCj
GC
GCs
HO
GS1
GS2
GS3
GF
HIF
HFS,
HFS2
HP
none
significant
significant
none
none
significant
significant (g)
very slight
none
significant
slight
significant (g) (k)
none
slight (j) (1)
significant
none
none
significant
slight (e)
none
none
significant
slight Ci)
none
significant
none
significant (d) (j)
none
significant (d)
none
none
none
significant (d) (f)
none
none
conditional (h)
unlikely (i)
none
significant
none
normally slight (d) (j)
none
(a) DO^-0.1 mg/1; upper region of sediment
(b) DO <0.1 mg/1; S04> 10-50 mg/1; below aerobic region
(c) D0<0.1 mg/1; SO, < 10-50 mg/1; below sulfate reduction
region
(d) likely decreasing with depth
(e) normally YgCGC2> YCgGC3
(f) if conditions favorable (e.g., available degradable
organics, no loxic levels of free sulfides)
(g) normally small suflate source in comparison to
diffusive transport and exhange due to disturbance
of bottom
(h) largely dependent upon diffusion of
free sulfides from sulfate reduction
region
J^i) light needed
(j) long term significance; time scale of
weeks to years
(k) likely to occur after sediment over-
turned or flushed with oxygenated
water
(1) not included as oxygen loss due to
relative slowness or reaction
-------
Insoluble carbon:
- HIC
(36)
Soluble carbon:
.
dZ
+ Y HIC
1 Li
Dissolved oxygen:
(37)
= °
0
d Z
YOCGC! - YOSGS1 - YOFGF
(38)
Free sulfide:
Pjc
If = °S
3z
YSCGC2
YCSGC3 -
GS2 - GS3
Sulfate:
f
(40)
134
-------
Insoluble iron (and other material which form insoluble sulfides):
HE = YFSHFSl - HIF + nYIFGF (41)
Insoluble sulfides:
- HFS2
Soluble iron (and other material which forms insoluble sulfides)
- D _ D + Y „_ ™
9t ~ °F 2 Udz + n ~ YFSGS2 ~ GF (43)
Elemental sulfur:
= YHFS ~ YHFS + nYGS
t SU1 ~ FSU2 SSUl (44)
Pyrite:
j\p
f = YpHFS2 - HP (45)
For simplicity, a number of the normally less significant material
transfers have been omitted from the model. As an example, the additions of
sulfate from reactions GS? and HFS have been omitted because these addi-
tions will normally be significantly smaller than the diffusive transport
and the exchange due to the disruption of the sediment.
BIOCHEMICAL MODEL II
The biochemical model shown in Fig, 37 and Table 18 is similar to that
shown above. Organics and available iron, however, are included in a single
component which is considered insoluble and pyrite is not included (pyrite
135
-------
YFOHFSI
INSOLUBLE
SULFIDES
(FS)
VHFS.
AVAILABLE
IRON
(AF)
YsuHSF.
DISSOLVED
OXYGEN
(0)
YFSUHFS2
ELEMENTAL
SULFUR
(SU)
YOCHCI
FREE
SULF1DE
(S)
YCSHC3
ORGANICS
(0
YSCHC2
YLCHC2
SULFATE
(L)
GS
FIG. 37 - BIOCHEMICAL MODEL II FOR ESTUARINE BENTHIC SYSTEM.
(Available Iron also includes other metals which
form insoluble sulfides; Zn, Sn, Cd, Hg and Cu.)
136
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TABLE 18. DEFINITION OF TERMS FOR BIOCHEMICAL MODEL II
OF ESTUARINE BENTHIC SYSTEM
GS = the oxidation of free sulfides
1
GS- = the loss of free sulfides through photosynthetic anaerobic sulfur bacteria,
HAF = the reaction of available iron with free sulfides,
HC = the aerobic biochemical degradation of organics,
HC = the degradation of organics by sulfate reduction,
HC- = the non sulfate reduction anaerobic biochemical degradation of organics,
HO = the dissolved oxygen demand due to other causes (e.g., oxidation of ammonia,
benthic plant respiration),
HFSi = the oxidation of insoluble sulfides with dissolved oxygen,
HFS2 = the reaction of insoluble sulfides with elemental sulfur to form pyrite,
Y = mass of sulfide produced per mass of organic carbon utilized in non-sulfate
reduction anaerobic decomposition,
Y_R = mass of available iron per mass of insoluble sulfide (primarily ferrous sulfide)
oxidized,
Y = mass of sulfide per mass of available iron reacted to form insoluble sulfides,
ro
Y _.. = mass of elemental sulfur used per mass of insoluble sulfide transformed to
pyrite,
YLC = mass of sulfate per mass of organic carbon utilized in sulfate reduction,
Y = mass of oxygen used per mass of organic carbon oxidized,
UL
YQP = mass of oxygen used per mass of insoluble sulfide oxidized,
Y = mass of oxygen used per mass of free sulfide oxidized,
Uo
Y = mass of sulfide produced per mass of organic carbon utilized in sulfate reduction,
Ygy = mass of elemental sulfur per mass of insoluble sulfides aerobically oxidized,
^SSU = mass °f elemental sulfur per mass of free sulfide oxidized, and
137
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TABLE 19. DEPENDENCE OF REACTIONS OF LOWER RESOLUTION MODEL
(MODEL II) ON BENTHIC CONDITIONS
Reaction
oo
GS1
GS3
HAF
HFS]
HFS-
Aerobic Region
significant (g)
none
very slight
significant (g) (k)
none
Anaerobic-Sulfate
Reduction Region DO
none
slight (10
significant
none
significant (e) (j)
Anaerobic, No Sulfate
Reduction Region
HCj
HC2
HC3
HO
significant
none
none
significant
none
significant
slight (d)
none
none
none
significant (e)
none
(f)
none
unlikely 00
conditional (i)
none
normally slight (e) (j)
(a) D0> 0.1 mg/1; upper region of sediment (h)
(b) D0<0.1 mg/1; S04> 10-50 mg/1; below aerobic region (i)
(c) D0<0.1 mg/1; S04<10-50 mg/1; below sulfate reduction
region
(d) usually YSOHC2> YCSHC3 (j)
(e) likely decreasing with depth
(f) if conditions favorable (e.g., available degradable (k)
organics, no toxic levels fo free sulfides)
(g) normally small sulfate source in comparison to
diffusive transport and exchange due to distrubance (1)
of bottom
light needed
largely dependent upon diffusion
of free sulfides from sulfate re-
duction region
long term significance; time scale
of weeks to years
likely to occur after sediment over-
turned or flushed with oxydenated
water
small fraction of oxygen demand,
primarily biological
-------
can be simply added). The increased simplicity of model 2, however, is
purchased by a loss of "reality" in the model. Nevertheless, model 2 may be
sufficient for most investigations and the greater simplicity may be advantageous,
The relative significance of the reactions of model 2 to benthic conditions
is given in Table 19. The biochemical reactions shown in Fig. 37 and Table 18
are combined with equations (34) and (35) to provide the following mathe-
matical description of the benthic system.
Organic carbon:
Dissolved oxygen:
2
f a °0 7T - UH - YOSGS1 - n-
d Z
(47)
Free sulfide:
f - °S 4 - "II * ^CSHC3 * YSCHC2 - Y!
oZ
- GS, - GS, (48)
Sulfate:
--
d Z
139
-------
Available Iron (and other materials which form insoluble sulfides):
- HAF (50)
Insoluble sulfides:
Elemental sulfur:
HFS1 - HFS2 (51)
- YFSUHFS2 + n YOSU
140
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SECTION XV
ENVIRONMENTAL IMPLICATIONS FOR ESTUARINE BENTHIC SYSTEMS
GENERAL IMPLICATIONS
The conditions within estaurine benthic systems have important influences
on the functioning of estaurine systems. The interfacial and benthic regions
are themselves significant regions of estuarine systems. Moreover, the
overlying water quality can be strongly influenced by the conditions within
the benthic system. Depletion of dissolved oxygen and toxic concentrations
of free sulfide may result. ,
Benthic conditions, particularly within tidal flat areas, may have
significant influences on air quality. The release of sulfur to the
atmosphere as a result of sulfate reduction within coastal regions may be
of the same order of magnitude as the buring of fossil fuels (25)(26)(27).
2
If one assumes that the total area of estuaries is 1,222,000, km (77)
2
then an average estuarine benthic atmosphere sulfur release of 0.1 gm/m -day
would be sufficient to be approximately equal to the annual world sulfur release
from the burning of fossil fuels, 50 X 10 tons/yr. This figure does not include
the influence of bays, seas, deltas,shoreline areas, saline lakes, etc.
The total magnitude of benthic sulfide release to the atmosphere and the
nature and extent of man's activities which would contribute toward an
increased sulfur release are subjects of major concern. The quantitative
evaluation of these subjects is beyond the scope of this report. It is
possible however to describe a variety of activities which can result in
a variety of changes within the benthic systems of environmental concern.
To facilitate the following discussion, reference will be made to the
141
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classification system described in Chapter VI and specifically to Table 6
and Fig. 12. The possible influences of man's activities on estuarine
benthic systems are briefly discussed. Reference to Fig. 2 will help to
clarify these discussions.
CHANGES IN ORGANIC DEPOSITION
A general increase of organic deposition to benthic systems can result
from the input to the estuary of additional organics (such as from waste
outfalls), from the input of inorganic nutrients which lead to an increase
of primary production within the estuary, or to the deposition of organics
re-suspended at some other location. Such increases result in a shift of
benthic states toward region 5 (see Fig. 12)- Dredging operations can,
however, remove organics from previously degraded systems and thus assist in
their recovery.
CHANGES IN INORGANIC DEPOSITION
Increased deposition of inorganic material can result in suffocating
benthic plants and animals. Toxic materials in these deposits can also
harm these communities. Again, controlled removal of toxic materials by
dredging may assist degraded systems to recover.
Potential problems can also occur from reduced sediment transport to
estuaries due to. upstream dams, jetties, channelization, and reduction of
seasonal sediment scouring. Such reduction may lead to a lowered input of
available iron to the systems and thus result in a shift toward regions 4
and 5 of Fig. 12. This shift would be most pronounced if a general increase
in organic deposition also occurred.
CONSTRUCTION OF DIKES, JETTIES, WHARVES, ETC.
Dikes, jetties, wharves, etc. can alter estuarine ecosystems in several
ways. These structures may provide a solid substream on which a highly
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diverse population of attached plants and animals may develop. However,
they can isolate regions from the estuarine system, thus drastically altering
their nature and function. Partial diking of a tidal flat region can lead
to an increased trapping of organics and fine particles. In addition, such
dikes can reduce the more significant periodic scour and thus reduce the
"recycling" of iron. Benthic systems within such regions may shift to a
type 5 system, as appears to be the case at site 5 investigated in this report.
HYDRODYNAMIC CHANGES
Deepening of channels, filling of tidelands, construction of dikes and
jetties, stabilization of banks and other such activities all serve to change
the hydrodynamic regime of estuaries. Changes in advective and diffusive
transport and scour can result in significant changes in organic deposition,
inorganic deposition, salinity distributions, temperature distribution,
distribution of pelagic life stages, inorganic nutrient distribution, dissolved
oxygen distribution and other environmental factors which influence the
estuarine ecosystem in complex (and often unknown) ways; often these changes
show themselves most graphically in altered plant and animal distributions.
Changes which tend to reduce scour may lead to a shift to type 5 systems.
Extremely unstable bottoms may prevent the establishment of stable benthic
communities. Reduction of seasonal salinity variations can disrupt biological
cycles, lead to an increased development of resident populations at the
expense of migrating populations and can contribute to a benthic system shift
toward type 5 by reducing the seasonaly variation in sulfate supply to
benthic systems.
TIDELAND FILLING
Filling tideland with dredging spoils or other materials can have wide-
spread adverse effects in an estuary. In general, tidal flats are highly
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productive areas contributing a major portion of the food to an estuary.
Covering of this valuable tideland can be critical, not only because food
sources are removed but because the total high tide volume of the estuary
is reduced. Many estuarine fishes live in the channels at low tide and
move over the tidal flats as the tide rises to feed (a habit of fishes
which may account for better fishing on an incoming tide). Likewise, shore
birds and a few mammals move into the exposed areas at low tide to feed.
The variety of species .which function at some stage of their life cycle in the
tidelands is great. When tidelands become.covered, diversity is thus likely
reduced.
Within a given estuary a wide range of localized inorganic and organic
benthic depostion rates, DO concentrations, sulfate concentrations, scour
velocities, and other factors determine estuarine benthic types that can be
expected. These conditions, moveover, will change with time. Thus, dif-
ferences can be expected over temporal and spacial dimensions of any estuary.
Within unpolluted regions of estuaries, benthic types 1, 2 and 3 will likely
predominate. The wide variety of benthic systems normally found in tidal
estuaries makes possible a larger biological diversity which, in turn,
likely contributes to the functioning of the entire estuarine system.
A wide range of man's activities can produce significant changes in
temporal and spacial distribution of benthic types. Such changes can result
in a decrease in the variety of benthic systems within space and time. Such
a reduction of system diversity, if extensive, would likely reduce biological
diversity. Man's activities might also result in a general shift, likely
toward benthic type 5 (region 5 of Fig.12) which would be considered -undesirable
because of lower DO concentrations and higher free sulfide concentrations.
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Environmental changes often cannot be explained by a simple casual relationship
to a single activity. Thus, the environmental impact of any particular activity,
must be considered along with a host of other activities (both man-made and
natural).
TRANSIENT CONDITIONS DUE TO DREDGING
Dredging of type 4 and 5 benthic systems would likely have a greater
immediate impact on water quality than would the dredging of types 1, 2 and
3. The release of free sulfide would be most objectionable because of its
oxygen demand and its toxicity. However, because of the rapid reaction
rate of free sulfides and oxygen in estuarine waters, the,short term
release of free sulfide during dredging of sludge deposits might be preferable
to the long term release by undisturbed deposits. Thus, in some cases,
dredging operations could be used to assist the recovery of a system which
had been degraded to a type 5 system. Such dredging, however, could release
heavy metals which had been held in the deposits as insoluble sulfides.
Heavy metals such as mercury and cadmium could have both short term and long
term toxic effects.
Increased turbidity due to dredging operations can decrease light
penetrations, and thereby reduce photosynthesis. It can cause mechanical
blockage of gills and ultimate suffocation of many species, simply because
the gills cannot absorb enough oxygen from the water. Likewise, food
filtering mechanisms, which are often associated with respiratory organs,
may also become blocked by too many particles in the water.
The settling of fine sediment from turbid waters over benthic species
may have catastrophic effects on life cycles if pelagic larva or eggs cannot
leave the sediments or if pelagic larvae are prevented from settling in a
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satisfactory environment. For example, oyster spat cannot attach to the
necessary shell substratum if this shell is covered with a layer of fine
sediment.
The type of benthic system that develops at a given location can be
dependent on the biological turnover previously discussed. A transient
environmental condition, such as a temporary depletion of dissolved oxygen,
could eliminate a community which is contributing to this turnover. A new
benthic type might then develop. The re-establishment of the previous com-
munity might not be immediately possible due to this change in type even
though the unfavorable environmental condition which caused this change had
passed. Thus, periodic unfavorable conditions could have a continuous
influence on benthic systems.
LONG TERM PARTICLE SIZE CHANGE
Significant long-term decreases in sediment particle sizes within
developed estuaries can occur (78) . Adequate data to demonstrate such
long-term decreases, however, are not available for most tidal estuaries.
Decreases in particle size may occur as a result of upstream dams, continued
dredging, construction of jetties, and other similar activities. Particle
size reduction could significantly decrease the permeability of deposits
and thus contribute toward reduced transport of exogenous hydrogen acceptors.
This would in turn lead to a reduction, to the left, of regions 1 and 2
of Fig. 12. A reduction of particle size may also impair the movement of
certain benthic animals within deposits.
SPOIL DISPOSAL
Sediment removal and spoil disposal can result in a shift in benthic
types at both the sites of sediment removal and disposal. As an example,
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the oxidation of sulfides and organics during the removal and disposal
operations might result in a shift from an original type 5 system at the
removal site to a type 2 or 3 system at both sites. The reverse may occur,
however, if during disposal, differential settlings of organics and in-
organics occurred. After settling, the spoil deposits may shift toward
type 5 due to the higher organics within the upper regions of the deposits.
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SECTION XVI
LOWER LEVEL RESOLUTION STUDIES
GENERAL
In the earlier phases of this study, a conventional viewpoint of
estuarine systems was employed which dealt with the estuary as an elongated
water body subject to tidal influences. A mathematical model was developed
which, though having a number of unique features, reflected the same
estuarine systems view as the common one and two dimensional estuarine models
Field studies of tidal hydrodynamics and mixing processes were conducted
to compliment the mathematical model studies. As previously discussed, the
study shifted emphasis to a finer level of resolution in which estuarine
benthic systems were examined.
The latter stages of the study were dominated by this finer level of
resolution view which has been described in the previous chapters. During
the course of this study, a number of subjects principally related to the
lower level resolution view were examined. The principal results of these
examinations are briefly discussed in the following subsections. Reference
is made to the publications which provide specific details and a description
of the general estuarine model employed is given in the appendix.
ADVECTION ERRORS
A basic component of most water quality models involves the advection or
movement of materials by flowing waters. The errors associated with several
finite difference models of advection were examined. These errors were
found to distort model results and these errors were most apparent when the
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advection of slug loads was simulated. The tendency to numerically spread
materials was found to be a common error, along with the production of
oscillations and a tendency to skew distributions. A method of estimating
and controlling these errors was developed and examined. Details of this
phase of the study are found in the following reference.
REFERENCE: Bella, David A., and Grenney, William J., "Finite-Difference
Convection Errors," Journal of the Sanitary Engineering Division, ASCE,
Vol. 96, No. SA6, Proc. Paper 7744, December, 1970, pp. 1361-1375.
ESTIMATING DISPERSION COEFFICIENTS; IN ESTUARIES
Two general methods of determining longitudinal dispersion coefficients
from estuarine salinity profiles (79) (80") (81) were examined. In addition
a third method was developed. The adequacy of these three methods was
examined through the use of the finite difference estuarine model. Methods
which employed the assumption of a steady state were found to be seriously
deficient when river flows varied. This general conclusion was in agreement
with the work of Ward and Fisher (82). The method developed in this study
was found to be the most accurate of the three tested. Details are provided
in the following reference.
REFERENCE: Bella, D.A., and Grenney, W.J., "Estimating Dispersion Coefficients
in Estuaries," (technical note), Journal of the Hydraulics Division, ASCE,
Vol. 98, No. HY3, March 1972.
SLACK WATER BUILD-UP IN ESTUARIES
A series of model runs were conducted in which tidal variations were
included. The Yaquina estuary served as the prototype, although a great
deal of the effort was not spent to model the Yaquina in detail. The
results indicated that pollutant profiles within estuaries were generally
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far more sensitive to variations in the biochemical reactions than to
the dispersion coefficients or tidal variations. This result had a major
influence in directing the study toward the biochemical reactions.
These studies did demonstrate, however, the high pollutant concentrations
could develop in the vicinity of an outfall during the slack water period.
This pollutant accumulation was studied through the mathematical model and a
field study.
A diffuser was installed across the main channel of the Yaquina River
about 35 Km from the mouth, at Newport, Oregon. Rhodamine-B was injected
at a constant rate for a ten hour period and more than 400 samples were
collected. The data indicate a significant build-up during periods of
slack water. The model results simulated average observed trends reasonably
well, however, calculated peaks were lower than the field observations.
Details are provided in the following reference.
REFERENCE: Grenney, W. J. and Bella, D. A., "Field Study and Mathematica
Model of the Slack Water Build-up of a Pollutant in a Tidal River,"
Limnology and Oceanography, Vol. 17, No, 2, March 1972.
TIDAL MEASUREMENTS
During the initial phases of this study, partial support was given to
field studies of tidal elevations and tidal currents in the Yaquina, Alsea,
and Siletz estuaries. Because of the changing emphasis of the study reported
herein, continued support for tidal measurement was later provided through
the Sea Grant Program (NSF), Ocean Engineering, Oregon State University.
Results of the supported studies are provided in the following reference.
REFERENCE: Goodwin, C.R., Emmet, E. W., and Glenne, B., "Tidal Study of
Three Oregon Estuaries," Bulletin No. 45, Engineering Experiment Station,
Oregon State University, Corvallis, Oregon, May 1970.
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SECTION XVII
REFERENCES
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on Estuarine Water Quality," Journal Water Pollution Control
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2. Bella, D.A. and J.E. McCauley, "Environmental Considerations for
Estuarine Dredging Operations," Proceedings of the World
Dredging Conference IV, New Orleans, Louisiana. December 1-3,
1971.
3. Bella, D.A., "Environmental Considerations for Estuarine Benthal
Systems," Water Research, Vol. 6, 1409-1418, 1972.
4. Schroeder, E.D. and A.W. Busch, "Mass and Energy Relationships in
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5. Baas Bscking, L.G.M. and E.J.F. Wood, "Biological Processes in the
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Amsterdam, Vol. 58, 160-181, 1955.
6. Orion Research, Inc., "Sulfide Ion Electrode, Model 94-16," Cambridge,
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8. Fenchel, T., "The Ecology of Marine Microbenthos. IV. Structure and
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10. Berner, R.A., "Diagenesis of Iron Sulfide in Recent Marine Sediments,"
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j.2. Chen, K.Y. and J.C. Morris, "Oxidation of Aqueous Sulfide by 02-
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13. Ostlund, H.G. and J. Alexander, "Oxidation Rate of Sulfide in Sea
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14. Avrahami, M. and R.J. Golding, "The Oxidation of the Sulfide Ion at
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15. Connell, W.E. and W.H. Patrick, Jr., "Reduction of Sulfate to Sulfide
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16. Servizi, J.A., R.W. Gordon and D.W. Martens, "Marine Disposal of Sedi-
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17. Bella, D.A., "Tidal Flats in Estuarine Water Quality Analysis," Oregon
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18. Ramm, A.E., "Some Aspects of the Sulfur Cycle in Tidal Flat Areas and
Their Impact on Estuarine Water Quality," Doctoral Thesis, Oregon
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19. Richards, R.A., "Chemical Observations in Some Anoxic, Sulfide-Bearing
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20. Ivanov, M.V., "Microbiological Processes in the Formation of Sulfur
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22. Colby, P.I. and L. L. Smith, Jr., "Survival of Walleye Eggs and Fry on
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23. Dimick, R.E., "The Effects of Kraft Mill Waste Effluents and Some of
Their Components on Certain Salmonoid Fishes of the Pacific
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24. Haydu, E.P., H.R. Amberg and R.E. Dimick, "The Effect of Kraft Mill
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25. Kellogg, W.W., R.D. Cadle, E.R. Allen, A.L. Larus and E.A. Martell,
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26. Hitchcock, D.R. and A.E. Wechsler, "Water Pollution and the Atmospheric
Sulfur Cycle," publication pending.
27. Hitchcock, D.R., "The Production of Atmospheric Sulfur from Polluted
Water," publication pending.
28. Moore, H.B., "The Muds of the Clyde Sea Area. III. Chemical and
Physical Conditions; Rates and Nature of Sedimentation; and
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29. Moore, H.B., "The Specific Identification of Faecal Pellets," Journal
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30. Carriker, M.R., "Ecology of Estuarine Benthic Invertebrates, a Per-
spective," In: Estuaries (G.H. Lauff, editor), American Assoc.
for the Advancement of Science, Pub. 83, 442-487, 1967.
31. Hata, Y., H. Kadota, H. Miyoshi, and M. Kimata, "Microbial Production of
Sulfides in Polluted Coastal and Estuarine Regions," In: Advances
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32. Martin, B.C., "The Effect of Mixing on the Oxygen Uptake Rate of Estuarine
Bottom Deposits," Masters Thesis, Oregon State University, Corvallis,
1969.
33. Martin, D.C. and D.A. Bella, "Effect of Mixing on Oxygen Uptake Rate of
Estuarine Bottom Deposits," Journal Water Pollution Control Federation,
Vol. 43, 1865-1876, 1971.
34. Crook, G.R., "In Situ Measurement of the Benthal Oxygen Requirements of
Tidal FlatTDeposits," Masters Thesis, Department of Civil Engineering,
Oregon State University, Corvallis, 1970.
153
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35. Crook, G.R. and D.A. Bella, "In Situ Measurements of the Benthal Oxygen
Requirements of Tidal Flat Areas," Proc. 25rd Industrial Waste
Conference, Purdue University, 1970.
36. McKeown, J.J., A.H. Benedict and G.M. Locke, "Studies on the Behavior
of Benthal Deposits of Paper-Mill Origin," Technical Bull. No. 219,
National Council of the Paper Industry for Air and Stream Improve-
ment, New York, September 1968. 28 p.
37. Rolley, H.L.J. and M. Owens, "Oxygen Consumption Rate and Some Chemical
Properties of River Muds," Water Research, Vol. 1, 759-766, 1967.
38. Edwards, R.W. and H.L.J. Rolley, "Oxygen Consumption of River Muds,"
Journal of Ecology, Vol. 53, 1-27, 1965.
39. Hanes, B.N. and T.M. White, "Effects of Sea Water Concentration on
Oxygen Uptake of a Benthal System," Proceedings of the 22nd
Industrial Waste Conference, Purdue University, Engineering Exten-
sion Series No. 129, 1967.
40. Stein, J.E. and J.G. Denison, "In Situ Benthal Oxygen Demand of Cellulosic
Fibers," In: Advances in Water Pollution Research, Proceedings
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Munich, Germany, Vol. 3, 1967. p. 181-197.
41. Pamatmat, Mario M. and K. Banse, "Oxygen Consumption by the Seabed.
^- In situ Measurements to a Depth of 180 m," Limnology and
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42. Fair, G.M., E.W. Moore and H.A. Thomas, "The Natural Purification of
River Muds and Pollutional Sediments," Sewage Works Journal, Vol. 13,
270-307, 756-779, 1941.
43. Baity, H.G., "Studies of Sewage Sludge," Sewage Works Journal, Vol. 10,
539, 1938.
44. Ogunrombi, J.A. and W.E. Dobbins, "The Effects of Benthal Deposits on
the Oxygen Resources of Natural Streams," Journal of the Water Pol-
lution Control Federation, Vol. 42, No. 4, 538-551, 1970.
45. U.S. Federal Water Pollution Control Administration, Middle Atlantic
Region, "An In Situ Benthic Respirometer," CB-SREP Technical Paper
No. 6, 1968, mimeographed.
46. Bella, D.A. and W.F. Dobbins, "Difference Modeling of Stream Pollution,"
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48. American Public Health Association. Standard Methods for the Exam-
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Acoel Flatworm, in an Estuarine Mudflat on the Central Coast of
Oregon," Doctoral thesis, Oregon State University, Corvallis,
1971.
50. Reeburg, W.S., "An Improved Interstitial Water Sampler," Limnology and
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51. Bertolacini, R.J. and J.E. Barney, II, "Colorimetric Determination of
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281-283, 1957.
52. Washington State Pollution Control Commission, "Pollutional Effects of
Pulp and Paper Mill Wastes in Puget Sound," Federal Water Pollu-
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54. Zobell, C.E., "Ecology of Sulfate Reducing Bacteria," Producers Monthly,
Vol. 22, No. 7, 12-29, 1958.
55. Postgate, John, "The Chemical Physiology of the Sulfate-Reducing
Bacteria," Producers Monthly, Vol. 22, No. 19, 12-16, 1958.
56. Rubentschich, L.I., "Sulfate Reducing Bacteria," Microbiology, Vol. 15,
443-456, 1946.
57. Redfield, Alfred C., "The Biological Control of Chemical Factors in
the Environment," American Scientist, Vol. 46, 204-221, 1958.
58. Downing, A.L. and G.A. Truesdale, "Some Factors Affecting the Rate of
Solution of Oxygen in Water," Journal of Applied Chemistry,
Vol. 5, 570-574, 1955.
59. Gloyna, E.F. and E. Ernesto, "Sulfide Production in Waste Stabilization
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60. Edwards, V., "Analytical Methods in Bacterial Kinetics," Doctoral Thesis,
University of California, Berkeley, 1967.
61. Wong, P.S., "Determination of Sulfate Reduction by Using Radio-Tracer
Methodology," unpublished graduate report, Oregon State University,
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62. Murthy, A.R.V. and K. Sharada, "Determination of Sulfide Sulfur in
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63. Jeffery, P.G., Chemical Methods for Rock Analysis, Pergamon Press,
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65. Cooper, T., Research Associate, Personal Communication, Oregon State
University, Corvallis, 1971.
66. Nakai, N. and M.L. Jensen, "The Kinetic Isotope Effect of the Bacterial
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67. Olsen, R.C., "Measuring Estuarine Benthal Sulfide Release Rate,"
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68. Branfield, A.E., "The Oxygen Content of Interstitial Water in Sandy
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69. Jansson, B.O., "The Availability of Oxygen for the Interstitial Fauna
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70. Edwards, R.W., "The Effect of Larvae of Chironomus riparius meigen on
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71. Rhodes, D.C., "Rates of Sediment Reworking of Intertidal and Subtidal
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72. Perkins, R.K. and O.C. Johnston, "A Review of Diffusion and Dispersion
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74. Duursma, E.K., "Molecular Diffusion of Radioisotopes in Interstitial
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75. Bungay, H.R., J.W. Whalen and W.M. Sanders, "Microprobe Techniques for
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76. Milhous, N.T., "Water Interchange in an Estuarine Tidal Flat," North-
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77. "National Estuarine Pollution Study," Report of the Secretary of the
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78. Fleming, G., "Sediment Balance of Clyde Estuary," Journal of the
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ADDITIONAL REFERENCES
Bella, D.A., "Modeling and Simulation of Stream and Estuarine Systems," Eighth
Annual Workshop, Association of Environmental Engineering Professors, Nassau,
Bahamas, 1972.
Dornhelm, R.B. and Woolhiser, D.A., "Digital Simulation of Estuarine Water Quality,"
Water Resources Research. Vol. 4., 1317-1328, 1968.
Fisher, H.B., "A Lagrangian Method for Predicting Dispersion in Bolina's Lagoon,
California," Open File Report, Geological Survey, Menlo Park, California, 1969.
Glenne, G., Personal Communication, Oregon State University, Corvallis, 1969.
Grenney, W. J. and Bella, D.A., "Field Study and Mathematical Model of the Slack-
Water Buildup of a Pollutant in a Tidal River," Limnology and Oceanography, Vol.
17, No. 2, 229-236, March, 1972.
Grenney, W. J., "Modeling Estuary Pollution by Computer Simulation," M.S. Thesis,
Civil Engineering, Oregon State University, Corvallis, 1970.
Ippen, A.T., "Estuary and Coastline Hydrodynamics," McGraw-Hill Publishing Co.,
New York, N.Y., 1966. 744 p.
157
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SECTION XVIII
PUBLICATIONS
1970 1. Bella, D.A., "Role of Tidelands and Marshlands in Estuarine
Water Quality," Proc. Northwest Estuarine and Coastal
Zone Symposium, 1970.
2. Bella, D.A. and W.J. Grenney, "Finite Difference Convection
Errors," Journal of the Sanitary Engineering Division,
ASCE, Vol. 96, 1352-1361, 1970.
3. Crook, G.R. and D.A. Bella, "In^ Situ Measurements of the
Benthal Oxygen Requirements of Tidal Flats," In: Proc.
23rd Industrial Waste Conference, Purdue University, 1970.
4. Goodwin, C.R., E.W. Emmett and B. Glenne, "Tidal Study of
Three Oregon Estuaries," Oregon State University Engineer-
ing Experiment Station Bulletin No. 45, Corvallis, 1970.
5. Crook, G.R., "In Situ Measurement of the Benthal Oxygen Require-
ments of Tidal Flat Deposits," M.S. Thesis, Civil
Engineering, Oregon State University, Corvallis, 1970.
6. Grenney, W.J., "Modeling Estuary Pollution by Computer Simu-
lation," M.S. Thesis, Civil Engineering, Oregon State
University, Corvallis, 1970.
1971 7. Bella, D.A. and J.E. McCauley, "Environmental Considerations
for Estuarine Dredging Operations," In: IV Proceedings
of the World Dredging Conference, New Orleans, Louisiana,
1971.
8. Ramm, A.E., "Some Aspects of the Sulfur Cycle in Tidal Flat
Areas and Their Impact on Estuarine Water Quality,"
Ph.D. Thesis, Oregon State University, Corvallis, 1971.
9. Martin, D.C. and D.A. Bella, "Effects of Mixing on Oxygen
Uptake Rate of Estuarine Bottom Deposits," Journal Water
Pollution Control Federation, Vol. 43, No. 9, pp. 1865-1876,
September 1971.
158
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1972 10. Bella, D.A., A.E. Ramm and P.E. Peterson, "Effects of Tidal
Flats on Estuarine Water Quality," Journal Water Pollu-
tion Control Federation, Vol. 44, No. 4, 541-556,
April 1972.
11. Bella, D.A. and W.J. Grenney, "Estimating Dispersion Coef-
ficients in Estuaries," Journal of the Hydraulics
Division, ASCE, Vol. 98, No. HY 3, 583-589, March 1972.
12. Bella, D.A., "Environmental Considerations for Estuarine
Benthal Systems," Water Research, Vol. 6, 1409-1418,
1972.
1973 13. Peterson, P.E., "Factors That Influence Sulfide Production in
an Estuarine Environment," M.S. Thesis, Oregon State
University, Corvallis, 1973.
14. Ramm, A..E. and D.A. Bella, "Aspects of the Sulfur Cycle in
Tidal Flat Areas" (title subject to revision), Limnology
and Oceanography (in press].
159
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SECTION XIX
FINITE DIFFERENCE MODEL
INTRODUCTION
The purpose of this appendix is to provide a description of the finite
difference estuarine model which was used in the early phases of this research
project. The model was developed for the research project and the following
three basic criteria were used in its selection:
(1) The model should be accurate.
(2) The model should be simple and flexible so that many revisions
could be made.
(3) The model should be reasonably efficient with respect to
computer time.
It was not the intent of the investigators to develop a standard computer
program for general use. We have found, however, that the finite difference
methods selected and described herein are relatively simple (particularly the
water quality model) and thus they should be of general use. Only a first
order biochemical reaction will be presented below, however, the methods pre-
sented can be simply adapted to a wide variety of biochemical reactions and
processes.
This appendix is based principally upon two references (Grenney, 1970
and Bella, 1972) and referral to them should be made for additional information.
GENERAL STRUCTURE
The main computer program is based on the finite-difference method de-
veloped by Bella and Dobbins (46). The stream channel is divided into equal
length segments (AX) and the mass within each segment is computed for finite
time increments (AT). In this model, AT is less than a tidal cycle. Segments
160
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are labeled beginning with segment one at the fresh water end of the channel
as shown in Figure Al, where N = segment number on the main channel and n =
interface number. The program is versatile in that irregular estuary configura-
tions can be simulated by appropriate arrangement of the segments. Tributaries
can be attached to the main channel as shown in Figure Al where K are seg-
ment numbers of the tributary intersecting the main channel at segment N. Mud
flats can be simulated by a series of adjacent short tributaries where material
is transferred across all four interfaces of each interior segment. A two-
dimensional effect can be achieved by superimposing two or more channels.
The program was written in Fortran IV for use on the CDC 3300 computer
at Oregon State University. Figure A2 is a flow diagram of the main program.
Four types of input data are required:
1. Finite-difference grid parameters AX and AT. The amount of numerical
error introduced by the finite-difference scheme was found to be
very sensitive to these parameters (See Bella and Grenney, 1970).
2. Estuary configuration consisting of the cross section area,
channel side slopes, and mean water depth at each segment and the
location and configuration of each tributary.
3. Hydraulic characteristics of the main channel and each tributary.
These data include magnitude of tidal wave at the mouth, speed of tidal
wave propagation, channel friction, and fresh water inflow rates.
4. Mass transfer parameters and initial conditions. Those data
include the dispersion and decay coefficients and the initial
pollution concentration in each segment. Also included is the
location of pollution sources and the rate of pollution injection
at each source.
161
-------
(tributary fresh
water flow)
C VH
? OJ
ri ,d
C M
C 4)
n-1
(K-2)
N
r-i
N
N+l 1
n
n+l
i ~ ?
1 o
0)
'C "M
FIG. Al - REPRESENTATION OF ESTUARY CHANNEL AND TRIBUTARY.
162
-------
Input data)
Write initial conditions
Increment time by an amount At
I Water quantity model
Water quality model j
no
Add pollutant to specific segments
by subroutine SOS INK
Is output desired at this time?
yes
no
Write output
Has time exceeded maximum time for run?
yes
FIG. A2 - FLOW CHART FOR THE MAIN COMPUTER PROGRAM.
163
-------
For each time increment, the program begins at the fresh water
end of the main channel (segment number one) and calculates flows in each
successive segment by means of the water quality model. As the calcula-
tions proceed down the estuary, each segment is checked for an intersecting
tributary. When tributaries are encountered, control is shifted to a sub-
routine which calculates tributary inflow. The main channel flow is adjusted
by the amount of the tributary flow and the program proceeds to the next segment
After flows have been calculated in all segments, the program returns to
segment one and again proceeds down the main channel calculating pollutant
concentrations by means of the water quality model. When tributaries are
encountered, control is shifted to a sub program which distributes the pollutant
in the tributary.
When the pollution distribution has been calculated for all segments in
the estuary, the program returns to segment number one and proceeds back down
the main channel checking for pollution sources. When outfall locations are
encountered pollution is injected by subroutine SOSINK.
Results are printed in predesignated times. Output includes the velocity,
area, dispersion coefficient and concentration in each segment.
WATER QUALITY MODEL
General
Conceptualize a stream as a series of mixed cells of length AX as shown
in Figure A3. A number of memory locations are established in the computer
to record the water quality and quantity conditions within each cell. The
longitudinal mixing, as an example, that occurs over a small time interval
164
-------
of length AT is simulated by numerically exchanging a given amount of water
and thus pollutant in each segment with the two adjacent segments. The amount
of such water exchanged between segments is dependent on the length of the
time interval, AT, the segment size, and the hydraulic conditions being
simulated in the stream. The pollutant mass at the end of the time interval is
given by the simple mass balance equation
/ Pollutant Mass\ / Pollutant Mass\ I Net Mass \
in segment at j = in segment at 1 + I exchanged I
\ time T+AT / \ time T / \ during AT I
This "pollutant exchange", of course, involves the addition and subtraction
of the pollutant masses recorded at the appropriate memory locations in the
computer. These computations are performed for each segment along the length
of the stream and are repeated for successive time intervals in order to
describe the change in pollutant mass in each segment (from which the con-
centration can be calculated) over time due to mixing.
The simultaneous advection and biochemical reactions of a pollutant can
also be simulated. Advection is simulated by transferring during each time
interval a portion of the pollutant in each segment to the segment immediately
downstream. Biochemical reactions are simulated by adding or subtracting
appropriate amounts for each segment over each time interval. A pollutant
FIG. A3 - STREAM CONCEPTUALIZED AS A SERIES OF MIXED CELLS
165
-------
outfall can be simulated by adding pollutant mass to the particular segment
at the outfall location for each time interval. The simultaneous action
of all of these processes is simulated by performing all such exchanges and
removals for all segments during each time interval. The process is repeated
over a sufficient number of time intervals (of length AT) so as to span the
time period of interest. If AX and AT are sufficiently small, the finite-
difference results will approximate the continuous processes being simulated.
Each stream segment (Figure A3) will be considered as completely mixed.
The concentration within any segment N at time T will be designated by C(N,T).
That is C(N,T) is averaged over the space interval, AX, but is not averaged
over the time interval, AT. Concentrations are recorded at the beginning and
end of time intervals.
The cross sectional areas are similarly defined. That is, A(N,T) designates
the average cross sectional area of segment N at time T.
For the present discussion, only first order biochemical reactions will be
considered. The first order reaction coefficient, K.., may vary with distance
and time. The value of K- for the entire segment N during the entire time
interval beginning at time T will be designated as K1(N,T).
The mean water velocity, U and the dispersion coefficients, DT, always
lj
appear as products with A, the cross sectional area. For simplicity these
products will be designated as UA, the total flow rate, and DA, the total
dispersion coefficient. In order to keep the number of subscripted variables
to two, the following notation is used. UA(N+1,T) and DA(N+1,T) equal the
average values of UA and DA at the interface of segments N and N+l over a
time interval which begins at time T. UA(N,T) and DA(N,T) equal the average
values of UA and DA at the interface of segment N and N-l over a time interval
which begins at time, T.
166
-------
The above finite-difference definitions are not standard definitions.
One must closely examine the definitions used in each modeling approach. In
the following sections, advection, dispersion, decay and pollutant addition
will be individually modeled in finite-difference form. A method of combining
these individual processes will then be presented.
Advection
The advection can be conceptualized in finite-difference terms as a water
transfer from upstream segments repeated for each time interval, AT (see
Figure A4). The volume of water transferred in each step is equal to the
flow rate between segments times the length of the time interval as shown
in Figure A4.
U
UA(N,T) UA(N+1,T)
N-l
N
N+l
FIG. A4 - FINITE-DIFFERENCE ADVECTION
The mass transferred into segment N will be equal to the volume of water
transferred from segment N-l to segment N (assuming direction of flow from
left to right) times the pollutant concentration in segment N-l. That is,
the pollutant mass transferred into segment N equals
UA(N,T)ATC(N-1,T)
167
-------
Similarly the pollutant mass transferred out of segment N equals
UA(n+l,T)ATC(N,T) (2)
The pollutant mass at the start of AT will be equal to
C(N,T)A(N,T) (3)
while the pollutant mass at the end of the time interval will be equal to
C(N,T+AT)A(N,T+ AT) (4)
Consider a mass balance of pollutant within segment N. One obtains:
(mass at end of AT) = (mass at start of AT) + (mass advected in during AT) -
(mass advected out during AT). Substituting equations (1), (2), (3) and (4)
into this mass balance leads to
C(N T+AT) = CCN,T)A(N,T)
HN,I+AJJ A (N,T+AT)
UA(N,T)C(N-1,T)AT
+ A(N,T+AT)AX
UA(N+1,T)C(N,T)AT ,,.,
A(N,T+AT)AX L J
Equation (5) is the finite-difference model of advection with variable
parameters when the velocity flows from segment N-l to segment N.
Should the velocity reverse direction, equation (5) must be replaced by
equation (6)
CfN T+AT) - C(N,T)A(N,T)
HN,l+AlJ A(N,T+AT)
UA(N+1,T)C(N+1,T)AT
A(n,T+AT)AX
UA(N,T)C(N,T)AT
A(N,T+AT)AX l }
Equation (5) and (6) are subject to the restriction
UAT < AX (7)
168
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Explicit Finite-Difference Model of Dispersion
Consider any stream segment N as illustrated in Figure A5. At time T,
the beginning of the time interval, equal elements of water, of volume w(N,T),
are exchanged between segments N and N-l. Similarly, equal elements of water,
of volume w(N-i-l,T), are exchanged between segments N and N+l. The elements
of water are exchanged and all cells are completely mixed over the time
interval AT.
The mass of pollutant leaving segment N during this exchange is:
w(N,T)C(N,T)+w(N + l,T)C(N,T) (8)
while the mass of pollutant entering segment N from segment N.-1 and N+l during
AT is :
w(N,T)C(N-l,T)+w(N+l,T)C(N+l,T) (9)
The change in mass in segment n during AT equals
[C(N,T+AT) - C(N,T)]A(N,T)AX (10)
setting the change in mass within segment N equation (10), equal to the mass
input to segment n minus the mass output from segment N leads to
C(N,T+AT) = C(N,T)
[C(N-1,T)-C(N,T)]
+ A MAX [C(N+1,T)-C(N,T)] (11)
From a mixing length description the dispersion coefficient may be
considered as
DT = qh1 (12)
L
in which q is the volume rate of water exchanged per unit cross sectional
area and h1 is the effective length of the exchange. From the difference
model described above and equation (12), one obtains:
169
-------
and
Dr =
WAX
L AAT
w =
D AAT
LJ
If
(13)
(14)
W(H,T) W(N+1,T)
^
N-l
N
N+l
FIG. AS - FINITE-DIFFERENCE DISPERSION
Using the notation previously given, one obtains
w(N,T) =
AX
[c(N_ljT)_c(NjT)]
and
w(N+l,T) =
DA(N+1,T)AT
AX
[C(N+1,T)-C(N,T)]
(15)
(16)
Substitution equations (15) and 16) into equation (11) leads to
C(N,T+AT) = C(N,T)
DA(N,T)AT
A(N,T)AX2
+ DA(N+1,T)AT
A(N,T)AX2
(17)
Equation (17) enables one to explicitly obtain the concentration in segment N
at the end of the time interval. Repeated use of equation (17) will closely
simulate the dispersion (and diffusion) process if AX and AT are sufficiently
small.
170
-------
It is reasonable from the above approach that the total volume of water
exchanged during the time interval should not exceed the volume of the segment.
That is:
w(N,T) + w(N+l,T) < A(N,T)AX (18)
Substituting (15) and (16) into (18) leads to
DA(N,T) DA(N*1,T) AX2
A(N,T) A(N,T) * AT
If DL and A do not vary with length, X, one obtains from equation (19)
the stability requirement for the standard explicit scheme for the approximation
of the diffusion equation:
D AT
AX
(20)
To prevent an oscillation error, one should accept the following requirement.
D AT
-— - < 1/2 (21)
AX
First Order Biochemical Reactions
If one assumes that the change of pollutant mass within any segment N that
occurs over a time interval of length, AT, is proportional to the pollutant mass
within the cell during the time interval and proportional to the length of the
time interval, one obtains
C(N,T+AT) = C(N,T)-K1(N,T)AT(1-6)C(N,T)+6C(N,T+AT)] (22)
in which e is a weighing function (0 <6< 1).
The Addition of Pollutants
Pollutants may be added to any stream segment. By equating the pollutant
mass in any segment N at the end of the time interval (T+AT) , to the pollutant
mass in the segment at the beginning of the time interval (T) plus the pollutant
mass added over the time interval, AT, one obtains:
171
-------
C(N,T+AT) = C(N,T) H- C23)
in which m(N,T) equals the average pollutant mass input rate into segment N
over the time interval beginning at time T.
Combining Dispersion, Advection, Decay and Additions
Equations (5), (17), (22) and (23) can be simply combined in order to
simulate the simultaneous occurrence of advection, dispersion, decay and addi-
tions. One merely utilizes each of these four equations sequentially. The
final concentration results of a given step serve as the initial concentrations
for the following step. As an example, equation (5) would be used to describe
the concentration changes due to advection alone. The results of equation (5)
would then be used for the initial conditions for the dispersion equation (17).
That is C(N,T+AT) from equation (5) would serve as C(N,T) in equation (17).
The results (T+AT) of equation (17) would then serve as initial (T) concentra-
tions for equation (22). The C(N,T+AT) values obtained from equation (22)
would serve as the C(N,T) values of equation (23). Change in the cross sectional
area, A(N,T), should be included during the advection step. The use of all
four equations would describe the pollutant concentration changes over the time
interval due to advection, dispersion, decay and additions. These steps would
then be repeated for successive time intervals (of length, AT) until the computa-
tions covered the desired time span being simulated. The computational sequence
of equations (5), (17), (22) and (23) can be changed in any order with no
significant change in the results.
Equations (5) and (6) produce a numerical mixing error. This error can
be described by an effective or pseudo dispersion coefficient given by
D = | [AX - UAT] (24)
In order to compensate for this error, D from equation (24) must be sub-
tracted from the actual dispersion coefficient used in equation (17) for each
time and length interval.
172
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WATER QUANTITY MODEL
The most simple method for estimating stream velocities is to assume
uniform flow throughout the estuary and apply a sinusoidal velocity at the mouth.
A more realistic method is to determine water surface elevations as a function
of distance and time and calculate flows from-known characteristics of the
channel. This can be accomplished by solving the continuity and momentum equa-
tion for unsteady flow. Although this method is accurate, a great amount of
computer time is required. A more efficient method has been to use changes in
water surface elevations to calculate average flows over small time intervals,
i.e., (average flow out of segment) = (average flow in) - (change in volume)
(Fisher, 1969). This method has been adopted for the present study and can be
represented in finite-difference terms as follows for flow in the direction
shown in Figure Al.
UA(N,T) = UA(N-1,T) + [A(N,T-1) - A(N,T)] |jr (25)
By using this approach, the problem is reduced to one of finding an efficient
means for predicting water surface elevations (H).
Fisher (1969), in studies of Bolinas Lagoon, California, used observed
values of H over a few tidal cycles. The use of tabulated values becomes
awkward for long period of analysis. Dorlhelm and Wollhiser (1968) predict
H by propagating a sine shaped tidal wave up the estuary. Tidal actions in
most real estuaries do not conform to this simple representation however.
Frequently a reflected sine wave can be used to predict tidal heights along
an estuary (Ippen, 1966). Consider the channel profile of length L shown in
Figure A6. An imposed wave is assumed to travel up the estuary from the mouth.
A hypothetical boundary exists at the end of the estuary which reflects a
portion of the incident wave. The water surface at a point x feet from the
173
-------
C"
fi
0
• r-i
-t-J
rd
-------
boundary can be predicted by superimposing the height of the reflected wave onto
the incident wave. The effects of friction can be approximated by assuming an
exponential reduction in wave height with distance. Mathematically, the tidal
height can be represented as (Glenne, 1969):
H = HQ + a[eyXcos(6T + kx) + 3e"yxcos (
-------
(Subroutine AREA]
No
Has program reached the end of
the current tidal cycle
Yes
Calculate ordinates of the imposed
tidal wave at the mouth of the
estuary for the next
tidal cycle
Based on the speed of wave propagation, and
friction of the channel, calculate or-
dinates of the incident wave at the
center of each channel seg-
ment
Calculate ordinate of reflected
wave and superimpose
on incident wave
Calculate average cross section
areas for each segment
Return
FIG. A7 - FLOW CHART FOR SUBROUTINE "AREA".
176
-------
For any particular time at the mouth of the estuary, T, the ordinate
of the imposed wave at any point, x, in the estuary can be calculated by
determinining the time required for the wave to move up the channel to that
point. This is the lag time and is represented by TRAV on Figure A8. For
this study, the time lag was expressed as:
TRAV = c(l-x)
where c is wave celerity, x is the distance from the fresh water end, and L
is the length of the estuary. More accuracy could possibly be achieved by
allowing c to vary as a function of depth; however, at the sacrifice of com-
puter time- Once T is located (Figure AS), the elevation is obtained by
J\
interpolating between points M and M + 1.
The ordinate of the reflected wave is obtained in a similar manner. A
portion of the incident wave is assumed to bounce off the boundary at the end
of the estuary and travel back towards the mouth at the same celerity. The
total lag time for the wave to travel from the mouth back to point x can be
calculated by:
TRAV = c(L+x)
Friction is incorporated in the model by reducing the ordinates by an
exponential function of the distance traveled by the wave, i.e., e p^
for the incident wave and e~V^ +X^ for the reflected wave. The water surface
elevation is then obtained by superimposing the ordinates of the incident and
reflected waves. Friction need not be represented by an exponential function;
however, it was selected so that existing methods (Ippen, 1966) could be used
to estimate the parameters u and k. It may not be the most realistic repre-
sentation, however, because the major frictional influence is exerted near the
mouth of the estuary and the frictional effect decreases with distance up
177
-------
Time
FIG. AS - TIDAL WAVE REPRESENTATION AT THE MOUTH OF AN ESTUARY
.2
s
. 1
x
rt
10
15
20
25
30
Number of increments in tidal cycle
FIG. A9 - ERROR INTRODUCED BY LINEAR INTERPOLATION
178
-------
the estuary. Conversely, in some real estuaries the effect of friction would
probably be least near the mouth and increase with distance up the estuary.
Cross section areas are calculated as a function of the water surface
elevation and side slopes of the channel at each segment.
In order to calculate flows by means of equation (25) it is necessary to
know the flow across the interface of segment one at the upper end of the
estuary and in every tributary. For a completely reflected wave all of the
tidal induced flow is reflected and, therefore, the flow across the first
interface is equal to the fresh water inflow. Physically this can be visualized
as a waterfall forming a complete barrier at the end of the estuary. However,
when the upstream boundary reflects only a fraction, 3, of the incident wave
amplitude, a certain amount of flow will be induced at segment one due to tidal
action beyond the boundary. When this flow is neglected, significant errors
may be introduced for values of g less than one.
The tidal induced flows at the boundary can be calculated by extending
the hydraulic model beyond the boundary a distance necessary for the -unreflected
portion of the wave to become significantly attenuated by friction. For
progressive waves with low friction this method may require excessive computer
time. One approach to reduce computer time and still approximate flows across
the first interface would be to increase the friction coefficient beyond the
boundary. Errors introduced by this approximation would have to be investigated
for each individual case.
If the reflected wave is not completely attenuated when it reaches the
mouth of the estuary, the calculated water surface elevation will not coincide
with the incident wave and a discontinuity will occur at the ocean boundary.
Ippen (1966) avoids the problem by applying the incident wave at the end of the
179
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estuary instead of the mouth. However, tidal fluctuations are generally
not recorded at the upper end of an estuary, and this approach is not always
practical.
If there is sufficient friction in the channel, or if g is low, the
discontinuity will be negligible. When a substantial discontinuity does exist
at the ocean boundary, a driving wave must be found such that superposition of
the reflected wave will result in water surface elevations which agree with
observed data. Finding a driving wave is difficult, but it can be done by
trial and error for short runs.
The water quantity model was computed for one tidal cycle by adjusting
the model to fit measured results (Grenney and Bella, 1972). The single tidal
cycle was repeated to simulate longer runs with a minimum computer time.
A description of program subroutines is given in Table Al and a listing
of the programs is provided in Table A2.
180
-------
TABLE Al - DESCRIPTIONS OF COMPUTER PROGRAMS
PROGRAM ESTREF: main program
a) Controls input/output and subroutines
b) Calculates initial conditions
c) For each AX and AT calculates: average flows, convection, dis-
persion, and decay
SUBROUTINE INPUT: reads input coefficients and identification number
scheme for finite-difference representation of estuary.
SUBROUTINE INCON: reads initial conditions values for variables;
calculates cross-sectional areas and stream surface widths for
mean water depth; and sets up initial concentrations in each
segment.
SUBROUTINE INAREA: calculates initial cross-sectional areas at each
segment given an initial tidal height at the mouth.
SUBROUTINE TIDE: calculates tidal height at the mouth as a function of
time.
SUBROUTINE AREA: propagates tidal wave inland and calculates cross-sectional
areas at each segment for each AT.
SUBROUTINE SOSI: provides input or removal of material at specified
segments and times to represent exogenous sources and sinks.
FUNCTION FLWU: fresh water flow entering the system at the inland
boundary.
181
-------
TABLE A2---LISTING OF COMPUTER PROGRAMS FOR ESTUARY MODEL
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lEXPK<300)»<1HK(!50li).UH( JoO), A( 300 • 2 ) . 1 1 ( 300 > » AA2C 300) > NroT»€LM»
2U.TH».OTM»T»TMAXiCKl»QU(JlO).U.»CI 1 • C ll»» RFPT »[>! H» MTUT.MQ. TCU TE»
3AnTH»ALFA»LASI .L)X»EL» '-'A?. J JJ. 1 RAVK< 300) , rtKC
10 CALL INplll
THAl»t.O-THA
OTH«nTM/60.0
1>T"UTM*60.0
CK1»CK1/86«00
RNTuT-NTOT
OTX»OT/DX
KIP-0
K»2
Jl 1«1
LASTP]«LAST+1
INITIAL
CALL 1NCON
CALL INARLA
UA{2)«FL«U(T)
C1(1)«C11
ClCLAST+U-ClN
W»ITF(?.«)0) NTUT»tL.1»OX,OT".ll. ThA,l M.LL
50 FPHdATflH .//•// 1H »3V,Ia. '1F12.3 //1H • 3X. 3K 1 ?. 3. I
HRITE(2«54) T
51 FnRMAKlH ./ Iri ,3X, Kl*. 2)
ICONl«l
ICON2»LAST+1
55
35 T«TtOTH
CALL A»tA(K)
AVt FLO^b AT ALL SfCTIl)l"!>
DH d N-2.LAST
NP«N+1
UA(2)"FLWU(T)
8 CONTINUE
L»l
on 72 N«?.LASI
T\ CALL SOSIIT.RATE)
C1CN)«C1(N)+RATE*I)TH/(A(M,J)+A(N.KJ)/2.0
L«L*1
72 CONTINUE
CDNVLCTIDN AT ALL SECTIONS
IF(UA(2» M>oO,60
61 F1"OA(2)*CH2)
OPA(2)«-UA(2)*(UX+UMCz)*nT/A(2»K))/2.0
GP TO 62
60 F1*UA(2)*CU U
OPA(?)« IIA(2)*(I)X-UAC2)*I1T/A{^.K. ) 1/2.0
182
-------
TABLE A2 (continued) LISTING OF COMPUTER PROGRAMS FOR ESTUARY MODEL
62 l!A»(?)«L'(i')
(in 63 N*?.LASl
MPsN+1
6*
fin in
6S F?»lJA(
OP AC UP )m UA{»ih>)»<0<-UA(NP)*i>T/A(IMW»K))/2.0
66
63 CONFTNUfc
C I'lSPFKSiUN «T ALL ...,„
F««t(C2ll)-C2C2))*')«A(/))*(i.o-ALf»)
DO 9 N«2»L»ST
-C21 N) )*uAA( "i+1 )
9 CDNFlNUt
UECAr Af ALL
00 11 N*?.LASl
Cl (N)«C?(lx)*Fj
it cnwriNiiE;
o a I p u I
IF(uUCKull)-| ) 25> -»b»2't
25 KntJ*KOU+l
WRTTE(2.S2) T
52 FnHHATtlH »/ 1H »3X. F I S . "5 )
WPI rF(?,"i6) iLOul •Ottl 1 J.C1 I 1 )
56 FHRMAKlM . I 5. F 1 2. 4, 26<. F 15. i| )
00 21 I»?.LASIPl
WRITE<2.">3> I.Ont I )»VELnr,4( I.KJ.CU I)
53 FOflMATUH . ^.F^.l.M'l.a.Fl?. t ,F1S.4)
21 rnMTlNlIt
24 inUM*.J
J»K
K-IOUM
IFCT-T^AX) 35>J2>32
32 IFtl-lKEPT) «1»«U.«1
41 CALL F*1T
SIIHKOUTINL INPUT
COMMON A01300J»HO( 100}»SO{ 3DO)»TRAVC 30(J),D( JOO).Il)(300)»ŁXP(300)»
lEXP«(300)»''»riKC50UJ' Jri( JOO) • A( 300.2 J .Cl ( 100J.AA?(300)>NTUT,ELM.
2l).THA,nTM»T,TMAx»CKl.()uiSO).LI »Cl l.tlN.ftEPI .OTH. MTOT »MO» fO. TE.
3AOTH.ALFA.LASI i.ljX.tL* UA». JJJ. TRAVKl 3UO) . rtKC
NTOT'FFIMC 1 )
ELM»FFIM( i>
THA«FFIN( 1 )
TMAX«FF1MC 1 )
CK1»FFIN( 1 )
NTlME-FFlNf 1)
00
-------
TABLE A2 (continued)---LISTING OF COMPUTER PROGRAMS FOR ESTUARY MODEL
on 3 I-I.LL
IDCI)«FFIN(l)
CONTINUE
10( I+l )«0
REPT-FFINU)
LAST«NTUT+1
RFTURN
EN!)
SUBKOUTlNfc INCON
COMMON A.tXP<300)»
1EXPR< 300) »OHKC500)»UH( 300) . AC 300.2) .Cl( 300) »AA2( 300) .NTiJT.ELMii
2UiTHA,l)TMpT,TMAXpCKl»nil(SO)»LL»Cll«ClN»KEPIi'DTH»MTOT»MO»TO»Tt»
3AnTri.ALF»»LASr.OX»tl.. UA». JJJ. f«AVKC3l>0».it.D) 12*12,13
12 R0( 1*1 )=1 35.0
li CONflNUe
SO(I)»SSS
TRAV(I)«TTT
D( I) "ODD
CKD-CCC
Y«Y+OX/5?80.0
1 CONTINUE
RETURN
ENO
SUBHOUTINt IMArtEA
COMMON AOMOO>»BO(300)>SO(30(»>TRAV(300).D(300>>IIH300>»EXP(300>*
1FXPH(300)>I1HKC>00}.IJH( JO 0).«( 300,2) »C1( 300), AA2t 300). N TUT. ELM.
2U,TriA,DTM.T.TMAX.CM.OUCiO).LL.C1 1 . C1N .REPT. 0 TH, MTUT . MO. TO, TE.
3AOTH.ALFA.LAST.OX.EL. OA?. JJJ. TRAVKC 300) .KKC
CALL TIDE
J»l
LASTP1«LAST»1
U«U/?280.
TRA(/(l)»RKC*EL/(AOTH*3<>00.)
EXP(l)«2.na28**(
TRAVRU)«T«AV(1)
EXP(LASTP1)»1.0
TRAV(LASTPl)«IKAVRtLASIPl ) •KXfPC I. AS IP 1 )»0.0
Y«DX/2.0
184
-------
TABLE A2 (continued)—-LISTING OF COMPUTER PROGRAMS FOR ESTUARY MODEI
TRAVX»(T-I>»/A!>TH»'1U
1N»LAST
DO 4 N*2.LAST
Z«EL+EL-Y
T»AV(lN)«rtKC*T/(AOTri*3oOO.)
Y.Y+OX
<» CONTINUE
DO S IN*1.LASlPl
TINC»TRAVXTRAV( IN)
M»T1NC
HTMC-M
OH(IN)»ŁXH(INJ*(UHK{rt}+(OHK(M+1)-U^K(M))*(TINt-HINC))
M»T1NC
OHR«EXPR(lN)*CUHKC'>l) + (UHKC'1*lJ-LiHKt-1))*(TIi
DH( IN)«i^
5 CONTINUE
RFTURN
COMMON Ani')00)»dU(3UU)>SO(3QO)*TftAV(30U)>0('iOO)>IU(JOO)»EXP(300J>
lEXPK(300)»l>MKCiOO)»LlH( JOO) , »{ 3uO. i>) .C 1 1 300 J . AA21 300) . NTUT . ELM.
2U.THA,OTM»T,TnAX.CKl»|)J(SO).LL»Cll.ClN.«FPr.I)TH.HTUT.m).TO>lE.
3AOTM.ALKA.LAS1 .OX»EL« UA?. JjJ.TRAVtU 30ID.KKC
IF(T-TE) l.^.i?
? TO»TE
TP«TO*(MTUT-MO)*AOTH
t TRAVX»( r-TOj/AUTH+MO
LASTP1*LAST+1
M«TINC
RINC«M
MaTINC
RINC«M
nHK*EXPR(l )*(DHK*{ f I NC-rtlNC ) )
DH( D»nH( i )+DH«
AC l.K)«AO( 1 )*UH( l
00 6 IN-?.LASTPl
TINC-TRAVXTRA^f IN)
M-TINC
RINC«M
TINC«TRAVX-TRAVH11M>
M.TINC
BTNC«H
) ) * ( T I^C-
A( IN.K)«AO(
AA21IN )»(A(
CONTINUE
AA?(?)«A( l.K)
RFTUBN
ENO
185
-------
TABLE A2 (continued)—-LISTING OF COMPUTER PROGRAMS FOR ESTUARY MODEL
SUrtHluTIML HUE
CfHMON A(K 3oO)»flOC3UO)»S(H3uO>»TKAVl30d),Ol J00)» IUC 300) • tXP< 300)
lEXP«(100)»i)HK(50U).Urt( JOO).AC300.i?>»Cl( 300) » AA2( 300 ) » NTU f >ELH.
2U.THA,OTMf T,Trt*X»CKl»OJ(SO).LL»Cll.ClN.REPT.bTH»HTOT»MO»TO»TE.
SAnTH.ALf A.LAST.DX.EL' UA?. JJ J. TKA VK ( U»n ) . HKC
Tn«r
MTOI«50
M0«18
TX»TO-1
DO 1 I^
DHK( I
Tx»TX+AOTH
1 CnNTI.MUE
RETURN
END
FUNCTION KLXIJI r j
Fl WUS950.0
RETURN
FNO
siiRKOuTiNt snsTt T.HA 1 1
IF(4.9-T) 1.2. Z
1 RATt-a.bFS
6H TO 3
? RATŁ«0.0
1 HFTURN
END
CARU CnilniT 2TH
186
-------
TECHNICAL REPORT DATA
(Please read Iwiifuctions on the reverse before completing)
FRPORT NO.
27
TITLE AND SUBTITLE
Tidal Flats in Estuarine Water Quality Analysis
3. RECIPIENT'S ACCESSIOf»NO.
5. REPORT DATE
June 1975
6. PERFORMING ORGANIZATION CODE
AUTHOR(S)
David A. Bella
8. PERFORMING ORGANIZATION REPORT NO.
EPA-660/3-75-025
PERFORMING ORG \NIZATION NAME AND ADDRESS
Department of Civil Engineering
Oregon State University
Corvallis, Oregon 97331
10. PROGRAM ELEMENT NO.
1BA025
11. CONTRACT/GRANT NO.
Grant 16070 DGO
12. SPONSORING AGENCY NAME AND ADDRESS
U. S. Environmental Protection Agency
National Environmental Research Center
200 S_W. 35th St.
Corvallis. Oregon 9733Q
13. TYPE OF REPORT AND PERIOD COVERED
Final
14. SPONSORING AGENCY CODE
15. SUPPLEMENTARY NOTES
16. ABSTRACT
The initial phases of the study involved mixing processes and tidal hydraulics; how-
ever, the study emphasis shifted to estuarine benthic systems as the importance of
these systems became more apparent. A conceptual model of estuarine benthic systems
was developed and a classification system of estuarine benthic deposits which is based
on the availability of hydrogen acceptors and reactive iron was developed.
Field studies demonstrated that estuarine sediments and overlying wastes could contain
significant concentrations of free sulfides which are toxic to a variety of organisms.
Field studies of benthic oxygen uptake and benthic sulfide release were conducted.
Water quality profiles within the deposits also were determined. A number of labora-
tory studies were conducted to determine the rate of sulfate reduction. Results from
experiments using extracts from benthic deposits and algal mats demonstrated a close
relationship between the rate of sulfate reduction and the sulfate and soluble organic
carbon concentrations. A general systems model of estuarine benthic systems was devel
oped. A variety of activities which could contribute to significant environmental
changes with estuarine benthic systems were identified.
Methods of determining dispersion coefficients from salinity profiles were examined
and an improved method was developed. The build-up of a pollutant in the vicinity of
the outfall during the slack water period of tide was studied through a field experi-
,ment and mathematical model
DS AND DOCUMENT ANALYSIS
DESCRIPTORS
estuaries
estuarine ecosystems
tidal hydraulics
mixing processes
benthic ecology
bottom deposits simulation
sulfides sulfide toxicity
b.lDENTIFIERS/OPEN ENDED TERMS C. COSATI Field/Group
Yaquina Estuary, Oregon
05C
18. DISTRIBUTION STATEMEN1
unlimited
19. SECURITY CLASS (ThisReport)
OF PAGES
20. SECURITY CLASS (Thispage)
22. PRICE
EPA Form 2220-1 (9-73)
A U. S. GOVERNMENT PRINTING OFFICE: 1975-699-188 133 REGION 10
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