Ecological Research Series
WATER QUALITY CRITERIA RESEARCH OF THE
U.S. ENVIRONMENTAL PROTECTION AGENCY
Proceedings of an EPA-sponsored Symposium
Environmental Research Laboratory
Office of Research and Development
.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 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 document is available to the public through the National Technical Informa-
tion Service, Springfield, Virginia 22161.
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EPA-600/3-76-079
July 1976
WATER QUALITY CRITERIA RESEARCH OF THE
U.S. ENVIRONMENTAL PROTECTION AGENCY
Proceedings of an EPA-sponsored Symposium on
Marine, Estuarine and Fresh Water Quality —
presented at the 26th annual meeting
of the AIBS, August 1975
compiled by
Technical Information Office
Corvallis Environmental Research Laboratory
Corvallis, Oregon 97330
CORVALLIS ENVIRONMENTAL RESEARCH LABORATORY
OFFICE OF RESEARCH AND DEVELOPMENT
U.S. ENVIRONMENTAL PROTECTION AGENCY
CORVALLIS, OREGON 97330
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DISCLAIMER
This report has been reviewed by the Corvallis Environmental Research
Laboratory, U.S. Environmental Protection Agency, and approved for
publication. Mention of trade names or commercial products does not
constitute endorsement or recommendation for use.
ii
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FOREWORD
The mission of the Environmental Protection Agency (EPA)
is primarily one of a regulatory nature, with responsibility
for establishing and enforcing environmental standards. The
establishment of standards must be preceded by, and based on,
sound, defensible research data. The Office of Health and
Ecological Effects within the Office of Research and Development
administers a variety of research programs to develop information
necessary for the establishment of such standards. The major
emphasis of this research has been directed toward the develop-
ment of scientific information for the establishment of water
and air quality standards. More recently, however, research
expertise increasingly is being directed toward the total
environmental picture (holistic or ecosystem approach), to
develop a sound basis for evaluating the ecological consequences
of all aspects of environmental pollution.
Proceedings of the two symposiums appearing in this text
were organized under the National Environmental Research Center
managerial mode. Subsequent reorganization has rendered each
laboratory (exception noted on title page of papers) an inde-
pendent entity with its own central research theme; in most
cases similar to the one under the NERC mode. Therefore,
questions concerning research activities of the EPA, in
research areas appearing in these proceedings, should be made
directly to the laboratories. Inquiries concerning other
aspects of water and air related ecological effects research
are invited. These may be made to the EPA Office of Health
and Ecological Effects, 401 M Street, SW, Washington, D.C.
20460 (RD-683).
oy/yAlbert, M.D.
Deputy Assistant Administrator
Health and Ecological Effects
Office of Research and Development
iii
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PREFACE
Early in 1975, the American Institute of Biological
Sciences (AIBS) invited the National Environmental Research
Center (NERC), Corvallis to participate in its 26th Annual
Meeting to be held on the Oregon State University campus
during 17-22 August 1975. A. F. Bartsch, Director of the
Center, accepted the invitation and Spencer A. Peterson,
NERC staff ecologist, was charged with organizing a one-
day symposium on the NERC research program.
At the time the invitation was extended, a diversity of
research work was being conducted by the nine laboratories
associated with NERC Corvallis. Research programs included
ecological effects of various pollutants on freshwater,
marine, and terrestrial ecosystems. Presentation of research
papers in all of these areas during one day was considered to
be impractical. Therefore, a decision was made to limit the
presentations to marine and freshwater research areas since
they were dominant at the time and to focus on water quality
criteria research. The symposium was divided into a fresh-
water and a marine segment. Each was designed to present a
cross-sectional representation of the types of research being
conducted at the NERC-associated laboratories and was not
meant to be all-inclusive. The freshwater program centered
around the transport and biological modeling capabilities of
the laboratories, cold climate aquatic biology, trophic
status of lakes in the Eastern United States, and the impact
of toxic substances on the freshwater environment. The
marine program centered on microbial and abiotic degradation
processes, the problem of trace metals, the effect of toxic
organics on the marine environment and the feasibility of new
stress-measuring methodology.
iv
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CONTENTS
Page
FOREWARD iii
PREFACE jv
MARINE AND ESTUARINE WATER QUALITY RESEARCH
OF THE ENVIRONMENTAL PROTECTION AGENCY
Structural Analysis of Stressed Marine Communities
R.C. Swartz, J.D. Walker, W.A. DeBen and F.A. Cole 3
Trace Metals in the Oceans: Problem or No?
Earl W. Davey 13
Persistence in Marine Systems
Kenneth T. Perez 23
Criteria for Marine Microbiota
V.J. Cabelli, A.P. Dufour, M.A. Levin, Paul Haberman 31
Impact of Chlorination Processes on Marine Ecosystems
D.P. Middaugh and W.P. Davis 46
Techniques to Assess the Effects of Toxic Organics on
Marine Organisms
David J. Hansen 63
The Effect of Subtle Temperature Changes on Individual
Species and Community Diversity
William C. Johnson II and Eric D. Schneider 77
FRESHWATER QUALITY CRITERIA RESEARCH
OF THE ENVIRONMENTAL PROTECTION AGENCY
Models for Transport and Transformation of Malathion in
Aquatic Systems
James W. Falco, Donald L. Brockway, Karen L. Sampson,
Heinz P. Kollig, and James R. Mauds ley 97
Shagawa Lake Recovery Characteristics as Depicted by
Predictive Modeling
D.P. Larsen and H.T. Mercier 114
A Mathematical Model of Pollutant Cause and Effect in
Saginaw Bay, Lake Huron
William L. Richardson and Victor J. Bierman, Jr. 138
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Page
Mathematical Model of Phytoplankton Growth and Class
Succession in Saginaw Bay, Lake Huron
Victor J. Bierman, Jr. and William L. Richardson 159
Implications of Resource Development on the North Slope
of Alaska with Regard to Water Quality on the
Sagavanirktok River
Eldor W. Schallock 174
Lake Eutrophication: Results from the National
Eutrophication Survey
Jack H. Gakstatter, Marvin 0. Allum and James M. Omernik 185
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MARINE AND ESTUARINE WATER QUALITY RESEARCH
OF THE ENVIRONMENTAL PROTECTION AGENCY
Eric Schneider, presiding
Director, Environmental Research Laboratory--Narragansett
Structural Analysis of Stressed Marine Communities
R.C. Swartz, J.D. Walker, W.A. DeBen and F.A. Cole
Trace Metals in the Oceans: Problem or No?
Earl W. Davey
Persistence in Marine Systems
Kenneth T. Perez
Criteria for Marine Microbiota
V.J. Cabelli, A.P.Dufour, M.A.Levin, Paul Haberman
Impact of Chlorination Processes on Marine Ecosystems
D.P. Middaugh and W.P. Davis
Techniques to Assess the Effects of Toxic Organics on
Marine Organisms
David J. Hansen
The Effect of Subtle Temperature Changes on Individual
Species and Community Diversity
William C. Johnson II and Eric D. Schneider
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Structural Analysis of Stressed Marine Communities
R.C. Swartz, J.D. Walker, W.A. DeBen,
and F.A. Cole*
ABSTRACT
Pollution often causes major changes in the structure
of marine communities. The impact of sewage sludge on
macrobenthic assemblages in the New York Bight and in
experimental microcosms is described as an illustration
of the effects of stress on species composition, density,
diversity and heterogeneity. Structure analysis provides
an exceptionally good method for assessing ecological
alterations at specific sites, but quantitative criteria
such as diversity indices should not be used as universal
regulatory standards. Field surveys should be closely
coordinated with laboratory investigations of the toxi-
city and accumulation of pollutants from those species
which dominated community structure and function prior
to human perturbation.
INTRODUCTION
Variations in species composition, density, diversity, and spatial -
temporal heterogeneity of multispecies assemblages are often used as
indicators of the effects of pollution on marine community structure.
Sometimes only one aspect of structure, usually diversity, is presented
as the sole biological criterion of stress. We will review the need for
more comprehensive structural analyses and their relationships with other
branches of pollution ecology, especially multispecies bioassays. For
illustrative purposes, data are given from field and laboratory investi-
gations of the effects of sewage sludge on the marine macrobenthos.
*Corvallis Environmental Research Laboratory, U.S. Environmental Protection
Agency, Newport Field Station, Newport, OR 97365
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METHODS
NEW YORK BIGHT SAMPLES
To illustrate different methods of structural analysis, two repre-
sentative samples were selected from a macrobenthic survey of the New
York Bight. Sample S was collected in the apex of the Bight near the
center of the sludge dumping site (40°23.8'N, 73°42.5'W) on 24 August
1973 and sample A at a relatively clean site approximately 23 km south
of Fire Island, New York (40°25.6'N> 73°11.TW) on 23 August 1973. The
samples were collected with a 0.05m2 Smith Mclntyre grab and sieved
through a 1.0 mm screen. Animals retained on the screen were preserved
in formalin and later identified to the species level.
INDICES OF COMMUNITY STRUCTURE
S ,
Density: Z n-/0.05m2
i=l 1
Diversity:
Area! Species Richness: S/0.05m2
Simpson's (1949) Index of Dominance (S.I.):
S n./n.-l\
1- S.I. = 1 - Z ]( 1 >
1=1 N(N-l)
Information Theoretical Diversity (Shannon-Weaver Equation):
i s
H' =^.(N log N - I nj log n^)
Fauna! Heterogeneity:
Czekanowski or Bray-Curtis Dissimilarity Index:
D.I. = 1 - W
where W = sum of lesser n-f/N for each species found in both samples
th
n-j = number of individuals of the i species
N = total number of individuals in a sample
S = total number of species in a sample.
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SLUDGE-BENTHOS EXPERIMENT
Digested municipal sewage sludge was obtained from the Bay Park
Sewage Treatment Plant, East Rockaway, New York. Sediments and test
specimens were collected in the lower Yaquina Bay, Oregon. The sedi-
ments were autoclaved and placed in a 3 cm deep layer at the bottom of
a polyethylene box (25 1 capacity, 1200 cm2 bottom area). The tanks
received a continuous flow of seawater entering at the surface at one
end and exiting through a stand pipe at the opposite end. Animals were
introduced to the tanks 48 hr before the sludge. The seawater was turned
off for 45 min while the sludge was added and allowed to settle. Dissolved
oxygen concentration reached a mimimum of 3 mg/1 during this period.
The same sludge sample was used in the first two experiments, con-
ducted 11-25 February 1974 and 1-15 April 1974. In the first, the poly-
chaetes Eupolynmia crescentis (5 individuals) and Glycinde polygnatha
(17); the molluscs Clinocardium nuttallii (11), Macoma nasuta (40), and
Transanella tantilla (25); and the amphipod Co r oph i urn sp i ni cp rn e_ (50)
were placed in two control tanks (no sludge) and three test tanks in
which sludge layers of 4, 20, and 45 mm were deposited. In the second
experiment the same number of individuals of all of the above species
except E_. crescentis were placed in two control and four test tanks re-
ceiving 1, 4, 5, and 8 mm layers of sludge. In both experiments sedi-
ments were sieved 14 days after the sludge was deposited and all living
individuals were recorded.
The third experiment was conducted 1-15 July 1974 with a different
sludge sample from the same treatment plant. The polychaete Glycinde
polygnatha (20 individuals); the molluscs Clinocardiurn nuttaTJii (10),
Macoma n"as~uta (40), Transanella tanti 11 a (15), and Cry^ptomya"caTifornica
(10); the amphipods Corophium spim'corne (50) and Paraph oxus epistomus
(17); and the cumacean Lamprops quadripiicata (6) were placed in two
control tanks and six test tanks each of which received a 4 mm sludge
layer. One of the test tanks was sacrificed 1, 2, 4, 7, 10, and 14 days
after the sludge was deposited. Both controls were sacrificed after
14 days.
RESULTS AND DISCUSSION
SPECIES COMPOSITION
The species composition, density, diversity, and dissimilarity of
macrobenthic collections at the sludge ground (station S) and a relatively
clean station (A) are given in Table 1. With the exception of Capitella
capitata, at least one individual of all species found at S was also pre-
sent at A. The fauna at S appears to be composed of the stress-tolerant
remnants of a more diverse benthic assemblage found in clean sandy sedi-
ments throughout much of the New York Bight. In particular, the gammarid
amphipods were abundant at A, but absent at S. The extreme dominant at
S, Capitella capitata, has never been found in several hundred grabs taken
in the vicinity of A.
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Total reliance upon indicator species as criteria of marine pollution
is undesirable, but the concept should not be ignored, the polychaete,
C. capitata. is an opportunistic species that can rapidly increase in abun-
dance during environmental disruptions. It is a cosmopolitan species often
found in areas of dredging, marine construction, cannery wastes, and sewage
deposits (Reish 1959, Wass 1967, Eagle and Rees 1973). Surveys by the
National Oceanic and Atmospheric Administration (1972) also indicated that
the absence of amphipods was a good indicator of the effects of sludge dumping
in the Bight.
TABLE 1. SPECIES COMPOSITION, DENSITY, DIVERSITY, AND
DISSIMILARITY OF MACROBENTHOS COLLECTIONS AT
THE SLUDGE GROUND (STATION S) AND A RELATIVE-
LY CLEAN STATION (A) IN THE NEW YORK BIGHT
Station S
Taxon
Number of
Individuals
Capitella capitata 916
Cancer irroratus 7
Unid. Nemertean 3
Cerebratulus sp. 2
Unicola irrorata 1
Nucula proxima 1
Total No. of Species 6
Total No. of
Individuals
930
Species Diversity (H1) 0.04
1- Simpson's Index
of Dominance
Station A
Taxon
Trichophoxus epistomus
Spi ophanes bombyx
A can th oh aus ton us" intermedius
Foraminiferan No. 1
Acanthohaustorius spinosus
Pseudunico1 a obTTquua
Unicola i/rorata
Cerebratulus sp.
Nucula proxlrna
Phyllodoce maculata
Acan th oh aus tori us_ mi1|si
Echi narcTi ni us parma
Unid. Nemertean
C1ymen_eJJ_a zonal is
Plus 11 separate species
represented by one
individual
0.03
Number of
Individuals
36
29
8
6
4
4
4
3
3
3
2
2
2
2
11
25
119
1.03
0.84
Dissimilarity Index between Stations S and A = 0.98
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_It is difficult to define the exact relationship between indicator
species and specific environmental factors. These species may not be
tolerant to all forms of pollution. The success of stress tolerant species
may be due to the elimination of less resistant competitors or predators
rather than a preference for altered habitats. The disappearance of stress
sensitive indicators may be caused by factors other than human perturbation.
However, if the more abundant organisms at a site are typically associated
with stressed ecosystems, it is probable that environmental conditions have
deteriorated. Wass (1967) suggested a pollution index based on the ratio
of the number of individuals of tolerant and intolerant species. Applica-
tion of the indicator concept at the multi-species level is preferable to
reliance upon a single species such as Capitella capitata.
DENSITY
The higher density of individuals at station S is due to the great
abundance of Capitella capitata. Excluding that species only 14 specimens
were collected at S, whereas 119 were found at A (Table 1).
Density comparisons based on the total number of individuals in entire
collections can be misleading if different phylogenetic and trophic assem-
blages are combined. Many of our samples from station A are dominated by
a very large (1-4 mm) arenaceous foraminiferan (Astrorhiza sp.?). It seems
unreasonable to include this species when comparing macrobenthic densities.
Changes in the abundance of individual or closely related groups of species
can be sensitive to minor ecological changes. Watling e_t aj_. (1974), for
example, found that sludge dumping off Delaware Bay had not caused serious
environmental damage, although the density of Nucula proxima, a deposit
feeder, had increased substantially due to organic enrichment of the sediment
DIVERSITY
Species diversity is a function of the number of species (richness)
and the distribution of individuals among the species (evenness) (Lloyd
and Ghelardi, 1964). This is a very broad ecological concept for which
a plethora of quantitative indices have been proposed.
Areal species richness or species density can be expressed as the
number of species (S) collected per unit effort or area. This is the
most basic concept of relative niche diversity. As a richness estimate,
species density is preferable to catch of species per unit number of
individuals (numeric species richness) because the latter is strongly
influenced by evenness patterns. Species density obviously is not an
estimate of the total number of species in a community and it is valid
only for comparative study. Constant sampling effort can usually be in-
corporated into survey designs and S's for different samples can be di-
rectly compared.
The degree of dominance by the more abundant species is an important
characteristic of the evenness of distribution of individuals among the
species. As dominance increases, "effective" diversity will decrease
even when the species density does not change. Simpson's (1949) index
gives the probability that two individuals drawn at random and without
replacement from a multispecies assemblage will belong to the same species.
7
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It is a good measure of dominance and its complement is positively related
to diversity.
To many ecologists species diversity implies an integrated measure of
both richness and evenness. The most popular index of "overall" diversity
is the Shannon-Weaver equation (H1). Because richness and evenness are not
necessarily correlated and have basically different theoretical significance,
H' may not always provide an adequate diversity analysis.
All aspects of diversity were substantially less at station S than at
station A (Table 1). Species density decreased from 25 to 6 species, H'
from 1.03 to 0.04, and the complement of Simpson's Index from 0.84 to 0.03.
These data demonstrate a major deterioration in benthic community structure.
The analysis of species composition and density indicate the differences in
diversity are due to the absence at S of many stress intolerant species and
the presence of a very large number of Capite11 a capitata.
FAUNAL HETEROGENEITY
The Bray-Curtis index is sensitive to differences in species compo-
sition and the relative abundance of individual species. The extremely
high value (0.98) between S and A clearly demonstrates the major differ-
ence between these assemblages (Table 1).
Analysis of faunal heterogeneity is more useful when applied to surveys
which include a large number of stations. Dissimilarity between all possible
pairs of samples can be calculated and the results expressed in dendrograms
which show hierarchial relationships between site clusters. Species clusters
can also be identified from the interspecies similarity of distributions be-
tween stations. A variety of dissimilarity indices and clustering strate-
gies are discussed by Clifford and Stephenson (1975). Boesch (1973) gave
a good example of the application of this kind of analysis in marine pollu-
tion research.
SLUDGE-BENTHOS EXPERIMENT
In the first two experiments, the survival of all species was reduced
when exposed to sludge layers >8 mm for two weeks (Table 2). Recovery of
living Corophium spinicorne, Clinocardium nuttallii, and Transanella
tantilla in test tanks receiving 4 or 5 mm sludge layers was substantially
less than in the controls. Only 4 C_. spinicorne survived when exposed to
1 mm of sludge while 21-37 of the 50 seeded specimens were recovered from
the control tanks. These results show that l') the response of the macro-
benthos is proportional to the quantity of sludge deposited, 2) all species
are not equally sensitive to sludge, and 3) a very thin layer of sludge
(1 mm) can affect some species over a short period of time (14 days). It
is not certain whether the impact is due to toxicity or indirect effects
such as burial or changes in sediment size distribution. The design only
crudely simulates the deposition of the settlable sludge fraction on the
bottom. Under field conditions, currents or wave action might keep a
larger proportion of the sludge in the water column.
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TABLE 2. SURVIVAL IN A MACROBENTHIC SPECIES ASSEMBLAGE EX-
POSED TO SEWAGE LAYERS OF 1-45 mm FOR TWO WEEKS
Individuals
Seeded
Species
Eupolynmia crescentis 5
Glycinde polygnatha 17
Corophium spinicorne 50
Macoma nasuta 40
Transanella tantilla 25
Clinocardium nuttallii 11
Individuals Recovered
Controls Sludge Layer (mm)
I II III IV 1 4 4 5 8 20 45
42*-- --2--00
15 17 16 14 13 10 15 12 8 6 1
35 42 21 37 4050110
39 37 39 39 37 35 19 29 9 1 1
18 21 24 21 20 15 2 3 0 0 0
99 11 11 7000000
*0nly 4 Glycinde seeded in Control II
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The third experiment was designed to determine if there was a tem-
poral pattern of mortality following exposure to a 4 mm sludge layer.
Interestingly, there was no substantial difference in survival of all
species between control and test tanks after 14 days (Table 3). Failure
to reproduce the results of the first two experiments might be attribut-
able to variations in the toxicity of sludge from the same treatment
plant since a different sample was used in the third experiment. If this
is true, source control of toxicants in some sewage sludges might signif-
icantly reduce their environmental impact.
These experiments demonstrate the feasibility of conducting multi-
species bioassays with ecologically important species collected from
ocean disposal sites. This permits a much closer coordination between
field and laboratory investigations than is possible with many conventional
bioassay organisms.
EFFICACY OF COMMUNITY STRUCTURE ANALYSIS IN POLLUTION RESEARCH
No single aspect of community structure provides an unequivocal
criterion of biotic response to stress. In particular, total reliance
on diversity indices should be discouraged because of their insensitiv-
ity to changes in the species composition. Further, the assumption that
diversity always decreases in response to ecological alteration is not
always true. Fish species diversity sometimes increases in the vicinity
of ocean outfalls and thermal discharges (Grimes and Mountain, 1971;
Turner, Ebert and Given, 1966). It is thus impossible to establish a
particular diversity index value as a universal regulatory standard for
an unacceptable level of pollution. However, diversity should be employed
as an important part of a more comprehensive investigation of community
structure.
Analysis of the condition of biotic assemblages in stressed areas
has several advantages over other methods in pollution ecology. The
biotic resource to be protected is examined directly. There is no un-
certainty about extrapolating laboratory results to field situations.
Structural characteristics such as species composition, density, diver-
sity, and heterogeneity are sensitive to individual and synergistic
effects of all forms of natural and pollutional stress, some of which
may not be immediately apparent. Spatial-temporal community patterns
can indicate both environmental impact and recovery following pollution
abatement.
We do not advocate community structure analyses as the "best" method
of assessing ecological alterations. They should be closely coordinated
with experiments on the toxicity and bioaccumulation of specific pollu-
tants. Similarly, bioassay organisms should represent the dominant taxo-
cenes which would occur in the absence of pollution at the stressed site.
Neither field nor laboratory studies by themselves can provide an adequate
basis for regulatory decisions.
10
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TABLE 3. SURVIVAL IN A MACROBENTHIC SPECIES ASSEMBLAGE EXPOSED
TO A 4 mm LAYER OF SEWAGE SLUDGE FOR 1 - 14 DAYS.
Species
Glycinde polygnatha
Corophium spinicorne
Paraphoxus epistomus
Lamprops quadriplicata
Macoma nasuta
Transanella tantilla
Clinocardium nuttallii
Cryptomya caliform'ca
Individuals
Seeded
Individuals Recovered
14 14
Controls
20
50
17
6
40
15
10
10
19
48
15
5
36
15
9
10
17
46
14
6
34
14
10
10
Exposure Time (da\
1 2
/
17
40
14
6
36
15
10
10
20
44
15
5
35
15
9
9
/s)
4 7 10 14
I mm Sludge Layer
20
37
13
4
34
13
6
10
14
43
14
4
30
14
9
10
17
31
14
3
33
13
9
10
17
40
15
6
32
13
8
9
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REFERENCES
Boesch, D. F. 1973. Classification and community structure of macro-
benthos in the Hampton Roads area, Virginia. Mar. Biol. 21: 226-244.
Clifford, H. T. and W. Stephenson. 1975. An introduction to numerical
classification. Academic Press, New York. 229 p.
Eagle, R. A. and E. I. S. Rees. 1973. Indicator species—a case for
caution. Mar. Poll. Bull. 4: 25.
Grimes, C. B. and J. A. Mountain. 1971. Effects of thermal effluent
upon marine fishes near the Crystal River steam electric station.
Florida Dept. Nat. Resources. Prof. Pap. Ser. No. 17. 64 p.
Lloyd, M. and R. J. Ghelardi. 1964. A table for calculating the
equitability component of species diversity. J. Anim. Ecol. 33:
217-225.
National Marine Fisheries Service. 1972. The effects of waste disposal
in the New York Bight. Final Report. Section 2: Benthic studies.
Sandy Hook Sports Fisheries Mar. Lab. Highland, New Jersey.
Final Rept. Feb. 1972, 277 pp. NTIS Rept. No. AD739532.
Reish, D. J. 1959. An ecological study of pollution in Los Angeles -
Long Beach Harbors, California. Allan Hancock Found. Occ. Pap.
No. 22. 119 p.
Simpson, E. H. 1949. Measurement of diversity. Nature 163: 688.
Turner, C. H., E. E. Ebert, and R. R. Given. 1966. The marine environ-
ment in the vicinity of the Orange County Sanitation District's
ocean outfall. California Fish & Game 52: 28-48.
Wass, M. L. 1967. Indicators of pollution. In Pollution and Marine
Ecology, T. A. Olson and F. J. Burgess (eds.). Interscience Pub!.,
John Wiley & Sons, N.Y., pp. 271-283.
Watling, L., W. Leathern, P. Kinner, C. Wethe, and D. Maurer. 1974. Evaluation
of sludge dumping off Delaware Bay. Mar. Pol. Bull.5: 39-42.
12
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Trace Metals in the Oceans:
Problem or No?
Earl W. Davey*
ABSTRACT
Increased input of mercury to the estuarine environ-
ment resulted in bioaccumulation in marine food chains
that affected man (Irukayama, 1966). Toxic effects of
other metals on marine animals have been demonstrated
under laboratory conditions. However, cause and ef-
fect between elevated environmental metals levels and
toxicity to marine animals has yet to be conclusively
demonstrated under field conditions. Municipal waste
water treatment plants, dredging and spoiling activi-
ties, and the dumping of sewage sludge and industrial
wastes are the major sources of metals to the marine
environment. These sources are likely to increase in
the near future unless the Federal Water Pollution Con-
trol Act Amendments of 1972 (PL-92-500) are carefully
enforced.
INTRODUCTION
Estuaries, because of the fact that they are landward extensions of
the sea, have become centers of industrial, commercial, and related activ-
ities. As a consequence, estuaries have received an increasing input of
metals due to the result of by-products of modern industry and technological
advancement. Metals can be introduced indirectly from contaminated rivers
and land runoff or directly by pumping from land based industries and muni-
cipalities, ship or barge discharges and aerial fallout (Merlini, 1971).
When viewed as a whole, ocean systems appear to be beyond compromise in
relation to its ability to dilute elemental introductions due to man's ac-
tivities -- after all, the continental masses are continually bathed in
their oceans and seas. Where then do problems occur?
*Environmental Research Laboratory, U.S. Environmental Protection Agency,
Narragansett, RI 02882
13
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Ocean waters, especially estuaries, are not uniformly mixed and
lack^of uniform dilution can cause local concentrations of metals.
Metals tend to be concentrated at air-sea, sediment-water, or fresh-
water-saltwater interfaces and boundaries between water and living or
dead particles (Fig. 1). Some metals discharged even in small quantities
can be accumulated to alarming and lethal levels by certain marine
biota. Seafoods harvested by man can become extensively impacted when
excessive metals are added to the sea. A classic example of the human
aspect of the problem first received considerable attention when mer-
cury poisoning occurred in Japan in 1953 through consumption of con-
taminated fish and shellfish (Irukayama, 1966).
It must also be recognized that it is not necessarily the total
amount of a metal present in seawater or marine sediments but the form
of the metal which may be important to consider with respect to the
effects metals may manifest on marine biota. Metals in seawater can be
operationally characterized as particulate (metals associated with par-
ticles larger than 0.45y) or dissolved metals (<0.45y). Dissolved metals
can be categorized further as in organically associated, organically
bound, i.e. chelated, or metal-organic compounds. Dissolved metal forms
are likely to interact with most marine biota; however, the effects may
differ if the metals are organically bound. If a metal such as copper
is chelated, there may be a reduction in metal toxicity response by or-
ganisms such as marine phytoplankton; whereas, if the metal is an organ-
ometallic like methy!-mercury, this compound is more toxic than the in-
organic form and can also be concentrated in food chains. Particulate
metals, probably occurring in high levels near industrial outfalls or
ocean clumping activities, are likely to affect filter feeding organisms
which ingest and concentrate particulate matter. Consequently, the form
of the metals may be the dictating factor in the response of marine biota
to heavy metals.
METHODS AND MATERIALS
Trace elements are essential to all life systems; yet excess amounts
are toxic. Also, non-essential elements such as mercury, cadmium, lead,
etc., can be toxicants and bioaccumulated to large quantities to affect
organisms within marine food chains including man. Therefore, a matrix
of existing toxicity and body burden data using marine species (including
various life stages) as one axis and metals (including various chemical
states and modes of application) as the other has been formulated in order
to assess, broaden, and validate the data base needed for criteria deci-
sion making. The metals matrix helps to point out information gaps, thereby
defining research goals; and it provides a basis for comparing metal levels
and their modes of application in laboratory toxicity and bioaccumulation
studies with levels and pathways defined in metal-problem areas in the
natural environment.
A summary of two metals matrices which were constructed mainly from
literature reviews by Ketchum et al. (1972), Eisler (1973), more recent
additions from the open literature, and in-house experiments performed
14
-------
Aerial fallout of PM
AIR-WATER
INTERFACE
SOM, PM
FOOD CHAIN
SEDIMENT-
WATER
INTERFACE
\
SM(lo)
SOM(lo)
PM(hi)-*-
sal>30%0
PM(lo)
SM(hi)
sal>IO%o
PM (hi)
SM(lo)
sal
-------
at the National Marine Water Quality Laboratory (NMWQL) are presented
in Tables 1 and 2.
RESULTS
The metals matrix indicates that there is information on only 36
elements and of these only 18 have toxicity data listed and of the 18
perhaps only four (Cd, Cu, Hg, and Zn) are sufficiently documented to
formulate good criteria. Since the NMWQL has had to respond to unex-
pected requests for elemental toxicity and bioaccumulation data, in
order to anticipate future requests, we have undertaken an in-house
program to develop acute and chronic marine bioassay information on a
wide spectrum of elements.
However, because there is infinite variety of combinations of marine
biota versus elemental compounds, a number of elements can be eliminated
from consideration in the following categories:
1. Elements such as mercury with sufficient information for good
water quality criteria
2. Major constituents of seawater s.a. Na, Mg, Cl, $0)4
3. Major constituents of marine organisms s.a. C, H, N, 0
4. Noble gasses, i.e. He, Ne, etc.
5. Elements which are short half-life isotopes
6. Rare earths
The remaining, approximately fifty elements, can be listed in priority
according to the following considerations:
1. Known toxicity to man
2. Information indicating elemental impact in the marine environment
3. The form of the element in seawater
4. No information available
On the basis of these considerations, elements are chosen for short-
term, acute bioassays. Acute bioassays involve a rapid response of a
single species to increasing concentrations of a toxicant. The results
of the acute bioassay are reported as the median tolerance limit (TLm of
TL5o) which signifies the concentration of toxicant that kills 50% of the
organisms within a specified time span, usually in 96 hours. Organisms
for acute bioassay are being selected from a wide range of representative
marine phylla and growth stages.
16
-------
TABLE 1. MATRIX OF ELEMENTS VERSUS MARINE BIOTA RESPONSE
Element
Aluminum
Antimony
Arsenic
Beryl lium
Cadmium
Chromium
Cobalt
Copper
Germanium
Gold
Iron
Lead
Manganese
Mercury
Nickel
Selenium
Silver
Yttrium
Zinc
Environmental
Oceans
Clean
0.01
0.0005
0.003
0.0000006
0.0001
0.000005
0.0001
0.003
0.00006
0.000004
0.0013
0.00003
0.005
0.00003
0.002
0.0004
0.00004
0.003
0.01
Spp. (Phyla)
Tested
4 (3)
2 (2)
6 (4)
1 (1)
34 (7)
13 (5)
1 (1)
48 (9)
2 (1)
1 (1)
1 (1)
14 (7)
2 (2)
43 (8)
17 (4)
5 (2)
9 (5)
1 (1)
28 (8)
Organism
Redfish
Algae
Copepod
Mummichog
Oyster
Algae
Copepod
Diatom
Diatom
Pinfish
Diatom
Ciliate
Oyster
Oyster
Algae
Copepod
Copepod
Oyster
Annelid
Most Sensitive Response
Level
88*
3.5
0.1
0.0001
0.015
0.0001
0.01
0.001
1.0
0.069
0.027
0.15
16.0
0.0056
0.0°D2
0.01
0.0033
0.001
0.05
Death
Inhib. cell div.
72hr LC5Q
Deer. enz. act.
Slow sex. devel .
Deer, culture yield
72hr LC5Q
Inhib. growth
Inhib. growth
Death
Cell clumping
Inhib. growth
LC50 of embryos
48hr LCcQ of emryos
Inhib. growth
96hr LC5Q
72hr LC5Q
Abnormal larvae (98%)
Abnormal larvae
* all concentrations expressed in mg/kg
-------
TABLE 2. MATRIX OF ELEMENTS VERSUS MARINE BIOTA BIOACCUMULATION
00
Element
Aluminum
Antimony
Arsenic
Barium
Beryl lium
Bismuth
Cadmium
Cerium
Ces i urn
Chromium
Cobalt
Copper
Gold
Iron
Lanthanum
Lead
Manganese
Mercury
Molybdenum
Nickel
Plutonium
Polonium
Rubidium
Ruthenium
Samarium
Scandium
Selenium
Silver
Strontium
Thorium
Tin
Titanium
Uranium
Vanadium
Yttrium
Zinc
Spp. (Phyla)
Tested
7 (1)
42 (10)
88 (12)
3 (1)
1 (1)
1 (1)
136 (12)
16 (5)
20 (6)
30 (4)
34 (7)
101 (8)
3 (1)
73 (8)
4 (2)
102 (7)
51 (5)
198 (15)
5 (3)
45 (6)
38 (7)
1 (1)
6 (3)
11 (7)
15 (3)
20 (-4)
11 (5)
18 (4)
18 (5)
5 (3)
2 (2)
6 (2)
1 (1)
6 (2)
2 (2)
130 (10)
Organism
Phytoplankton
Octopus
Squid (gills)
Phytoplankton
Phytoplankton
Phytoplankton
Abalone (digest, gland)
Fish
Algae
Zooplankton
Zooplankton
Squid (liver)
Mollusc
Annelid
Fish
Algae
Algae
Algae
Zooplankton
Zooplankton
Algae
Fish
Algae
Sponge
Annelid
Annelid
Octopus
Squid (liver)
Algae
Octopus
Phytoplankton
Phytoplankton
Fish
Pteropod
Mollusc
Mollusc
Level
Reached
5000*
0.92
198
262
8.4
7.7
1162.7
64
0.64
260
110
15,160
282
42,800
57
3100
226
7400
36
480
21,000 (CF) #
61 pCi/gm wet wt
2.3
10,000 (CF)
3.6
26.4
71
1044
4160
9.2
101
940
21
290
1000 uCi
99,220
Toxic
to
Man
_
+
+
-
+
+
+
-
-
+
-
-
_
_
_
+
_
+
_
_
+
+
-
_
_
_
+
-
.
+
-
-
+
-
_
-
* all values in mg/kg, except where noted
# CF - concentration factor
Toxic to Man: + yes; - no
-------
Elements having low TLso are in turn chosen for long-term chronic
bioassays. Chronic bioassays involve a continuous exposure to a sublethal
concentration of the toxicant. In the chronic bioassay, any biological
response, such as reduction of growth or reproduction, behavior change,
histopathological change, etc., can be used to monitor the effect of the
element or the species. Also, test organisms are analyzed to determine
possible bioaccumulation of the element which could in turn indicate a
potential pathway back to man.
DISCUSSION
A definite need exists to carefully inventory all natural and man-
made element sources which might impact the marine environment. Table 3
is a generalized inventory. Assessing potential ocean pollutants (Robinson,
et al., 1974) has presented an extensive and important approach for bud-
geting pollutants; however, the report deals only with the metals iron,
copper, and plutonium and concludes that plutonium is the only element of
potential global pollution. Similar assessment should be made for all
elements; however, these assessments should be focused at more localized
areas, such as coastal or estuarine areas as well as on a global scale.
These inventories would highlight elements of major environmental concern
which should be carefully bioassayed in the laboratory. Also, these
budgets should point out specific areas of high metal impact in the
United States.
Field investigations of metal impacted areas throughout the U.S. are
necessary in order to determine the extent, fate and effects of metals on
marine biota. Have metals per se directly or indirectly caused environ-
mental damage and, if so, to what extent? What are the inputs, rates,
routes, and reservoirs of metals within impacted areas? Special consid-
eration should be given to areas of:
1. Mining activities
2. Smelters
3. Industrial outfalls, especially metal plating industries
4. Sewage outfalls
5. Desalinization plants
6. Offshore ocean disposal areas for industrial wastes, sewage
sludge, and dredge spoils
However, Cross and Duke (1974) have emphasized that it is essential
that present efforts be continued and new efforts initiated to determine
baseline levels of trace metals in marine organisms and the environmental
variables that affect them. These studies should be conducted not only
in contaminated environments such as Long Island Sound, New York Bight,
and the Southern California Bight, but also in relatively pristine or
uncontaminated environments. The concentration of any trace metals can
19
-------
TABLE 3. INORGANIC CHEMICALS TO BE CONSIDERED AS POLLUTANTS OF THE MARINE ENVIRONMENT
World production
Element Natural cone in metric tons/year Routes of entry Pollution
_____ sea water uq/1 (1968) into the sea categories
H (acids) pH=8 (alk=0.0024) ? D,A III c
Be 0.001 250 U IV c ?
Ti 2 1,000,000 A ? IV b ?
V 2 9,000 A IV a ?
Cr 0.04 1,500,000 R(U) IV c ?
Fe 10 480,000,000 D,R IV c
Cu 1 5,000,000 D,R IV c
Zn 2 5,000,000 D,R III c
Cd 0.02 15,000 A,R II c
Hg 0.1 9,000 A,R I b
Al 10 8,000,000 D,R IV c
CN - ? D,R III c
Pb 0.02 3,000,000 A,R I a
P - ? D IV c
As 2 60,000 D II c
Sb 0.45 60,000 U IV c
Bi 0.02 3,800 U IV c ?
Se 0.45 1,000 U III c ?
F 1,340 1,800,000 D,R IV c ?
D dumping, A through atmospheric pollution, R through rivers (runoff) or pipelines
U unknown
I-IV order of decreasing menace; a worldwide, b regional, c local (coastal, bays, estuaries, single dumping).
Referenced from FAO Fisheries Reports, No. 99 Suppl. 1. Report of the seminar on methods of detection, measure-
ment and monitoring of pollutants in the marine environment: Inorganic chemicals, Panel 3. Dyrssen, D.,
C. Patterson, J. Ui and G. F. Wei chart
-------
be highly variable both within and between species and influenced by a
number of environmental variables. Until we understand the variability
that exists in healthy ecosystems, it may be difficult to identify a
contaminated ecosystem. Also, because trace metals occur naturally in
the marine environment as a result of weathering and volcanic activity,
the problem of determining the contribution of anthropogenic additions
of trace metals to natural levels in marine organisms is more difficult
than with halogenated hydrocarbons or refined petroleum products.
Other questions concerning potential metal pollutants which need
to be answered are as follows:
1. Are certain industries, such as power plants, producing exces-
sive metal inputs which should be controlled?
2. Can elemental transformations occur within marine areas to
produce more lethal and/or bioaccumulated compounds such as
methyl-mercury? If so, which elements are capable of trans-
formation and under what circumstances?
3. Dredge spoils removed from navigational channels are often
taken from areas which act as traps for sediments laden with
river and estuarine bourne waste. What are the long-term
effects of these ocean dumped dredge materials upon the
cleaner shelf areas? How should ocean dumped materials be
handled to lessen environmental impact in disposal areas?
4. Liquid effluents from waste water treatment plants probably
will be a major contribution of trace metals to estuarine
and coastal waters during the next several decades. Efforts
should be made to evaluate the impact that these discharges
will have on concentrations of trace metals in harvestable
marine species that complete a major portion of their life
cycle in coastal areas.
According to Schroeder (1973) environmental pollution by toxic
metals is a much more serious and insidious problem than is pollution
by organic substances such as pesticides, weed killers, sulfur dioxide,
oxides of nitrogen, carbon monoxide, and other gross contaminants of
air and water. Most organic substances are degradable by natural pro-
cesses; no metal is degradable. Elements in elemental form or as salts
remain in the environment until they are leached by rains into rivers
and into the sea. Therefore, every effort must be made to slow down
the environmental build-up of those elements which are toxic and can
cause degenerative diseases.
The solution to problems of metal waste disposal might be expected
to be dilution into the vastness of sea. However, because metals can
be concentrated by geological, chemical, and especially biological pro-
cesses in the sea, the solution to metal disposal problems is not dilu-
tion. The solution must be to stop pollution at its source by the
development of the proper technology to control and recycle metal wastes.
Hopefully, metal wastes entering the marine environment should be reduced
if application of the Federal Water Pollution Control Act Amendments of
1972 (Public Law 92-500) to apply the best available technology to mini-
mize environmental pollutants are carefully enforced.
21
-------
REFERENCES
Cross, F.A. and T.W. Duke. 1974. Trace metals. In: 6.V. Cox (ed.)f
Marine Bioassays. Marine Technology Society, Washington, D.C.:
308 p.
Eisler, R. 1973. Annotated bibliography on biological effects of
metals in aquatic environments. (No. 1-567) U.S. Environmental
Protection Agency Report R3-73-007; 287 p.
Irukayama, K. 1966. The pollution of Minamata Bay and Minamata
disease. In: Third International Conference on Water Pollution
Research. Water Pollution Control Federation, Washington, D.C.:
13 p.
Ketchum, B. 1972. Marine aquatic life and wildlife. In: Water
Quality Criteria 1972. National Academy of Sciences, Washington,
D.C.: 594 p.
Merlini, M. 1971. Heavy-metal contamination. In: D. Hood (ed.),
Impingement of Man on the Oceans. Wiley-Interscience, New York,
N.Y.: 738 p.
Robinson, E., L. Falk, B. Ketchum, and S. Piotrowicz. 1975. Metallic
wastes. In: E. Goldberg (ed,). Assessing Potential Ocean Pollu-
tants. National Academy of Sciences, Washington, D.C.: 438 p.
Schroeder, H.A. 1975. Elements and Living Systems. Plenum Press,
New York: 360 p.
22
-------
Persistence in Marine Systems
Kenneth T. Perez*
ABSTRACT
When various stressors and/or disturbances are applied
to a system, regulatory agencies are confronted with
the problem of determining what resulting systems
changes are "acceptable". In general, previous studies
have been arbitrary or unrealistic. We have attempted
to overcome the above inadequacies by: (1) viewing
systems as holistic, (2) assuming that some systems
can be miniaturized for experimental purposes, and (3)
attempting to define the persistence limits of a system.
Experimental microcosms simulating a complex marine
coastal system are described. Some preliminary results
of such systems to different artificial sewage stresses
are also presented.
Ecologists today more than ever before are being asked to predict
the consequences of changes in total systems caused by various disturb-
ances. A possible strategy for establishing limits for such changes is
the subject of this paper.
Conceptually, two general system responses are possible when a dis-
turbance of some intensity and duration is relaxed: the system either
recovers its "original" structural and functional state, i.e., it per-
sists or_ it does not recover, i.e., the changes due to the disturbance
are irreversible. This view is similar to that of Rolling (1973) but
more explicitly stated by Innis (1975). Given that the system persists,
one is also interested in the speed of recovery. Thus, the limits for
systems change proposed in this communique are the ability to recover
to some previously defined state or control condition. If system changes
were confined to the limits for recovery then the resource or system
would be maintained by definition.
*Environmental Research Laboratory, U.S. Environmental Protection Agency,
Narragansett, RI 02882
Co-investigators in this study are Scott Nixon, Candice Oviate and Jan Northby
nf +ha I In i \;QVC i" +\/ n-f Dhnrlci Tclan/H
of the University of Rhode Island.
23
-------
The establishment of persistence limits based upon previous theor-
etical studies is subject to question. The data base and principles from
which most complex models are formulated is derived from components- of a
system (May, 1973; Patten, 1971). These components are usually experi-
mentally isolated from the system as a whole before dynamical studies are
performed. The point is, if systems are holistic (Gallopin, 1971), then
even given a detailed knowledge of its parts will not enable the descrip-
tion of the total system. It had been shown (Walters & Efford, 1972)
that a complex model derived from isolated components of a system pro-
vided limited dynamical information.
Experimental field studies of the recovery of disturbed systems are
(1) few in number and (2) difficult to perform and interpret. First, it
is extremely difficult to impose or relax disturbances on natural marine
systems. Second, one is usually not allowed to jeopardize the system for
experimental purposes. Third, because of the problems in replicating
systems and/or knowledge of other uncontrolled disturbances, the estab-
lishment of cause and effect relationships is difficult.
Previous studies of laboratory microcosms have had major shortcomings.
To my knowledge, no persistence measures as described above have been made,
In fact, most of the microcosms lacked a full complement of species (see
Levandowsky, in press); those that did (e.g., Odum and Chestnut, 1970;
Whittaker, 1961) failed to provide sufficient physical properties (e.g.,
turbulence, water turnover) resulting in systems changes not observed in
similarly impacted areas in the field. One of the unique properties of
natural systems is their complexity. I would ask, "Why has this property
been minimized by the majority of microcosm studies?" Presented below is
an experimental marine system which attempts to overcome many of the in-
adequacies of previous studies.
Average Seasonal Light
Quality and Quantity
Narragansett
Bay Water
Benthic Tank
Pelagic Tank
•i i.
""' Rotating
Paddle
Exhaust
Air Pump to
Circulate
Pelagic Water
though the
Benthic Tank
Figure 1. Experimental microcosm.
24
-------
Our microcosms consisted of an interconnected pelagic and benthic
phase (Fig. 1). Pelagic water was continuously circulated over the
benthic community. The size of the pelagic phase (150 1) was dictated
by our resources. However, the surface area of the benthos (167 cm2)
was based upon the surface area to volume ratio in Narragansett Bay.
All container surfaces were scrubbed daily so that the only surface
area available for settling was the benthic sediments (i.e., the 167 cm2).
Natural changes in surface temperatures in the microcosms were reproduced
by continuously passing Bay water on the outside of the microcosms. Sal-
inity was monitored weekly. However, the West Passage of Narragansett
Bay, the system being simulated, exhibits small salinity changes (<2 o/oo)
Natural turbulence leyels in each tank were simulated by adjusting the
speed and reversing time of paddles such that the dissolution rates of
"sour ball" candy was approximately equal to that of the Bay (Table 1).
TABLE 1. DISSOLUTION RATES (gm/min) OF "SOUR BALL" CANDY
IN NARRAGANSETT BAY (NB) AND RHODE ISLAND SOUND
(RIS) DURING CALM AND WINDY PERIODS AND IN THE
EXPERIMENTAL MICROCOSMS
NB RIS Experimental
0 cm wave height 30-60 cm wave height microcosms
X 0.174 0.246 0.153
RGE 0.168-0.178 0.233-0.264 0.150-0.160
N 6 6 4
As a result of the continuous mixing, no differences in water chemistry
existed between the top and bottom of the tanks. This condition is simi-
lar to that found in the West Passage of the Bay. Water turnover (10 1/48
hrs) was based upon the flushing time of the Bay (Kramer, 1975) and was
accomplished by the removal and replacement of 10 liters of water every
48 hours. All water was hand-carried so as to eliminate mechanical damage
due to pumping. Light regimes were based upon the quality and quantity of
light found at 3 m, the depth at which the average light intensity for the
water column occurred in West Passage during early spring. Because light
is effectively extinguished on the bottom of the Bay, all benthic chambers
were dark.
The biotic composition of the microcosms was based upon densities per
unit volume for the pelagic phase and per unit area for the benthic phase.
This meant, for example, that fish larvae were experimentally excluded
from our systems since the highest densities found in West Passage was
5xlO~3/l. However, some larvae were observed in our tanks and probably
entered during the egg stage for we screened our water to lOOOy for pur-
poses of reducing replicate variability. It should be noted that the
inclusion of pelagic macroscopic forms such as ctenophores and fish larvae
25
-------
is possible; while ctenophores are major predators in the Bay they were
not present in any significant numbers during this time of the year (early
spring). The benthic organisms were collected from an anchor dredge,
screened of large living and dead material through V mesh, diluted with
seawater and mixed uniformly, allowed to settle in the benthic chambers
with flowing seawater for five days prior to place in the pelagic tanks.
Animal macrofauns (<1 mm) in the benthic chambers were counted at this
time and any large organisms not found in the boxes due to sampling error
were added in appropriate densities to each box. The seawater in the tanks
was intermixed for three weeks before any experiments were performed.
The objectives for our first experiment were to (1) compare different
variables in the microcosms to the field, and (2) to determine the persis-
tence (as defined above) of the microcosms exposed to fixed (i.e., time
independent) levels of domestic sewage (the disturbance). The sewage was
collected locally in large volume, partitioned to 500 ml fractions and then
frozen. The sewage was added, following the three week intermixing period,
to the tanks initially at three concentrations, 0.01, 0.1 and 1.0%; control
tanks received deionized water in proportions equal to the high sewage tanks.
However, the amount of sewage added thereafter only took into account the
48 hour seawater additions, i.e., the simulated influx water volume. There
were three replicates for each test condition and the control.
Some of the variables in the microcosms were measured continuously,
others discretely. The latter were the structural descriptions for the
benthos (macro- and meiofauna) and plankton. Specific size (1-50U) fre-
quency distributions of "particles" were made three times a week using-a
modified Coulter counter. Continuously measured variables in the water
column were ATP, nutrients (NH3, N03, N02, P04) and chlorophyll (in vivo
fluorescence and extracted). Measures of ATP, S=, C02, particulate organic
carbon (POC), and trace metals in sediments, and POC, DOC, dissolved or-
ganic carbon (DOC), total organic nitrogen, total metabolism, nitrogen,
and carbon fluxes in the pelagic phase are in progress. As a general com-
ment, we have found that the methods for measuring ATP in the sediments
and DOC, POC in the water column require close attention with respect to
accuracy.
The species richness of our microcosms prior to sewage addition was
representative of field levels using the above density criteria. The num-
ber of phytoplankton species ranged from 10-12. While we did not count the
number of holo and meroplanktonic species at the start of the experiment,
Martin (1964) averaged 19 such species during the early spring over a three
year period in West Passage. The benthic macrofauna consisted of 10 species.
The largest organism in the benthos was a polychaete worm, Nephtys, 4-5 cm
in length. The meiofauna (<1 mm) and bacteria were not enumerated at this
time.
We found a reasonable correspondence between the Bay and control ATP
levels (Fig. 2), thus suggesting realism of our microcosms with respect to
the pelagic bacteria and phytoplankton. The reason for the slightly higher
ATP concentrations in the controls as compared to the Bay during the inter-
mixing period (Table 2) is unknown; this relative difference remained, how-
ever, during the sewage exposure despite increases in Bay concentrations,
26
-------
-------
i.e., the controls appeared to follow the Bay but slight "constant" dif-
ferences existed. Differences between the sewage tanks treated alike also
existed. However, it was possible to detect significant (« = 0.05) non-
linear changes in ATP for all sewage levels over time. During the inter-
mixing period all tanks had equal ATP concentrations. A direct relation-
ship between ATP and sewage addition was observed most of the time following
the intermixing period. Chlorophyll concentrations (Fig. 3) showed similar
results although the differences between the Bay and the control tanks were
greater.
TABLE 2. AVERAGE ATP CONCENTRATIONS (ygm/1) FOR NARRAGANSETT
BAY (NB) AND LABORATORY MICROCOSMS PRIOR TO AND DUR-
ING EXPOSURE TO DOMESTIC SEWAGE
"Intermixing period" (17 days) Sewage exposure period (98 days)
x Sx n x Sx n
NB
control
0.01%
0.1%
1.00%
sewage
sewage
sewage
1
2
2
2
2
.54
.52
.40
.31
.40
0.1
0.1
0.1
0.1
0.1
35
78
19
14
68
5
7
7
7
7
2.
3.
4.
6.
8.
27
71
26
46
39
0
0
0
0
0
.200
.309
.194
.410
.334
42
43
43
43
43
The "sinks" for the high productivity in the pelagic phase were (1)
the 48 hour losses (i.e., flushing) and (2) the benthos. After 98 days of
sewage, the benthic communitites exposed to high sewage were characterized
by high organic layers at the sediment surface (36.9 mgC/1) as compared to
the controls (24.9 mgC/1). Polydpra lignae, a polychaete worm, occurred
in moderate density due perhaps to high sediment sulfide concentrations (a
sulfide bacterium almost completely covered the surface). It should be noted
that such a community was the result of organic loading and not low water
column oxygen levels, since oxygen concentrations near or exceeding satur-
ation (67-107%) of pelagic water was continuously circulated over the surface
of the sediments. The benthic communities treated with intermediate sewage
were dramatic; high densities of Polydpra were immediately evident. Tube
densities of 20-30/cm2 were "found. Presumably, lower organic loads with low-
er sediment anaerobiosis possibly either (1) provided "optimal" conditions
for Polydora to dominate and/or (2) excluded competitors and/or predators.
The real contrasts to the high and intermediate sewage levels were found in
the control and low sewage tanks. Almost no Polydora were present (1 tube/
cm2) in the control and low sewage tanks, but rather the initially stocked
community represented by Nephytys and Yoldia. As a result of these findings,
cursory surveys of the benthos along a sewage gradient were made in Narra-
gansett Bay. Areas suspected of high pollution were devoid of most living
material in the sediments. However, in slightly polluted areas densities
(5-25 tubes/cm2) of Pol.ydora similar to that found in our intermediate sew-
age microcosms were observed. Unpolluted locations showed benthic commun-
ities similar to that of the control tanks with Polydora at very low densities
28
-------
80
60
o>
NS
m
Q_
O
01
O
O
40
20
0
Bay wafer
Control tanks
Low sewerage
Medium
High
0
Figure 3.
f25
50 75
TIME,.Days
100
|l25
Mean pelagic chlorophyll ^changes in Narragansett Bay and experimental microcosms
prior to and during sewerage addition.
-------
(0-1 tube/cm2). This preliminary result suggests that the laboratory sim-
ulation of this marine system is somewhat realistic. Field determinations
of nutrient and organic loading and its relationship to benthic communities
is required before any further conclusions are made.
The main objective of this study is yet to come. Systems changes, as
described above, have occurred as a result of the disturbances applied.
The question now is, will the disturbed systems return to the structural
and functional state of the "control" systems once the sewage input is re-
laxed? If so, then what is the time course? — weeks, months, years? We
will continue to monitor the systems and follow their recovery since the
recent cessation of sewage.
REFERENCES
Gallopin, G.C. 1971. A Generalized Model of a Resource-Population system.
I General Properties Oecologia (Berl.) 7: 382-413.
Holling, C.S. 1973. Resilience and Stability of Ecological Systems.
Ann. Rev. Ecology and Systematics 4: 1-23.
Innis, G. 1975. Stability, Sensitivity, Resilience, Persistence. What
is of Interest? Proc. Siam-Sims Confer: S.A. Levin (Ed.) Ecosysterns
Analysis and Prediction, pp. 131-140.
Kramer, J. 1975. Analysis of a Plankton Based Temperate Ecosystem: An
Ecological Simulation Model. Univ. Rhode Island, Ph.D.
Levandowsky, M. In Press. Multispecies cultures and Microcosms. In
Marine Ecology (Ed.) 0. Kinne.
May, R.M. 1973. Stability and Complexity in Model Ecosystems.
Princeton Univ. Press, Princeton, N.J.
Martin, J.H. 1964. A study of the Relationships Between the Environment
and the Zooplankton of Narragansett Bay. URI, Masters, pp. 109.
Odum, H.T. and A.F. Chestnut. 1970. Studies of Marine Estuarine
Ecosystems Developing with Treated Sewerage Wastes, Annual Report
for 1969-1970. Institute of Marine Sciences, Univ. of North
Carolina, Chapel Hill and Moorehead City, N.C., 1-363 p.
Patten, B.C. (Ed.) 1971. Systems Analysis and Simulation in Ecology.
Vol. 1, Academic Press, N.Y., 607 p.
Whittaker, R.H. 1961. Experiments with Radiophosphorous Tracer in
Aquarium Microcosms. Ecol. Monbgr. 31: 157-187.
30
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Criteria for Marine Microbiota
V.J. Cabelli, A.P. Dufour and M.A. Levin*
and Paul W. Haberman**
INTRODUCTION
The examination of estuarine and coastal waters for microbial indi-
cators of quality can be and, in some instances, already is important in
assessing the ecological and human health impact of industrial, agricul-
tural and sanitary pollutants discharged therein. Such pollutants may
affect microbial activities in marine ecosystems in a number of ways.
Toxic organic and inorganic chemicals can destroy those microorganisms
responsible for essential biological transformations such as minerali-
zation, nitrification, etc. or those which act as food sources for higher
life forms. Nutrients discharged into the water from sewage or industrial
wastes or man-induced changes in the physical environment may permit the
multiplication of pathogens for fish or other marine fauna. On the other
hand, toxicants and changes in the physical conditions may affect the sus-
ceptibility of the fauna to microbial infection and disease. The activi-
ties of marine microorganisms themselves are important in that some species
degrade pollutants such as pesticides, detergents and other organic molecules,
some accumulate pollutants and pass them up through the food chain, and some
activate certain chemicals to more toxic forms, i.e., methyl mercury.
Unfortunately, in most of the interactions noted above, the relation-
ships among the pollutant, the adverse ecological effect and some microbial
indicator thereof have not been quantified sufficiently; and, hence, micro-
bial guidelines and standards are not available. This, of course, is under-
standable because of the complex interrelationship involved.
Pollution associated effects on human health are more amenable to quan-
tification. There are a number of microbial pathogens of man which can be
discharged into receiving waters via fecal wastes or which can multiply
therein under the influence of nutrient pollution. However, in terms of
microbial activity, there is only one target species (man), two routes of
transmission (recreational use of marine waters and consumption of marine
biota as food), and two basic processes (infectious disease and bionitoxi-
cation). The above factors notwithstanding, a major constraint in quanti-
fying health effects and modeling dose-response relationships stems from the
logistic and ethical restrictions placed on studying the target species, man.
^Environmental Research Laboratory—Narragansett, U.S. Environmental Protection
Agency, West Kingston, RI 02881
**Center for Policy Research, New York, NY
31
-------
Because of the above considerations, the directness and immediacy of
the impact on man via his health and well being, and man's egocentricity
as a species, it is not surprising that human health effects guidelines
and standards for water quality generally have antedated those for ecolo-
gical effects by a number of years. Furthermore, it is both understandable
and defendable that such microbial guidelines and standards have been set
forth based upon a limited quantity of epidemiological data. The reality
is that the construction of waste disposal systems will continue and that
receiving water criteria are needed so that they may be translated into
effluent guidelines and standards, including the degree of treatment re-
quired, the siting of sewage outfalls, and the locating of sludge disposal
areas.
As used in the studies to be described, a health effects, recreational
water quality criterion may be defined as a scientific or objective require-
ment of a condition to be fulfilled for the protection and promotion of the
health and safety of the public. It is a set of facts or data upon which a
decision or judgement may be based. Many such criteria are developed through
intensive and extensive epidemiological studies1'2. The criteria may be used
in two ways. They can be extrapolated into guidelines and standards for post-
ing or closing beaches as "unsafe" for use. A more desirable use of the
criteria, since it increases rather than decreases the available resource, is
their translation into effluent guidelines and standards as noted above.
Figure 1 is a graphic representation of the health effects, recreational
water quality criteria as defined above. It is a slightly modified version
of that described by Cabelli and McCabe . The salient point is that a cri-
terion is expressed as a quantitative relationship between some index of health
effects among swimmers to some measure of the quality of the water. Once an
"acceptable risk" is determined, an appropriate guideline can be extrapolated
from the criterion. The setting of an "acceptable risk" has social and eco-
nomic implications as well as the health effects input inherent to the criter-
ion. Therefore, the two translations may, and probably will, have different
"acceptable risk" levels and, hence, different guidelines and standards.
Existing criteria, guidelines, and standards at the local, state and
national level have been reviewed elsewhere4 as have the data base required
for their development, the types of experimentation or observation which can
be used to produce the data, available Information, and the shortcomings of
this information2'5. As of 1969 when the present study was designed or even
1972 when it was initiated, the available epidemiological data came primarily
from Stevenson's prospective studies6, Moore's restrospective study7 and some
scattered case reports8'9. The analysis of a recreation associated outbreak
of shigellosis along the Mississippi River below Dubuque, Iowa10 provided
some additional data. For a number of reasons none of these studies indi-
vidually or together provided the totality of the required epidemiological
and microbiological data.
The United States Environmental Protection Agency has responded to the
need for additional data by a long term epidemiological-microbiological
program whose objective is to define the relationships noted earlier. The
overall program calls for (1) a three year study at selected New York City
beaches to define the relationships (including a pretest of the epidemio-
logical and microbiological methods), (2) trials at some "subtropical
32
-------
B -
acceptable risk
background illness
WATER QUALITY INDICATOR DENSITY
Figure 1. Graphical illustration of health effects water quality criteria and their
extrapolation to guidelines.
-------
location to examine the variable of climate, (3) the development of a math-
ematical model (the regression line and confidence limits of the relation-
ship described in Figure 1), and (4) spot testing of the model at a number
of geographically distinct sites. The findings to be presented summarize
the data obtained during the first two years of the program.
EXPERIMENTAL DESIGN
The study being conducted at the New York City beaches is a prospec-
tive epidemiological investigation in which (1) the potential participants
(primarily family groups) are approached at the beach in the course of week-
end trials, and individuals who swim in the midweeks immediately before and
after a trial are eliminated from the study, (2) swimming is rigorously de-
fined as significant exposure of the head and face to the water, (3) measure-
ments for a number of potential water quality indicators are made during the
course of the trials at the test beaches, and (4) follow-up information con-
cerning symptomatology and demography is solicited by phone some 8-10 days
after a trial (Table 1). The experimental design and the rationale for its
TABLE 1. SEQUENCE OF EVENTS FOR EPIDEMIOLOGICAL-
MICROBIOLOGICAL TRIALS
Day of Day
Week Number
Acti vi ty
Function
Saturday 1
(Beach Interview, Water
Sampling)
Sunday 2
Monday 3
Monday 10
(Beach Interview, Water
Sampling)
Reminder Letter
Phone or mail Interview
(a) Obtain name,address, phone, etc.
(b) Reject pretrial midweek swimners
(c) Query on beach activity
(d) Assay of water samples
As Above
(a) Provide name of physician
(b) Reminder to note illness
(a) Obtain illness information
(b) Reject post-trial midweek
swimmers
(c) Obtain demographic information
34
-------
use have been described elsewhere5. In the first two years of the New
York study two sites (beaches) were used. The first beach, located at
Coney Island around 22-24th Street (Figure 2), was desianated as "barely
acceptable" (BA) and was the most polluted beach available which was not
posted as unsafe for swimming. The second, located at Arverne or Riis
Park at the Rockaways, was designated as "relatively unpolluted" (RU) and
was the least polluted beach available at which the populations were demo-
graphical ly similar to the BA beach. Thereby, attack rates for symptoms,
symptom groups (i.e., gastrointestinal, respiratory, "other") and a "sev-
erity index" (stayed home, stayed in bed, sought medical advice) was ob-
tained for the four groups (swimmers and nonswimmers at both beaches) and
for the various demographic subgroups (sex, age, ethnicity, socioeconomic
status). The symptoms for which information was solicited are given in
Table 2. The soliciting of health effects information in the context of
symptoms rather than specific disease is consistent with the first basic
tenant of the experimental design, i.e., there would be no prejudgement
as to which diseases were "important" in the context of swimming assoc-
iated health effects.
TABLE 2. SYMPTOMS FOR WHICH QUERIES WERE MADE
Gastrointestinal
Vomiting
Diarrhea
Stomachache
Nausea
"Other"
Fever (>100°C)
Headache (more than few hours)
Backache
General
Sunburn
Skin rash, itching skin
Red, itchy, or watery eyes
Respiratory
Sore throat
Bad cough
Chest cold
Runny or stuffed nose
Earache or runny ears
Sneezing, wheezing, tightness
in chest
"Severity" Index
Home because of symptoms
In bed because of symptoms
Medical help because of symptoms
The second tenant of the study was that the "correct" indicator would
be treated as an unknown; this required density measurements and, at times,
the development of enumeration methods for a number of potential water
quality indicators. Water samples, used to obtain the density measurements,
were collected at "chest level" from two sites at each beach about every two
hours during the period of maximum swimming (11:00 a.m. to 5:00 p.m.). A
number of potential indicators are listed in Table 3 along with designations
as to those for which measurements have or will be made. A review of po-
tential health effects water quality indicators and the methods used for
35
-------
u>
UPPER
BAY
BROOKLYN
LOWER
BAY
01234
J*
20th St.
I I I
KILOMETERS
COME
ATLANTIC OCEAN
Figure 2. Test beaches at Coney Island ("barely acceptable") and the Rockaways ("relatively
unpolluted") in New York City.
-------
their enumeration is beyond the scope of this report. This h
by Bonde'l. Such a review, including usage rationales is bei
for publication. Papers describing the enumerative
developed for this program have been published '2-179
preparation.
has been done
ng prepared
methods used in or
and others are in
TABLE 3. POTENTIAL HEALTH EFFECTS WATER QUALITY INDICATORS
Indicator
Status
Indicator
Status
a -
b -
C «•
d -
e -
Total Col i forms
Fecal Col i forms
E. coli
Klebsiella
Enterobacter-
Citrobacter
Enterococci
C. perfringens
C. albicans
Bifidobcateria
examined in 1973
examined in 1974
to be examined i
may be examined
examined in 1973
a
a
a
a
a
a
b
c
c
and 1974
n 1975
in future
, discontinued
Enterovi ruses d
Coliphage c
Salmonella
Shi gel la d
P. aeruginosa a
A. hydrophila a
V. parahemolyticus b
RESULTS
The pretest (Phase I) trials conducted during the summer of 1973 con-
firmed the applicability of the epidemiological and microbiological method-
ology5. Although the study population contained only 1300 individuals, a
statistically significant increase in the rate of gastrointentinal (GI)
symptoms for swimmers relative to nonswimmers was observed at the "barely
acceptable" but not the "relatively unpolluted" beach (Table 4). Cochran's
chi square method as descrived by Fleiss'8was used for the statistical
analysis. Increases in respiratory, "other" and "severe" symptoms also were
obtained at the Coney Island beach, but these were not statistically signi-
ficant at the P=0.05 level. With the exception of respiratory symptoms,
smaller increases (swimmer minus nonswimmer) were obtained at the Rockaways
beach. The microbiological findings were described previously^.
37
-------
TABLE 4. SYMPTOM RATES BY CATEGORY FOR 1973
Symptom Rate in Percent at
Symptom Type Coney Island Rockaways
S NS A s MS
474 167 484 197
Resp. 12.9 10.2 2.7 18.0a'b 11.7 6.3
GI 7.2 a 2.4 4.8 8.1 4.6 3.5
Other 9.9 6.6 3.3 9.1 8.6 0.5
"Severe" 5.9 4.2 1.7 6.0 5.6 0.4
^Significantly (P = 0.05) higher than nonswimmers.
Significantly (P = 0.05) higher than other beach.
S-Swimmers; NS-nonswimmers; ^-difference; Resp.-respiratory; Gl-gastro-
intestinal; Other-general symptoms; "Severe"-stayed home, stayed in bed
or sought medical help.
The 1974 findings (Table 5) essentially confirmed the 1973 results.
TABLE 5. SYMPTOM RATES BY CATEGORY FOR 1974.
Symptom Rate in Percent at
Symptom Type
Resp.
GI
Other
"Severe"
Coney Island
S NS A
1961
7.2
4.2 a
7.3
3.8
1185
6.4 0.8
2.6 1.6
6.7 0.6
2.9 0.9
S
2767
8.3
3.9
8.6
3.0
Rockaways
NS A
2156
7.8
3.5
7.7
2.6
0.5
0.4
0.9
0.4
Significantly (P=0.05) higher than nonswimmers.
Significantly (P=0.05) higher than other beach.
S-Swimmers; NS-nonswimmers; A-Difference; Resp.-respiratory; Gl-gastro-
intestinal; Other-general symptoms; "Severe"-stayed home, stayed in bed
or sought medical help.
38
-------
Although the differential rates for most of the symptom categories were
lower in 1974, because of the larger study population a statistically
significant increase in the rate of 61 symptoms among swimmers relative
to nonswimmers again was obtained at the "barely acceptable" beach. In
addition, the most sensitive portions of the Coney Island population were
identified as children, Latin Americans and low to middle socioeconomic
status individuals. Finally, the validity of the responses obtained as
to gastrointestinal symptomatology was examined by calculating the rates
of those symptoms considered highly reliable (all instances of vomiting;
diarrhea only when "severe" or with fever; stomachache and nausea only with
an accompanying fever) for each of the subgroups and comparing these to
overall GI rates for the corresponding groups. The trends, and in most
cases the statistical significance, were comparable. These findings were
important in demonstrating that certain differences probably were not spur-
ious, in confirming the acceptability of the methodology, and in identify-
ing the sensitive portions of the population for future studies. In addition,
there are aspects and implications of the data relating the health effects
of swimming per se, year to year variability in the nonswimming ("Background")
rates of gastrointestinal symptomatology, and demographic differences in re-
porting symptoms which will be considered in a later publication. However,
this type of analysis does not speak to the overall objective of the program,
that is, defining the relationship or association of health effects to water
quality indicators as described in Figure 1.
In the context of the present experimental design, the data can be
analyzed to yield the criteria in two ways. It can be obtained from re-
gression analysis of the data obtained during a given summer by consider-
ing the symptom rates and the corresponding indicator densities for each
trial (day) as a single point on the line; in this instance, one capitalizes
on temporal (day to day) and spacial (distance) variability in pollution
levels reaching the beaches. The second approach is to analyze the data
across summers. Thereby, the overall symptom rates and associated indica-
tor densities for all the trials at each beach during a given summer are
combined to yield a single data point.
Correlation coefficients for the differential rate of gastrointestinal
symptoms against the various water quality indicators were obtained from the
1974 data as shown in Table 6. The regression lines for £. coli, Klebsiella
and fecal coliforms, the indicators with the highest correlation coefficient,
are shown in Figure 3. Another set of regression lines should be obtained
from the 1975 data.
When the data were examined across summers, four points were obtained
for each indicator (Figure 4). Since an additional two points should be
obtained from the 1975 trials, a statistical analysis was not attempted.
However, inspection alone confirms the close relationship of GI symptoma-
tology to _E. coli densities, although fecal streptococci, Klebsiella and
Aeromonas hydrophila also produced close fitting lines. The regression
lines for both total and "severe" GI symptoms against E_. coli densities
are shown in Figure 5. It is of interest that the "IE. coli" lines ob-
tained by both methods of analysis were quite similar; the differential
rates for total GI symptoms associated with mean E_. coli densities of
200/100 ml were 3.8 and 3.6%.
39
-------
4 .
JS
o
zo 0
OK
?i 4
^-: 2
_i
< cc
^£ o
a: ^ 4
U. L_
t< «
o o: *
Klebsiello sp.
Fecal Conforms
o
o
E. coli
10 100
MEAN DENSITY/100 ml of WATER, PER TRIAL
1000
Figure 3.
Relationship of the differential rate of gastrointestinal symptoms (swimmers minus
nonswimmers) to the mean density of the water quality indicators as obtained from
the trial by trial analysis of 1974 data. The lines were drawn from a least squares
analysis. The three regression lines are for the three indicators which gave the
best correlation coefficients: E_. col i, 0.771; fecal coliforms, 0.673; Klebsiella,
0.664; »-Rockaways, A-Coney Island.
-------
10
2 o
^c
>2
3>o
z i-
o°- 10
Z5
±z
*- 0
c/>~
^e>
<§•
•-a. K)
z_
WS2
or2"
£u 0
u- |_ U
lfb<
o o:
10
o
Enterobocter- Citrobocter
A
. ° A
Total coliforms
^
-
A.hvdroDhila A 1973 CONEY ISL.
o 1973 ROCKAWAYS
A 1974 CONEY ISL.
o 1974 ROCKAWAYS
j*^* — ^^
P.aeruginosa
: . ° A*
Fecal coliforms Klebsiella
-
^
o
• A
Fecal streptococci
-
^
< 1^1 i 1 1 1 1 1 i i i i 1 1 1 1 1 i i i 1 1 1 1 1
1 10 100
A __
Q ^
A • • " A
E. coli
^tL— *"
L^-.— r-i i 1 1 1 1 A i it i i 1 1 1 1 i i i i i i 1 1
10 100
MEAN INDICATOR DENSITY PER 100 ml
Figure 4. Relationship of the differential rate of gastrointestinal symptomatology to indi-
cators densities as obtained from the analysis of 1973 and 1974 data. Each point
represents the overall GI symptom rate and mean indicator density for all the
trials conducted at the beach during that summer.
-------
50
30
to
COT
-------
TABLE 6. RELATIONSHIP OF INDICATOR DENSITY TO THE
DIFFERENTIAL (SWIM-NON-SWIM) ATTACK RATE
FOR GASTROINTESTINAL SYMPTOMS (1974)*
Indicator Correlation Coeff.(r)
I- Co1i 0.711
Klebsiella 0.664
Fecal Coliforms 0.673
Total coliforms 0.549
Fecal streptococci 0.453
Pseudomonas aeruginosa 0.191
*0btained from six trials at Coney Island (BA) and Rockaways (RU) beaches.
The high attack rate of 3-4% above background (nonswimmers) appears
to be somewhat disturbing. However, it must be borne in mind that the
individuals in question neither died or required hospitalization. In all
probability most of these cases would not have been reported to public
health authorities except in an "outbreak" situation. Nevertheless, they
do represent a measureable and, hopefully, predictable health effect whose
economic and social import must be considered in setting an "acceptable
risk". As noted previously, additional data should be forthcoming from
trials being conducted at the New York City beaches during the summer of
1975 and at a "subtropical" site the following year. Therefore, a complete
statistical analysis of the health effects vs indicator data obtained to
date was not attempted; and the information to follow is meant to be des-
criptive and indicative of the relationships which can be expected.
The overall program to develop health effects recreational water
quality criteria is far from complete. However, the data obtained thus
far are quite encouraging. E_. coli and fecal streptococci appear to be
the best indicators examined thus far; and, if the data being obtained
this summer are consistant with the findings to date, interim criteria
should be available in 1976. "Subtropical" site selection is in progress;
and the nature of the model is envisioned, that is, the regression line of
GI symptomatology vs E_. coli or fecal streptococci densities and the con-
fidence limits around the line.
-------
REFERENCES
1. Mood, E.W. Health Criteria for the Quality of Coast Bathing Water.
Report of the Special Advisory Committee on Health Criteria for
the Quality of Recreation Waters with Special Reference to Coastal
Waters and Beaches, World Health Organization. December, 1974.
2. Cabelli, V.J., M. A. Levin and A. P. Dufour. Recreational Water Quality
Criteria. Submitted to Critical Reviews in Environmental Control.
3. Cabelli, V.J. and L.J. McCabe. Recreational Water Quality Criteria.
News of Environmental Research in Cincinnati. November 11, 1974.
4. Mechalas, B.J., K.K. Hekimian, L.A. Schinazi and R.H. Dudley. 1972.
An investigation into recreational water quality. Water Quality
Criteria Data Book, Volume 4. 18040 DAZ 04/72 Environmental Pro-
tection Agency.
5- Cabelli, V.J., M.A. Levin, A.P. Dufour and L.J. McCabe. 1974. The
Development of Criteria for Recreational Waters. International
Symposium on Discharge of Sewage from Sea Outfalls.
6. Stevenson, A.M. 1953. Studies of bathing water quality and health.
Jour. Am. Pub. Health Assn. 4_3:529.
7. Moore, B. 1959. Sewage contamination of coastal bathing waters in
England and Wales. A bacteriological and epidemiological study.
J. Hyg. 57_:435.
8- Flynn, M.J. and D.K.B. Thistlewayte. Sewage Pollution and Sea Bathing.
Second International Conference on Water Pollution Research, 1964.
9- Ciampolini, E. 1921. A study of the typhoid fever incidence in the
health center district of New Haven. Unpublished Report.
10- Morbidity and Mortality Weekly Reports. Shigellosis Associated with
Swimming in the Mississippi River, National Center for Communicable
Disease, U.S.D.H.E.W., Vol. 23, No. 46, 1974.
11. Bonde, G.J. 1966. Bacteriological Methods for Estimation of Water
Pollution. Hlth. Lab. Sci. 3i, 124.
12. Levin, M.A. J.R. Fischer, and V.J. Cabelli. 1974. Quantitative Large-
Volume Sampling Technique. Appl. Microbiol. 28, 515.
13. Cabelli, V.J. 1973. The Occurrence of Aeromonads in Recreational Waters,
Abst., ASM. p. 32.
14. American Public Health Association. Standard Methods for the Examination
of Water and Wastewater. 13th ed, 1971.American Public Health
Association Inc., New York.
44
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15. Levin, M.A. and V.J. Cabelli. 1972. Membrane filter technique for
enumeration of Pseudomonas aeruginosa. J. Appl. Microbiol. 24:864.
16. Dufour, A.P. and V. J. Cabelli. 1975. Membrane filter procedure for
enumerating the component genera of the Coliform group in seawater.
J. Appl. Microbiol. 29_:826.
17. Levin, M.A., J.R. Fischer, and V.J. Cabelli. 1975. Membrane filter
technique for enumeration of Enterococci in marine waters. J. Appl
Microbiol. 30;66.
18. Fleiss, J.L. Statistical Methods for Rates and Proportions. John Wiley
and Sons. 1973. New York City p. 109-114.
-------
Impact of Chlorination Processes
On Marine Ecosystems
D.P. Middaugh and W.P. Davis*
ABSTRACT
The use of chlorine as a disinfectant and antifoul ing
agent is reviewed. Chemical reactions of chlorine in
aquatic environments are discussed, with particular
emphasis on the formation of halogenated organic con-
stituents in freshwater and marine systems. Studies
of the effect of chlorinated sewage effluents and cool-
ing water from generating stations on marine organisms
and ecosystems are summarized.
INTRODUCTION
Chlorine gas has been used as an industrial bleaching agent since
1800 and has become one of the most versatile chemicals known. In fresh-
water it is used in drinking and recreational water as a disinfection
agent, a biocide for slime and fouling control, and in the treatment of
municipal wastes to control pathogens. In these applications, vast quan-
tities of chlorine are used, and find their way into natural ecosystems.
The toxicity desired in disinfection and biocide applications can contin-
ue on with undesirable effects on wildlife and the environment. Recent
detection of halogenated organics in the drinking water of 80 cities un-
derscores the need for responsible assessment of the management and effects
of our chlorination processes, and the environmental costs incurred.
Some of the most accurate statistics on the rate of chlorine use exist
for the State of Maryland. Chlorine discharge from Maryland into the Chesa-
peake Bay is presently estimated to be 1.1 x lO1*5 g/yr from municipal sewage
treatment plants and 0.1 x 1010 g/yr from power generating facilities (Block
and Helz, 1975). It is estimated (but still to be confirmed experimentally)
that 1 percent of these totals may become halogenated organic compounds
which would persist in the environment (Jolley, 1973). Thirty-three other
states border marine ecosystems where some form of chlorine discharge cur-
rently persists.
*Environmental Research Laboratory—Gulf Breeze, U.S. Environmental Protection
Agency, Bears Bluff Field Station, Wadmalaw Island, SC 29487
46
-------
The purpose of this paper is to compile the scarce data presently
available for chlorine effects upon aquatic life relevant to estuarine
and marine ecosystems. The chemistry of chlorine is briefly reviewed
to point out some of the unique features of chlorination in marine waters.
Although some data exist on the effects or residual chlorine and a limi-
ted number of by-products upon specific organisms, virtually no informa-
tion is available on transport processes, persistence, bioaccumulations
and the fate of halogenated compounds from chlorination processes.
CHEMISTRY OF CHLORINE
Chlorine is presently manufactured by a variety of methods, including:
the electrolysis of brine,
2NaCl + 2H20 + electnc current > 2NaOH + C12 + H2
the salt process,
3NaCl + 4HN03 " 3NaN03 + Cl2 + NOC1 + 2H20
and the hydrochloric acid oxidation process,
4HC1 + 02 450"650°C > 2C12 + 2H20.
CHLORINE IN FRESHWATER SYSTEMS
Chlorine gas dissolves rapidly in water and hydrolyses,
C12 + H20 < • HOC1 + H + Cl .
This hydrolysis is nearly complete and only when the pH is below 3.0,
or the chlorine concentration over 1000 mg/1 is there any measurable quan-
tity of molecular chlorine present. The oxidizing capacity of chlorine is
retained in the hydrolysis product, hypochlorous acid. Hypochlorous acid
dissociates to form,
HOC! *=r H+ + CIO".
This reaction is pH dependent. For a neutral pH (7.0) at 20°C, the equi-
librium is approximately 75 percent HOC! and 25 percent CIO". For a pH of
8.0, the reverse is true with approximately 25 percent HOC1 and 75 percent
C1(T (Sawyer and McCarty, 1969).
The addition of hypochlorite salts to water forms hypochlorite ions
followed by hypochlorous acid,
Ca (C10)2 = Ca +2 Clo",
i
and H + CIO" « HOC1.
If ammonia or organic amines are present in the water, they will react with
hypochlorous acid to form
47
-------
chloramines,
NHs + HOC1 = NH2C1 + H20.
Like the ionization of hypochlorous acid to H + + CIO ~, the reaction
rate between ammonia and hypochlorous acid is pH dependent, occurring
most rapidly in solutions with a pH of 8.3. This reaction is also depen-
dent upon temperature and the ratio of ammonia to hypochlorous acid.
Monochloramines react with hypochlorous acid to form di- and tri-
chloramines,
and
NH2C1 + HOC1 = NHC12 + H20,
2NH2C1 + HOC1 = NCI 3 + H20.
Low pH favors a shift in eguilibrium toward the formation of di- and tri-
chloramines. Fair et al. (1948) determined that at pH 5.0, the ratio was
16 percent monochloramine and 84 percent dichloramine. For a pH of 8.0,
the ratio was 85 percent monochloramine and 15 percent dichloramine. Tri-
chloramine is found in significant quantities only at pH values of less
than 4 (McKee and Wolf, 1963).
Ingols et al. (1953) determined that hypochlorous acid and monochlor-
amine in freshwater will react with various, organic constituents. Some of
these reactions resulted in the formation of organic monochloramines al-
though none were persistent (Table 1). The formation of chlorinated organic
TABLE 1. SUMMARY OF REACTIONS OF CHLORINE WITH ORGANIC
COMPOUNDS IN FRESHWATER (MODIFIED FROM INGOLS
ET AL. 1953)
Organic Substrate
Hypochlorous Acid
Monochloramine
Alanine
Cysteine
Glycylglycine
Glycylglycylglyci ne
Tyrosine
Hemi n
Pyruvic Acid
RS03H
Oxidative
Hydrolysis and
Deaminization
Ketone
Violent Change
Organic Monochloramine
RSSR
Terminal Organic
Monochloramine
Organic Monochloramine
Irreversible Addition
or Oxidation
48
-------
compounds during chlorination of sewage effluents and power plant cooling
waters has,firecently been documented (Jolley, 1973; Jolley et al. 1975).
Isotopic Cl tracers and high-resolution anion-exchange chromatography
were used to separate over 50 chlorine containing constituents from chlor-
inated secondary effluents. Seventeen of these were tentatively identified
and quantified (Table 2).
TABLE 2. TENTATIVE IDENTIFICATIONS AND CONCENTRATIONS OF
CHLORINE CONTAINING CONSTITUENTS FROM CHLORINATED
SEWAGE EFFLUENTS (MODIFIED FROM JOLLEY, 1973)
Identification Cone, of Organic
Compound yg/1
5-Chlorouracil 4.3
5-Chlorouridine 1.7
8-Chlorocaffeine 1.7
6-Chloroguanine 0.9
8-Chloroxanthine 1.5
2-Chlorobenzoic Acid 0.26
5-Chlorosalicylic Acid 0.24
4-Chloromandelic Acid 1.1
2-Chlorophenol 1.7
4-Chlorophenylacetic Acid 0.38
4-Chlorobenzoic Acid 0.62
4-Chlorophenol 0.69
4-Chlororesorcinol 1.2
3-Chloro-4-Hydroxybenzoic Acid 1.3
4-Chloro-3-Methyl Phenol 1.5
CHLORINE IN MARINE SYSTEMS
Major sources of chlorine contamination in the marine environment are
related to postchlorination of secondary sewage effluents with outfalls
located on coastal and estuarine waters, and chlorination of seawater used
for cooling of thermal electric generating plants (White, 1972, 1973; Mar-
kowski, 1959).
The addition of chlorine to seawater results in a complex series of
chemical reactions, the most obvious one frees bromine,
C12 + 2Br" = 2C1" + Br2.
This reaction goes to completion and is the basis for the manufacture of
bromine from seawater (Lewis, 1966).
The industrial extraction of bromine from seawater requires that the
pH be reduced below 3.0, so that molecular chlorine can release molecular
49
-------
bromine. The hydrolysis products from adding chlorine to seawater, HOC!
and CIO , will also release bromine from the bromide ion in the form of
hypobromous acid and hypobromite ion,
CIO" + Br = BrO + Cl ,
and
BrO" + H" i=" HOBr.
Houghton (1946) has also suggested that chlorination of water containing
free ammonia and bromine may result in the formation of bromamines. Jo-
hanneson (1958) added chlorinated water to a sodium-ammonium salts solu-
tion buffered to pH 8.3. This resulted in the formation of monobromamine
and some monochloramine. The addition of sodium hypochlorite solution
produced mostly monochloramine. The hypochlorite in solution apparently
reacts with both the bromine and ammonia,
CIO" + Br" = BrO" + Cl",
and
CIO" + NH3 = NH2C1 + OH".
Injection of chlorine gas may result in localized acidity, favoring the
first reaction above, which is rapid at pH values of less than 8.0. The
second reaction is favored when chlorine is added as sodium hypochlorite
since there is no accompnying reduction in the normal pH of 8.0-8.3.
When ammonia is present in seawater, it will react with hypobromous
acid to form monobromamine. Monobromamine in turn will react with hypo-
bromite ion to form dibromamine,
NH3 + HOBr = NH2Br + H20,
and
NH2Br + BrO" = NHBr2 + OH".
In addition, monobromamine at near neutral pH will form monobrom-
ammonium which dissociates into ammonium ion and free bromine (Johanneson,
1960).
,+
NH9Br + H * NH-Br"1
and L 6
Block and Helz (1975) have prepared a reaction series model to illus-
trate the theoretical degradation processes occurring after the addition
of chlorine to natural, saline waters (Figure 1). Compounds in each suc-
cessive level can give rise to ones on a lower level. In general, com-
pounds occurring on lower levels will not contribute to the formation of
those in the levels above.
The reaction occurring between levels I and II is a result of chlorine
decay from a diatomic gas to hypochlorous acid, hypochlorite ions and sodium
hypochlorite. As pointed out by Moore (1951) and Lewis (1966), this reaction
occurs rapidly and goes to completion within seconds after the addition of
chlorine. The inclusion of sodium hypochlorite within level II is based on
the results of work by Sugam and Helz (1975, unpublished manuscript).
50
-------
II
III
IV
Cl.
HOC1, OC1", NaOCl etc.
NH2C1, NHC12, NH2Br,
NHBr2, BrO", HBrO
Organic chlorine and
bromine compounds
Cl", Br"
Figure 1. Degradation processes for chlorine in saline waters
(modified from Block and Helz, in preparation)
The chemical composition and abundance of products formed from level
II to level III is a function of physical and chemical parameters of the
water including but not limited to temperature, pH, ammonia, and bromine,
available as reaction components. In seawater it is possible that the
predominant species would be bromamines, especially if NH. ions are less
abundant than Br" ions. 4
Level IV includes halogenated organic constituents which may be formed
by level II or level III species, including chloramines, hypobromite and
bromamines. The stable end products in level V occur through a diverse
group of mechanisms taking place in steps I-IV.
Charge balance results in one atom of Cl passing from level I to le-
V to each atom passing from level I to level II. Reduction of hypo-
chlorite by Br~ or Fe^ and Mrr may release Cl ~ from level II to level V.
Movement of Cl" from level III to level V can also occur in a number of
ways, the most obvious, suggested by Laubusch (1971), involves the des-
+ „,.„.•.,• -c _ui ,•„„ ...u__ *i.- ""- ratio is large.
vel
truction of chloramines when the
Some of the chlorinated organics identified by Jolley
sistent and the decay from level IV to level V is probably
relative to decay from levels I through III to level V.
(1973)
a slow
are per-
process,
51
-------
TOXICITY OF CHLORINE IN ESTUARINE ENVIRONMENTS
The relative toxicity of chlorine in water is related to the amount
and proportions of free and residual chlorine. Several investigators have
found that free chlorine is generally more toxic to freshwater organisms
than chloramines (Douderoff and Katz, 1950; Merkens, 1958), even though
the toxicity of the various forms of chlorine were of the same order of
magnitude. Rosenberger (1971) and Basch and Truchan (1973), found that
dichloramine was more toxic than monochloramine in freshwater. A com-
prehensive review paper by Brungs (1973) summarizes the toxic effects of
residual chlorine on freshwater aquatic organisms.
In seawater, Holland et al. (1960) determined that dichloramine is
apparently more toxic than monochloramine and that the chloramines were
more toxic than free chlorine. These findings may reflect the complex
chlorine-bromine reaction kinetics suggested by Johanneson (1958, 1960)
and Lewis (1966).
CHLORINE TOXICITY TO MARINE PHYTOPLANKTON
The effects of chlorination and thermal pollution on phytoplankton
productivity have been investigated in some detail (Table 3). Carpenter
et al. (1972) observed an 83 percent decrease in the productivity of
phytoplankton passed through the cooling system of a nuclear generating
plant on Long Island Sound.
TABLE 3. SUMMARY OF TOXIC EFFECTS OF CHLORINATED
WASTES AND WATER ON MARINE PHYTOPLANKTON.
Species
Phytoplankton
Chlamydomona
Skeletonema
costatum
Phytoplankton
Phytoplankton
Toxicant
used
C12 injection
hypochlorite
solution
hypochlorite
solution
Cl- injection
Measured
residual
chlorine
mg/1
0.05-0.40
0.69-12.9
0.18-2.4
0.32
0.01
0.075-0.25
—
Duration
of
Test
12 hrs +
4 hrs incu-
bation
5 min
5 min
2 min
45 min
24 hrs
15 min
Effect(s)
50-98% loss
Of productivity
Reduced growth
rate
None up to 0.29
mg/1 ; greater
amounts inhibit-
ed growth
55% decrease in
ATP
77% decrease in
ATP
50% decrease in
growth
91% reduction in
photosynthesis
Reference
Carpenter et
al. (1972)
Hirayama and
Hirano (1970)
Gentile et al .
unpublished da
(1972, 1973)
Hamilton et al
(1970)
52
-------
Intake water was chlorinated at a rate of 1.2 mg/1 with a residual
of 0.4 mg/1 measured at the discharge. Addition of 0.1 mg/1 chlorine at
the intake with nondetectable residuals at the outfall decreased produc-
tivity by 79 percent. Essentially no decreases in productivity were ob-
served when phytoplankton passed through the cooling system without
addition of chlorine. Hirayama and Hirano (1970) measured the effect of
chlorination on the photosynthetic activity of Skeletonema costatum and
found that cells were killed when subjected to T.5 to 2.3 mg/1 chlorine
for 5 and 10 minutes.
Gentile (1972, 1973 unpublished data, National Marine Water Quality
Laboratory, West Kingston, RI) observed a 55 percent decrease in the ATP
content of marine phytoplankton exposed to 0.32 mg/1 residual chlorine
for two minutes and a 77 percent decrease after 45 minutes of exposure
to chlorine concentrations as low as 0.01 mg/1. A 50 percent depression
in the growth rates of 10 species of marine phytoplankton exposed to
chlorine concentrations ranging from 0.075 to 0.25 mg/1 for 24 hours was
also measured.
Morgan and Stress (1969) used photosynthetic rates to evaluate the
response of estuarine phytoplankton passed through the cooling system of
a steam electric power station on the Patuxent River, Maryland. The photo-
synthetic rate increased with an 8°C rise in temperature when ambient wa-
ter temperatures were 16°C or less. Inhibition occurred when ambient tem-
peratures were above 20°C. In a related study, conducted at the same site,
Hamilton et al. (1970) measured a 91 percent decrease in primary produc-
tivity during intermittent chlorination.
CHLORINE TOXICITY TO INVERTEBRATES
Muchmore and Epel (1973) investigated the effects of chlorination of
wastewater on fertilization in marine invertebrates (Table 4). Unchlor-
inated sewage (from the Pacific Grove, California STP) was a weak inhibi-
tor of fertilization in the sea urchin, Strpnqylocentrotus purpuratus.
Exposure of gametes of the sea urchin to^a 10 percent unchlonnated
sewage-seawater mixture typically reduced fertilization success by 20 per-
cent. A 0.5 percent dilution of moderately chlorinated sewage (11 mg/1
TRC undiluted) significantly reduced fertilization. It was also deter-
mined that chlorination had more effect on sperm cells than on eggs. Eggs
incubated for 5 minutes in a 0.77 mg/1 hypochlorite solution and subse-
quently washed to remove the hypochlorite showed no reduction in fertility.
Incubation of sperm at a 0.07 mg/1 hypochlorite concentration resulted in
a loss of fertilization ability. This was attributed to a loss of sperm
motility which was not restored after washing to remove the hypochlorite.
Gametes of the echiuroid, Urechis caupcu and sperm of the annelid worm,
Phragmatopoma caliform'ca, were not as sensitive to chlorine toxicity.
A number of power plant related studies have been conducted to deter-
mine the effect of chlorination of seawater on fouling organisms. Waugh
(1964) observed no significant difference in the mortality of oyster lar-
vae, Ostrea edulis, exposed to 5 mg/1 chlorine for 3 minutes at ambient
temperature, computed to control mortality. Exposure of larvae to thermal
stress (10°C above ambient) and 10 mg/1 chlorine for 6 to 48 minutes also
had no significant effect on survival 64 hours after treatment. Barnacle
53
-------
nauplii, Eliminius modestus, showed more acute sensitivity to chlorine.
Residual chlorine concentrations in excess of 0.5 mg/1 caused heavy mor-
tality and reduced growth for survivors.
TABLE 4. SUMMARY OF TOXIC EFFECTS OF CHLORINATED
WASTES AND WATER ON MARINE INVERTEBRATES
Species Toxicant
used
Stronaylo- chlorinated
centrus sewage effluents
purpuratus
(gametes)
Urechis
caupo
(gametes)
Phrogmatopoma
californica
(sperm)
Elminius residual
modestus chlorine
Melita Cl, injection
nitida i
Gammurus sp.
Bimaria
franciscana
Balanus sp.
Acartia tonsi
Anemones residual
chlorine
Mussels
Barnacles
My til us Cl,
edulis infection
Measured
residual
chlorine
mg/1
0.02
0.11
0.03
0.13
0.2
1.0
0.2
1.0
2.0
5.0
2.5
2.5
4.5
2.5
2.5
10.0
2.5
1.0
10.0
2.5
1.0
10.0
2.5
1.0
0.02
0.05
Duration
of
Test
5 min
5 min
5 min
5 min
5 min
5 min
5 min
5 min
10 min
3 min
5 min
3 hrs
48 hrs
96 hrs
3 hrs
4 days
5 min
5 min
1. 2, 4, 8
hrs/day for
10 days
8 days
15 days
1, 2, 4, 8
hrs/day for
10 days
5 days
15 days
1. 2, 4 hrs/
day for 10
days
4 days
7 days
A few hrs
Effect(s)
None
100% inhibition
of fertilization
None
99% inhibition
of fertilization
22% inhibition
of fertilization
100% inhibition
of fertilization
22% loss of
motility
86% loss of
motility
Death and inhib-
ited growth
None
None
27% mortality
72% mortality
97% mortality
25% mortality
96 hrs after
exposure
None
80% mortality
90% mortality
None
100% mortality
100% mortality
None
100" mortality
100% mortality
95-100% mortality
100% mortality
100X mortality
Detachment and
migration
Reference
Muchmore and
Epel (1973)
Waugh (1964)
McLean (1972,
1973)
Turner et al .
(1948)
James (1967)
54
-------
McClean (1973) simulated the conditions encountered by marine organ-
isms passing through a power plant on the Patuxent River, Maryland. In-
take chTorination to 2.5 mg/1 residual, entrainment for approximately 3
minutes and sustained exposure to elevated temperatures for up to 3 hours
were used as experimental parameters. While barnacle larvae, Balanus sp.
and copepods, Acartia tonsi, were not affected by a 3 hour temperature
stress of 5.5 and 11°C above ambient; exposure to 2.5 mg/1 residual chlorine
for 5 minutes at ambient temperatures caused respective mortality rates of
80 and 90 percent. The amphipod, Melita nitida, and the grass shrimp,
Palaemonetes pugio, showed a delayed death response after exposure to 2.5
mg/1 TRC for 5 minutes. Near 100 percent mortality was observed for both
species 96 hours after exposure to the chlorine residual. McLean (1972)
showed that established colonies of the euryhaline colonial hydroid, Bimeria
franciscana. were not greatly affected by 1 and 3 hours of exposure to
4.5 mg/1 TRC.
Turner et al. (1948) determined that continuous treatment of seawater
conduits with 0.25 mg/1 chlorine prevented fouling during a 90 day interval
when the flow velocity was 52 cm/second or less. Intermittent treatment
with 10 mg/1 residual chlorine for 8 hours a day was ineffective in pre-
venting fouling by anemones, mussels and barnacles.
James (1967), working in Great Britain, observed that residual chlor-
ine concentrations of 0.02 and 0.05 mg/1 caused detachment and movement
of mussels in the direction of water flow through an aquarium with eventual
elimination of the mussels. He concluded that the most effective way to
prevent fouling by mussels was not to kill, but to discourage settling in
cooling water systems by continuous low level chlorination.
Markowski (1960) compared the occurrence of marine organisms on con-
crete slabs placed in the intake and outfall canals of an electric gener-
ating plant. Chlorine was injected into the condensers of this plant for
two hours a day at a concentration between 1 and 2.5 mg/1. No vegetation
was observed growing in the intake canal where dense animal populations
occurred (predominantly invertebrates, Coelenterata and Polyzoa). The
outfall canal contained a prolific growth of algae, Enteromorpha sj^. but
fewer invertebrates. Balanus improvisis, which was col 1 ected witFsome
regularity from the intake canal was never observed in the outfall canal.
The mollusk, Eubranchus sp. was more abundant on the intake slabs than in
the outfall.
CHLORINE TOXICITY TO ESTUARINE FISH
Tsai (1968, 1970, 1975) has observed decreases in the abundance and
occurrence of brackish water fish species in certain areas of the Upper
and Little Patuxent Rivers receiving chlorinated sewage effluent. Tsai
suggests that chlorinated sewage effluent may also block the upstream mi-
gration of such semi-anadromous species as the white catfish and white
perch. He attributed the "blocking effect" to chlorination products
rather than reduced dissolved oxygen or pH resulting from organic decom-
position of the effluent (Table 5).
Tsai (1973) measured the diversity index of fish upstream and downstream
of 98 sewage treatment plants in Virginia, Maryland and Pennsylvania. Sewage
treatment plants were categorized as Type I engineering facilities (sludge
55
-------
TABLE 5. SUMMARY OF TOXIC EFFECTS OF CHLORINATED WASTES
AND WATER ON MARINE AND FRESHWATER FISHES
Species
Cyprlnus
carpi'o
eggs
(Freshwater)
Freshwater
and brackish
fishes
L. xanthurus
Morone sp.
Pomatomus
saltatrix
C. regal is
Brevoortia
tyrannus
L. xanthurus
0_. nerka
0. gorbuscha
^Freshwater)
£. gorbuscha
0. tshawytscha
Morone
americana
Menidia
menidia
F. hetero-
cli tus
Trinectes
macula tus
Pleuronectes
platessa
eggs
larvae
Toxicant
used
4-Chlororesor-
cinol
5-Chlorouracil
(0.001 mg/1)
chlorinated
sewage effluents
chlorinated
sewage effluents
sodium hypo-
chlori te
chlorinated
sewage effluents
residual
chlorine
residual
chlorine
free chlorine?
Measured
residual
chlorine
mg/1
0.6-2.0
0.07-0.28
0.09
0.14
0.28
0.02-0.026
0.16
0.5
0.5
0.08
0.08
0.03
0.03
0.04-0.08
0.70
0.12
0.032
0.026
Duration
of
Test
3-7 days
Long-term
May-June, 1973
96 hrs
24 hrs
6 hrs
24 hrs
72 hrs
80 min + 10°C
10 min + 10°C
thermal shock
10 m1n
10 min
10 min
10 min
8 days
72 hrs
96 hrs
48 hrs
96 hrs
Effect(s)
Reduced hatch
Decreased popn.
size and diver-
sity
Probable kill
5-10 million
fish
50% mortality
50% mortality
50% mortality
100% mortality
100% mortality
50% mortality
50% mortality
Avoidance.
Avoidance
Avoidance
Avoidance
None
50% mortality
50% mortality
50% mortality
50% mortality
Reference
Gehrs et al .
(1974)
Tsai (1968, 1970,
1973)
Virginia State
Water Control
Board (1974)
Virginia Inst.
Marine Science
for VSWCB (1974)
Servizi and
Martens (1974)
Stober and
Hanson (1974)
Meldrim et al.
(1974)
Alderson (1972)
activation, aeration, sedimentation and filtration) with effluent chlorina-
tion; Type II engineering facilities with chlorination and an effluent hold-
ing lagoon and Type III engineering facilities with a lagoon and effluent
chlorination at the lagoon outlet. Reductions in the number of fish, num-
ber of species and the species diversity index were significant downstream
of Type I and III plants. These reductions were attributed to total residual
chlorine levels and turbidity. Diversity indices showed no significant
changes in downstream areas associated with Type II plants.
Massive fish kills occurred on the James River, Virginia during May-
June, 1973 (Virginia State Water Control Board, 1974). Species affected
by the kill included spot, Leiostomus xanthurus; white perch, Morone amer-
i cana; bluefish, Pomatomus saltatrix; grey seatrout, Cynoscion regal is and
menhaden, Brevoortia tyran'nus. A majority of the fish kill in the James
River occurred adjacent to sewage treatment plants. Total residual chlor-
ine (TRC) levels as high as 0.7 mg/1 were observed in the James. Effluents
from both plants showed more than 3.0 mg/1 TRC.
56
-------
Distress symptoms of fish dying included, spiral swimming patterns,
broken vertebral columns, listless floating, inverted swimming, distension
of the air bladder in some, loose body scales, mucous on the skin and hem-
orrhaging along the fins and body surface.
Live box tests conducted adjacent to the James River sewage treatment
plant (STP) demonstrated a correlation between rates of effluent chlorina-
tion and mortality of juvenile spot and croaker. With an average daily
chlorine feed of 1200 pounds (total flow of water was approximately 10 mgd
during tests) and a measured residual chlorine level of 3.0 mg/1,caged
fish suffered 100 percent mortality within 20 hours. After a cutback to
a chlorine feed rate of approximately 400 pounds per day, only 20 percent
mortality was observed among caged fish after 20 hours.
On-site aquaria tests confirmed the results of the cage tests. Water
from an area adjacent to the outfall of the James River (STP) was pumped
through aquaria containing juvenile spot. Mortalities ranged from 91 to
100 percent after 40-85 minutes of exposure prior to the cutback in chlor-
ination. After chlorination rates were reduced, mortalities were 0-26
percent after 120 minutes of exposure.
Continuous flow laboratory bioassays were also conducted. The 96
hour LC§Q for juvenile spot was estimated at 0.09 mg/1 TRC. The esti-
mated 24 hour LC5Q was 0.14 mg/1 and the 6 hour LC50, 0.28 mg/1 TRC.
Separate field studies on the spot, Leiostomus xanthurus, found up
to 40 percent of juveniles from the 1973 year class exhibited deformities
in the vertebral column. These abnormal forms are identifiable as a dis-
tinct year class in 1975 population samples from the Chesapeake Bay, (Chao
Labbish, personal communication).
A study of the effect of chlorinated sewage effluents on sockeye
salmon, Onchorhynchus nerka, and pink salmon, 0_. gorbuscha, has been
conducted by Servi-zi and Martens (1974). They used three study sites to
conduct cage bioassays. The first, Site I, was adjacent to a primary
treatment plant with effluents chlorinated following settling and.dis-
charged through a 600' pipe line directly into the receiving stream.
Site II was on a stream receiving wastes from an activated sludge plant
in which chlorinated effluents were discharged into a large effluent
holding lagoon and retained for 30 to 60 days. Site III was located on
a stream receiving effluents which were chlorinated as they left a non-
aerated lagoon.
Measured chlorine residuals in the receiving stream at Site I ranged
from 0.02-0.26 mg/1. These concentrations resulted in 100 percent mor-
tality of caged sockeye fingerlings placed 30, 60 and 250 feet below the
effluent discharge point. Additional tests indicated that the primary
effluent without chlorination was also toxic. However, fish exposed to
the unchlorinated effluent lived ten times longer than ones exposed when
effluents were being chlorinated. Toxicity of the unchlorinated effluents
was attributed to MBAS and ammonia.
Tests at Site II indicated that chlorinated effluents retained for
30 to 60 days were not toxic to sockeye fingerlings and alevins and pink
salmon alevins after 26 days of exposure.
57
-------
In tests at Site II, with fingerling sockeye salmon, chlorinated
sewage effluents (measured TRC 0.85 mg/1) resulted in 50 percent mor-
tality after 48 minutes. Fifty percent mortality occurred after 13
hours of exposure to the unchlorinated effluents. Sublethal exposures
of fingerling sockeye salmon to the effluents from Site III (1-3 hours
of exposure to 0.22 mg/1 TRC) resulted in gill damage, including hyper-
pi asi a, swollen epithelial cells, and separation of epithelium from
pillar cells.
The toxicity of chlorine and heat to pink, Oncorhynchus gorbuscha,
and Chinook salmon, 0_. tshawytscha, has been determined by Stober and
Hanson (1974). Juveniles of each species were tested in seawater at five
residual chlorine concentrations, ranging from 0.05-1.0 mg/1, and four
temperatures from t 0-10°C. Salmon were exposed to each matrix for 7.5-
60 minutes. A decrease in the tolerance of both species to residual
chlorine was observed with increased temperature and exposure time. The
most toxic effect was observed at a t of 9.9-10°C where the LT^Q (lethal
time for 50 percent mortality) ranged from approximately 10 minutes at
0.5 mg/1 TRC for chinooks to 80 minutes for pinks.
Meldrim et al. (1974) in flowing water bioassays studied the effect
of chemical pollutants on estuarine organisms. They found that white
perch, Morone americana, consistently avoided TRC levels as low as 0.08
mg/1 at temperatures from 7-17°C. Silversides, Menidia menidia, also
avoided 0.08 mg/1 TRC at temperatures from 8-28°C but showed a preference
for 0.08 mg/1 TRC when fish acclimated to 7°C were exposed at 12°C. Mum-
mi chogs, FundijUjs he^terocjjjtus, and hog chokers, Trinectes maculatus,
avoided TRC levels as low as 0.03 mg/1.
Alderson (1972) found that the 48 and 96 hour Tlm of free chlorine
for plaice larvae, Pleuronectes platessa, was 0.032 and 0.026 mg/1 res-
pectively. Eggs were not affected when exposed to 0.075 and 0.04 mg/1
free chlorine for 8 days, indicating that the egg membrane gives consider-
able protection over long periods. The 72 and 192 hour Tlm for eggs was
0.7 and 0.12 mg/1 TRC respectively.
Gehrs et al. (1974) tested the sensitivity of carp eggs, Cyprinus
carpio, to two of the compounds identified by Jolley, 4-Chlororesorcinol
and 5-Chlorouracil. Significant reductions in the hatchability of non-
water hardened carp eggs were observed in concentrations of each compound
as low as 0.001 mg/1.
In California, Young (1964) observed tumor-like sores around the
mouth of white croakers, Genyonemus 1ineatus, collected near the Hyperion
sewage outfall in Santa Monica Bay. While there was no direct evidence
to link the occurrence of lesions with chlorinated sewage effluents, a
general decline in fitness of croakers and other species found in close
proximity to the outfall area was observed.
58
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REFERENCES
Alderson, R. 1972. Effects of low concentrations of free chlorine on
eggs and larvae of plaice, Pleuronectes platessa L. In: Marine
Pollution and Sea Life. Fishing News, Ltd., (London): 312-315.
Basch, R.E. and J.6. Truchan. 1973. Calculated residual chlorine con-
centrations safe for fish. Interim Report, Michigan Water Resources
Commission, Bureau of Water Management, Lansing, Michigan.
Block, R.M. and G.R. Helz. 1975. Biological and chemical implications
of chlorine in estuarine and marine systems. Chesapeake Sci., In
press.
Brungs, W.A. 1973. Effects of residual chlorine on aquatic life. Jour.
Wat. Pollut. Contr. Fed. 45(10): 2180-2192.
Carpenter, E.J., B.B. Peck and S.J. Anderson. 1972. Cooling water
chlorination and productivity of entrained phytoplankton. Marine
Biology 16: 37-40.
Doudoroff, P. and M. Katz. 1950. Critical review of literature on the
toxicity of industrial wastes and their components to fish. Sew.
and Ind. Wastes 22(11): 1432-1458.
Fair, G.M., J.C. Morris, S.L. Chang, I. Weil, and R.P. Burden. 1948.
Chlorine as a water disinfectant. Jour. Amer. Water Works Assoc.
40: 1051-1061.
Gehrs, C.W., L.D. Eyman, R.L. Jolley and J.E. Thompson. 1974. Effects
of stable chlorine-containing organics on aquatic environments.
Nature 249: 675-676.
Gentile, J.H., J. Cardin, M. Johnson and S. Sosnowski. 1972. The effects
of chlorine on the growth and survival of selected species of estuarine
phytoplankton and zooplankton. Unpublished manuscript, National Marine
Water Quality Laboratory, West Kingston, RI.
Gentile, J.H., S. Cheer, and N. Lackie. 1973. The use of ATP in the
evaluation of entrainment. Unpublished data. National Marine Water
Quality Laboratory, West Kingston, RI.
Hamilton, D.H., D.A. Flemer, C.V. Keefe, and J.A. Mihursky. 1970.
Power plants: Effects of chlorination on estuarine primary productivity.
Science 169: 197-198.
Hirayama, K. and R. Hirano. 1970. Influence of high temperature and
residual chlorine on marine phytoplankton. Marine Biology 7: 205-213.
59
-------
Holland, G.A., J.E. Lasater, E.D. Neumann, and W.E. Eldridge. 1964.
Toxic effects of organic and inorganic pollutants on young salmon
and trout. State of Washington, Dept. of Fish Res. Bull. No. 5,
264 p.
Houghton, G.U. 1946. The bromine content of underground waters. Part
II: Observations on the chlorination of water containing free
"ammonia and naturally occurring bromide. Jour. Soc. of Chemical
Industry 65: 324-328.
Ingols, R.S., H.A. Wyckoff, T.W. Kethley, H.W. Hodgen, E.L. Fincher,
J.C. Hildebrand, and J.E. Mandel. 1953. Bactericidal studies of
chlorine. Industrial and Engineering Chemistry 45: 995-1000.
James, W.G. 1967. Mussel fouling and use of exomotive chlorination.
Chem. and Ind. 24: 994-996.
Johanneson, J.K. 1958. The determination of monobromamine and mono-
chloramine in water. Analyst 83: 155-159.
Johanneson, J.K. 1960. Bromination of Swimming Pools. Am. Jour.
Publ. Health 50: 1731.
Jolley, R.L. 1973. Chlorination effects on organic constituents in
effluents from domestic sanitary sewage treatment plants. Ph.D.
Dissertation, Univ. of Tennessee. 339 p.
Jolley, R.L., C.W. Gehrs, and W.W. Pitt. 1975. Chlorination of cooling
water: A source of environmentally significant chlorine-containing
organic compounds. Proceeding of the 4th National Symposium on
Radioecology, Con/all is, Oregon.
Laubusch, E.J. 1971. Chlorination and other disinfection processes.
In: Water Quality and Treatment. Am. Water Works Assoc. 654 p.
Lewis, B.G. 1966. Chlorination and muscle control. I_. The Chemistry
of chlorinated seawater. A review of the literature. Central
Electric Res. Lab., Lab. Note No. RD/L/N/106/66.
Markowski, S. 1959. The cooling water of power stations: A new factor
in the environment of marine and freshwater invertebrates. Jour.
Animal Ecol. 28: 243-258.
Markowski, S. 1960. Observations on the response of some benthonic
organisms to power station cooling water. Jour. Animal Ecol.
29: 349-357.
McKee, J.E., and H.W. Wolf. 1963. Water Quality Criteria. 2nd Ed.,
Publ. 3A, Calif. State Water Quality Control Board, Sacramento. 548 p,
McLean, R.I. 1972. Chlorine tolerance of the colonial hydroid, Bimeria
franciscana. Chesapeake Sci. 13: 229-230.
60
-------
McLean, R.I. 1973. Chlorine and temperature stress in estuarine in-
vertebrates. Jour. Wat. Pollut. Contr. Fed. 45: 837-841.
Meldrim, J.W., J.J. Gift, and B.R. Petrosky. 1974. The effect of
temperature and chemical pollutants on the behavior of several
estuarine organisms. Icthyological Assoc. Inc. Bull. No. 11:1-129.
Merkens, J.C. 1958. Studies on the toxicity of chlorine and chloramines
to the rainbow trout. Water and Waste Treatment Jour. 7:150-151.
Moore, E.W. 1951. Fundamentals of chlorination of sewage and wastes.
Water and Sewage Works 98: 130-136.
Morgan, R.P., and R.G. Stress. 1969. Destruction of phytoplankton in
the cooling water supply of a steam electric station. Chesapeake
Sci. 10: 165-171.
Muchmore, D., and D. Epel. 1973. The effects of chlorination of waste-
water on fertilization in some marine invertebrates. Marine Biology
19: 93-95.
Rosenberger, D.R. 1971. The calculation of acute toxicity of free chlorine
and chloramine to coho salmon by multiple regression analysis. Thesis,
Michigan State Univ., East Lansing, Michigan.
Sawyers, C.N., and P.L. McCarty. 1969. Chemistry for Sanitary Engineers.
2nd Ed., McGraw-Hill, New York.
Servizi, J.A., and D.W. Martens. 1974. Preliminary survey of toxicity
of chlorinated sewage to sockeye and pink salmon. Pacific Salmon
Fisheries Comm. Progress Report No. 30: 1-42.
Stober, Q.J., and C.H. Hanson. 1974. Toxicity of chlorine and heat to
pink, Onchrhynchus gorbuscha and Chinook salmon, 0_. tshawytscha.
Trans. Amer. Fish. Soc. 103 (3): 569-576.
Sugam, R., and G.R. Helz. 1975. Apparent ionization constant of hypo-
chlorous acid in seawater. Unpublished manuscript, Univ. of Maryland,
College Park, Maryland.
Tsai, C. 1968. Effects of chlorinated sewage effluents on fishes in
upper Patuxent River, Maryland. Chesapeake Sci. 9(2): 83-93.
Tsai, C. 1970. Changes in fish populations and migrations in relation
to increased sewage pollution in Little Patuxent River, Maryland.
Chesapeake Sci. 11(1): 34-41.
Tsai, C. 1973. Water quality and fish life below sewage outfalls.
Trans. Amer. Fish. Soc. 102(2): 281-292.
Tsai, C. 1975. Effects of sewage treatment plant effluents on fish:
A review of the literature. Chesapeake Research Consortium.
Publ. No. 36: 1-229.
61
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Turner, H.J., D.M. Reynolds, and A.C. Redfield. 1948. Chlorine and
sodium pentachlorophenate as fouling preventives in seawater
conduits. Indus, and Engin. Chem. 40: 450-453.
Virginia State Water Control Board. 1974. James River Fish Kill 73-025.
Bureau of Surveillance and Field Studies, Division of Ecological
Studies. 61 p.
Waugh, G.D. 1964. Observations on the effects of chlorine on the larvae
of oysters, Ostrea edulij^ L., and barnacles, Eliminius modestus,
Darwin. Ann. Appl. Biol. 54: 423-440.
White, G.C. 1972. Handbook of Chlorination. Van Nostrand Rheinhold
Co. New York 744 p.
White, G.C. 1973. Disinfection practices in the San Francisco Bay Area.
Jour. Wat. Pollut. Contr. Fed. 46: 89-101.
Young, P.M. 1964. Some effects of sewage effluents on marine life.
Calif. Fish and Game 50(1): 33-41.
62
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Techniques to Assess the Effects of
Toxic Organics on Marine Organisms
David J. Hansen*
ABSTRACT
Acute static or flow-through bioassays generally
have been used to set marine water quality stan-
dards, but few new bioassay techniques are avail-
able to determine long-term effects of one or more
toxicants on survival, growth and reproduction of
individual species of mollusks, arthropods or fish
and on communities of estuarine organisms. Not
only has the duration of bioassays increased from
96 hours or less to periods of from one month to
two years, but the complexity has increased as well.
Effects of toxicants on the entire life-cycle of an
oviparous estuarine fish, Cyprinodon variegatus, can
now be studied; one bioassay with endrin has been
completed. This fish typically develops from an
embryo to maturity in 10 weeks, with about 70% sur-
vival overall. Females produce an average of eight
eggs per day and fertilization success exceeds^0%.
Effects of a polychlorinated biphenyl, Aroclor^l254,
and a pesticide, toxaphene, on developing communities
of estuarine animals have been investigated. These
studies provide data for prediction of pollution-
induced shifts in composition of estuarine animal
communities.
INTRODUCTION
Bioassays are probably the most useful technique available to the
biologist for predicting the potential hazard of a chemical. Bioassays
vary considerably in complexity and utility and each procedure has its
own particular advantages and disadvantages. They range from relatively
simple acute static and flow-through bioassays, to complex chronic entire
life cycle and community bioassays. Flow-through acute bioassays usually
provide a more sensitive measure of stress than do static bioassays whereas
entire life cycle and community bioassays provide a better estimate of
"safe" concentrations from which water quality criteria can be derived.
*Environmental Research Laboratory, U.S. Environmental Protection Agency,
Gulf Breeze, FL 32561
63
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In spite of the importance of developing sound marine water quality
criteria to protect aquatic life, the quantity and quality of sound re-
search aimed at evaluating effects is limited. Results of a recent sur-
vey conducted by the Water Quality Committee of the American Fisheries
Society indicated that the funds and manpower spend on research to develop
water quality criteria were comparatively small and that bioassays were
of short duration and predominantly used freshwater fishes. The Committee
on Water Quality Criteria of the National Research Council summarized
marine bioassay data for a total of 70 organic chemicals in Table 6, pages
484 to 508 of the Blue Book, 1972 (NAS-NAE Committee on Water Quality
Criteria, 1972). A summary of this appendix (Table 1) helps to quantify
the findings of the American Fisheries Society Committee. All of the 317
experiments with phytoplanters were static tests. Of the 332 experiments
with estuarine animals only 12 percent were flowing water bioassays, only
16 percent included statistical treatment and few received chemical analyses
to determine the actual concentration of the chemical in the test water.
Significantly, less than two percent of all of these bioassays upon which
water quality criteria may be recommended were dynamic tests lasting longer
than 96 hours and no tests were on a complete life cycle of an animal or
on communities of organisms.
TABLE 1. BIOASSAY METHODS USED TO OBTAIN TOXICITY DATA
ON THE EFFECTS OF ORGANIC CHEMICALS ON MARINE
ORGANISMS AS REPORTED IN APPENDIX III, TABLE 6,
P. 484-590 OF "WATER QUALITY CRITERIA," 1972 -
THE BLUE BOOK
Organism
Totals
Kinds and Numbers of Bioassays
Static
<96 hrs.
327
Dynami c
<96 hrs.
<96 hrs.
<96 hrs.
Plants
An i ma 1 s
37
290
280
53
0
29
0
13
333
29
13
The purpose of this paper is to describe some recent improvements in
bioassay procedures used at the Gulf Breeze Environmental Research Labor-
atory (GBERL) to test the effect of toxicants on estuarine animals. These
are: (1) improved techniques for conducting constant-temperature and-
salinity acute bioassays in which the concentration of the toxicant is
measured and the data are treated statistically; (2) in-house and extra-
mural bioassays on sensitive larval stages of crabs and shrimp; (3) de-
velopment of methods to bioassay a portion, or the entire life cycle, of
grass shrimp (Palaemonetes pugio) and the sheepshead minnow (Cyprinodon
variegatus) and" (4)development of methods to assess the effects of
toxicants on entire communities of benthic macroinvertebrates.
64
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CONCLUSIONS
1. Acute 96-hour flow-through bioassays should be conducted using uni-
form methods on representative species from several phyla of estuarine
organisms. Use of uniform methods with appropriate statistical and
chemical analyses makes comparisons between tests reliable. Recent
acute bioassays indicate that previous tests have underestimated acute
toxicities of many organic chemicals.
2. In addition to acute bioassays, information is needed on the effects
of chemicals on sensitive life-stages and entire life-cycles of es-
tuarine organisms, as well as on communities of organisms, in order
to set sound water quality criteria. Some methods necessary to con-
duct these experiments are available and additional procedures are
presently being developed.
BIOASSAY TECHNIQUES
ACUTE BIOASSAYS
Acute toxicity experiments are usually conducted to determine the
quantity of chemical that will adversely affect a certain percentage of
the test organisms in a short period of time. This information is.used
to make comparisons of relative toxicity and relative sensitivity.' Com-
parisons become most reliable if bioassay methods are uniform and the
tests are conducted to obtain statistically valid data supported by chem-
ical analyses of the test water. Data from this type of bioassay, al-
though more difficult and costly to obtain than data from simpler screening
tests, are required by EPA because of the Agency's regulatory responsibilities
Acute bioassay methods used at 6BERL have changed since joining EPA.
When our laboratory was part of the Bureau of Commercial Fisheries, Jack I.
Lowe was in charge of the acute bioassays. From 1963 to 1972 he conducted
flow-through bioassays that usually lasted 48 hours on over 200 chemicals
on oysters, penaeid shrimp, fishes and occasionally crabs. His data were
used to help in pesticide registration and to develop label restrictions.
Acute flow-through bioassays are now being repeated on some of these chem-
icals to provide 96-hour LC50 data backed by statistical and chemical
analyses. The results of recent experiments continue to show that panaeid
shrimp are usually more sensitive to the chemicals tested than oysters,
grass shrimp or estuarine fishes (Table 2). The acute toxicity of these
chemicals, except methoxychlor, in our tests exceeded that of acute bio-
assays published in the Blue Book (NAS-NAE Committee on Water Quality
Criteria, 1972).
Recent acute bioassays have been conducted using water of constant
temperature and salinity to improve comparisons of the results of these
tests. Bioassays of DDT, heptachlor (99%), heptachlor epoxide, lindane
and methoxychlor in Table 2 were all conducted at 2.5°C and 20 °/oo salinity.
The salinity was controlled by an inexpensive device in which appropriate
amounts of fresh and saltwater were added through solenoid valves that were
controlled electrically by a photocell that sensed changes in water density
detected by a floating hydrometer (Bahner and Nimmo, 1975a). This device
has been used successfully for periods of up to 9 months to maintain con-
stant (±1 °/oo) salinity in bioassays.
65
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TABLE 2. NINETY-SIX HOUR LCSO'S AND 95% CONFIDENCE INTERVALS
FOR THE SPECIES OF ESTUARINE ORGANISM MOST SENSITIVE
TO SELECTED ORGANIC CHEMICALS IN FLOW-THROUGH BIO-
ASSAYS. USUALLY THE AMERICAN OYSTER, TWO FISHES AND
TWO ARTHROPODS WERE TESTED. CONCENTRATIONS IN WATER
WERE MEASURED BY ELECTRON-CAPTURE GAS CHROMATOGRAPHY
CHEMICAL
Chlordane
DDT*
Dieldrin
Endrin
HCB
Heptachlor
(74%)
Heptachlor
(99%)*
Heptachlor
Epoxide*
Lindane*
Methoxychlor*
Toxaphene
SENSITIVE
SPECIES
Pink Shrimp
Brown Shrimp
Pink Shrimp
Pink Shrimp
Pink Shrimp
Pink Shrimp
Pink Shrimp
Pink Shrimp
Pink Shrimp
Pink Shrimp
Pinfish
96 HOUR
LC50 (yg/1)
0.4(0.3-0.6)
0.1(0.1-0.2)
0.7(0.4-1.2)
0.04(.02-.05)
>25
0.1(0.07-0.1)
0.03(0.02-0.04)
0.04(0.001-0.1)
0.2(0.1-0.2)
3.5(2.8-4.4)
0.6(0.5-0.7)
REFERENCE
Parrish et^ al_. ,
Schimmel e_t al . ,
Parrish et aj . ,
Schimmel e_t aj_. ,
Parrish et al . ,
Schimmel ejt aj_. ,
ii
n
ii
Bahner and Nimmo
Schimmel ejt al_. ,
1975
unpubl .**
1973
I974a
1974
unpubl .**
n
n
M
, 19755
unpubl -**
*Less than five species of estuarine animals tested.
**Steven C. Schimmel, Gulf Breeze Environmental Research Laboratory, Gulf
Breeze, Florida 32561
SENSITIVE LIFE STAGE BIOASSAYS AND ENTIRE LIFE CYCLE BIOASSAYS
Chronic bioassays on sensitive life stages and on entire life-cycles
of estuarine organisms are usually conducted to determine the quantity of
chemical that can be tolerated by an organism throughout its life or during
a critical portion of its life. Data from this type of bioassay are es-
pecially important in deriving water quality criteria. Water quality
criteria are most frequently obtained by multiplying the 96-hour LC50 of
the most sensitive species tested by an arbitrary application factor, to
protect that species--and hopefully, the ecosystem-- from chronic effects
of a pollutant. The arbitrary application factor for persistent pollutants
66
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is usually about 0.01 (NAS-NAE Committee on Water Quality Criteria, 1972).
Scientifically derived application factors can be obtained by comparing
data from acute bioassays and bioassays in which a fish or invertebrate
is exposed to the chemical throughout its entire life cycle. The factor
is obtained by dividing the concentration not affecting survival, growth
or reproduction in entire-life-cycle bioassays by the 96 hour LC50 for
that species (Mount and Stephan, 1967; Eaton, 1973).
Sensitive Life Stage Bioassays
Marine toxicologists have not been able to experimentally derive
application factors based on exposures throughout a marine animal's life
cycle because techniques for maintaining cultures throughout entire life
cycles were lacking. Therefore, it is necessary to develop and use meth-
ods that provide toxicity data on sensitive stages of the life-cycle of
saltwater species. Our laboratory has funded grants or contracts to look
at the effects of pesticides on larval development of dungeness crabs,
Cancer magister; blue crabs, Cal1i nectes sapi dus; and a mud crab, Rhithro-
panopeus harrisii. Chemicals that are being or have been investigated in-
clude captan, carbofuran, chlordane, DEF, malathion, methoxychlor, mirex,
propanil, trifluralin, 2,4-D and juvenile hormones. We also supported
research on the effects of methoxychlor and mirex on embryo, larval, ju-
venile and adult striped mullet, Mugil cephalus (Lee et al., 1975).
i
Research on sensitive stages of estuarine organisms at GBERL is pri-
marily on larval and postlarval grass shrimp (Palaemonetes pugijj). and
embryos and fry of the fishes Cypri nodon yari eqatus. Fundulus si mil is, £_.
heteroclitus, Leiqstomus xanthurus, Menidia menidi a and Morone saxatilis.
Recently published papers on this research irTcTude"those of Hansen et al.,
(1975), Middaugh et al., (1975), Parrish et al. (1975) and Schimmel et al.
(1974a, b). This research has primarily been on the effects of toxicants
in water on development and survival of early life-stages. Recent research
(Hansen et al., 1973) on the effects of a PCB, Aroclor 1254 in the eggs of
the sheepshead minnow, C. van'eqatus, indicated that certain concentrations
of PCB in eggs are lethal to embryos and fry (Figure 1). If this PCB af-
fects other fishes similarly, residues exceeding 5 parts per million in
eggs would decrease survival of fry.
Entire Life Cycle Bioassays--
Chronic, entire life-cycle bioassays are routinely conducted by fresh-
water toxicologists, but saltwater toxicologists have only recently developed
similar procedures. Freshwater chronic bioassays can be conducted with blue-
gills (Lepomis macrochirus), fathead minnows (Pimephales promelas), brook
trout (Salvelinus fontinaTis), water fleas (Daphm'a magnaj and other fishes
and invertebrates (Eaton, 1973).
Chronic bioassays using the estuarine fish Cyprinodon yariegatus are
possible (Schimmel and Hansen, 1974). This oviparous fish develops from
an embryo to maturity in about 10 weeks, with about 70% survival overall.
The fish spawns readily in an aquarium, producing about 8 eggs per day
(Figure 2). Total egg production seems unrelated to fish size but fre-
quency of spawning and egg fertility appear to be size-dependent (Schimmel
and Hansen, 1974). Females begin producing eggs at 27 mm standard length.
67
-------
1OO
50%
HATCH
AROCLOR 1254 IN
EGGS (jug/g)
EMBRYOS
Figure 1. Effect of Arocloi®1254 in eggs of sheepshead
minnows on the survival of embryos and fry.
68
-------
In one experiment, 19 fish less than 35 mm long produced an average of
8.2 eggs per day, and 15 fish 35 mm and longer averaged 7.8 eggs per day.
The smaller fish produced eggs more consistently (50% of the days vs. 31%)
with greater fertility (94% fertility vs. 79%) than the larger fish. As a
result of this and other information, a tentative method for entire life-
cycle bioassays using this fish has been suggested (Hansen and Schimmel,
1975). Recently, sheepshead minnows were exposed to endrin and to hepta-
chlor to determine the effect of these pesticides on reproduction.
Sheepshead minnows were exposed to 0.025, 0.077, 0.12, 0.31 or 0.77 yg/1
of endrin measured in water during an entire life cycle bioassay that lasted
25 weeks. This bioassay consisted of three parts: (1) the exposure began
with embryos and continued through embryonic development, hatching of fry and
growth of the fry to adulthood; (2) continued exposure of adult fish to moni-
tor spawning success, including egg production and fertility; and (3) the
bioassay ended following a 28-day exposure of embryos and fry obtained from
spawning fish. The apparatus used was that of Schimmel et al., (1974b) and
the methods were similar to those of Hansen and Schimmel~Tl9T5).
6O
so
y 40
O
w 30
>-
u
Z
Ul
D 20
O
1O
AVERAGE NUMBER =8.O
1-10 11-20 21-30 31-40 41-50
NUMBER OF EGGS PER DAY
Figure 2. Number of eggs spawned by breeding pairs of
sheepshead minnows (Cyprinodon variegatus).
69
-------
Sheepshead minnows were affected by endrin in this entire life cycle
bioassay (Table 3). Embryos in 0.31 and 0.72 yg/i of endrin hatched sooner
than embryos in water free of endrin. Fry in 0.72 yg/i began to die one
day after hatching and all were dead by day 9. Fry in 0.31 yg/l began to
die two days after hatching and over half were dead by day 12. Survival of
juvenile fish was unaffected. Survival of spawning females was reduced in
0.31 yg/1 and their eggs were less fertile than were those of control females,
Survival of fry from eggs spawned by fish exposed throughout their life to
0.31 yg/1 - and possibly 0.12 yg/1 - was decreased.
TABLE 3. EFFECTS OF ENDRIN ON SHEEPSHEAD MINNOWS (CYPRINODON
VARIEGATUS) EXPOSED THROUGHOUT THEIR ENTIRE LIFE-
CYCLE.CONCENTRATIONS OF EXPOSURE WERE: CONTROL,
0.025, 0.077, 0.12, 0.31 and 0.72 yg/1
Generation Life Stage
F- Embryo
Fry
Juveniles
Effect Concentration, yg/1
Early hatching
Death
Decreased growth
No effect
0.31, 0.72
0.31, 0.72
0.31
Adults
Death of spawning
F2
Embryos and
fry
females
Decreased fertility
of eggs
Death
0.31
0.31
0.31
The effects of technical heptachlor on reproduction and development of
Cyprinodon variegatus was studied in a similar experiment, except that it
began with juvenile fish rather than embryos. Measured concentrations of
exposure were 0.71, 0.97, 1.9, 2.8 and 5.7 yg/l of technical heptachlor
(heptachlor and trans-chlordane) and a control. In the first four weeks
of the experiment, some juvenile fish died in 2.8 and 5.7 yg/i of technical
heptachlor. Thereafter, few fish died until the reproductive portion of
the experiment began at week 8. Heptachlor also affected reproduction by
reducing number of spawnings, number of eggs, fertility of the eggs and
survival of fry from fertile eggs.
Experiments are being conducted to determine techniques required to
conduct entire life-cycle bioassays with the grass shrimp (Palaemonetes
pugio). The effect of light and temperature on initiation and success of
spawning has been investigated. Larval and postlarval shrimp have been
used in bioassays to determine effects of certain PCB's on larval develop-
ment and metamorphosis. Results indicate that grass shrimp will spawn
readily, larvae will develop successfully, the species will be sensitive
to toxic chemicals and, therefore, would be excellent for entire life-cycle
bioassays.
70
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COMMUNITY BIOASSAYS
Bioassays can be used to predict how communities of estuarine organisms
will respond to a toxicant. Bioassays in which only one species or organism
is exposed to a chemical can be used to predict how a community may respond
if a number of species from various phyla have been tested under similar con-
ditions. Predictions from this type of data are questionable, particularly if
little is known about how species interact in the community. Predictions can
also be made using data obtained from field studies, but these predictions may
also be questioned because of problems with inadequate controls and lack of
replication. An alternative approach is to conduct laboratory studies in
which communities of organisms are exposed to a chemical and effects deter-
mined. This approach can be valuable if laboratory communities resemble ones
in the field and if enough replicates and concentrations are used so that
statistical analyses can be made and trends observed.
I have completed two bioassays to determine the effects of Aroclo 1^1254,
a polychlorinated biphenyl (PCB), and toxaphene, an insecticide, on the devel-
opment of estuarine communities. The numbers, species and diversity of ani-
mals that grew from planktonic larvae in contaminated aquaria were compared
with those that grew in identical aquaria that were not contaminated. In each
bioassay, sea water with its natural complement of plankton flowed into each
of 10 replicate sand-filled aquaria for each of three toxicant concentrations
and a control. Planktonic larvae colonized the sand and walls of each aquar-
ium. At the end of the experiments—4 months for the PCB, 3 months for toxa-
phene--organisms were collected in a Imm-mesh sieve, preserved and later
identified.
Aroclor(B)l254 altered the composition of communities of estuarine ani-
mals that developed from planktonic larvae in salt water that flowed through
10 aquaria contaminated with 1 or 10 yg/1 (Hansen, 1974). Communities that
developed in 10 control aquaria and 10 aquaria that received 0.1 yg/1, of PCB
for four months were dominated (>75%) by arthropods, primarily the amphipod
Corophiurn yolutator (Figure 3). In aquaria receiving 1 and 10 yg/1 , the num-
ber of arthropods decreased and the number of chordates, primarily the tunicate,
Molgula manhattensis, increased; over 75% of the animals in 10 yg/1 aquaria
were tunicates.Numbers of phyla, species, and individuals (particularly
amphipods, bryozoans, crabs, and mollusks) were decreased by the presence of
this PCB, but there was no apparent effect on the abundance of annelids,
brachipods, coelenterates, echinoderms or nemerteans (Table 43. The Shannon-
Weaver index of species diversity was not altered by Aroc1orCvl254.
TABLE 4. EFFECT OF AROCLOR®1254 ON THE NUMBER OF PHYLA,
SPECIES AND INDIVIDUALS AND ON THE SHANNON-WEAVER
INDEX OF SPECIES DIVERSITY IN COMMUNITIES OF ES-
TUARINE ORGANISMS THAT DEVELOPED IN SAND-FILLED
AQUARIA IN A 4 MONTH BIOASSAY
Aroclor®1254 (,.9/1)
Control 0.1 1 10
Phyla
Species
Individuals
Species diversity
9
52
1776
1.82
7
34
2043
1.26
7
43
1421
2.21
5*
25*
657
1.70
*Statistically different from controls, cc = 0.05.
71
-------
AROCLOR 1254
10O
CONTROL
(MSI/I)
1O.O
Figure 3. Effect of Aroclor® 1254 on the structure of communities of estuarine organisms
-------
In a similar experiment, the insecticide toxaphene also altered the
structure of communities that developed in sand-filled aquaria. Concen-
trations of exposure were 0.1, 1 and 10 yQ/1. The number of mollusks
(primarily gastropods) tripled, annelids [primarily capitellids) doubled
and arthropods were almost eliminated in aquaria contaminated by 10 yg/1
of toxaphene (Table 5). Similar numbers of pelecypods were found in all
aquaria, however, the height (distance from hinge to distal valve edge)
of Morton's cockles (Laevicardium mortoni) was significantly reduced by
10 yg/1 of the insecticide (Figure 4).
TABLE 5. AVERAGE NUMBER OF ANIMALS IN 10 CONTROL AQUARIA
AND 10 AQUARIA THAT FOR THREE MONTHS RECEIVED
0.1, 1 OR 10 ug/1 OF TOXAPHENE. RANGE IN PARAN-
THESIS
Toxaphene
Phylum Control 0.1 1.0 10.
Mollusca 124(65-146) 170(98-274) 142(65-237) 373(245-489)
Annelida 56(19-97) 62(33-90) 66(31-126) 110(82-182)
Arthropoda 32(2-257) 155(1-523) 9(1-63) 0.4(0-1)
Coelenterata 3(0-21) 3(0-19) 10(0-44)
Other 0.1(0-1) 0.1(0-1)
73
-------
40
20
ui
Q.
z
111
D
o
40
20
EFFECT OF TOXAPHENE ON HEIGHT
OF COCKLES
CONTROL
N=437
1.0 MO/ I
N = 467
0-1 MO/1
N = 554
10.0
N = 431
04 8 12 16 20
04 8 12 16 20
HEIGHT (MILLIMETERS)
Figure 4. Effect of toxaphene on the height (distance from hinge to
distal edge of valve) of Morton's cockles collected from
a community of estuarine organisms.
40
20
40
20
74
-------
REFERENCES
Bahner, Lowell H. and Del Wayne R. Nimmo. I975a. A salinity controller
for flow-through bioassay. Trans. Am. Fish. Soc. (In press).
Bahner, Lowell H. and Del Wayne R. Nimmo. 1975b. Methods to assess com-
binations of toxicants, salinity and temperature on estuarine animals.
Proc. 9th Annu. Conf. on Trace Substances in Environ. Health, Columbia,
Missouri, June 10-12, 1975. (In press).
Eaton, John G. 1973. Recent development in the use of laboratory bioassays
to determine "safe" levels of toxicants for fish. In Bioassay Tech-
niques and Environmental Chemistry. Glass, Gary E.H["Ed.). Ann Arbor
Science Publishers, Inc. Ann Arbor, Mich. 48106.
Hansen. David J., Steven C. Schimmel and Jerrold Forester. 1973. Aroclor^
1254 in eggs of sheepshead minnows: Effect on fertilization success
and survival of embryos and fry. Proc. 27th Ann. Conf. S.E. Assoc.
Game Fish. Corrni. 1973. p. 420-426.
Hansen, David J. 1974. Aroclor^'1254: Effect on composition of developing
estuarine animal communities in the laboratory. Contrib. Mar. Sci .
18: 19-33.
Hansen, David J. and Steven C. Schimmel. 1975. Entire life-cycle bioassay
using sheepshead minnows (Cyprinodon variegatus). Fed. Regis-. 40 (123),
part II: 26904-26905.
Hansen, David J^and Steven C. Schimmel and Jerrold Forester. 1975. Effect
of Aroclor(B)l016 on embryo, fry, juvenile and adult sheepshead minnow
( Cypri nodon van' egatus ) . Trans. Am. Fish. Soc. (In press).
Lee, Jong H., Colen E. Nash and Joseph R. Sylvester. 1975. Effects of
mirex and methoxychlor on striped mullet, Mugi 1 cephalus L. U.S.
Environmental Prot. Agency, Ecol . Res. Ser. EPA-660/3-75-015, 18pp.
Middaugh, D.P., W.R. Davis and R.L. Yoakum. 1975. The response of larval
fish, Leiostomus xanthurus, to environmental stress following sub-
lethal cadmium exposure. Contrib. Mar. Sci. (In press).
Mount, Donald I. and Charles E. Stephan. 1967. A method for establishing
acceptable toxicant limits for fish-malathion and the butoxyethanol
ester of 2, 4-D. Trans. Am. Fish. Soc. 96(2): 185-193.
NAS-NAE Committee on Water Quality Criteria. 1972. Water Quality Criteria,
1972. Ecol. Res. Ser. xx + 594 pp. U.S. Environmental Protection
Agency, EPA-R3- 73-033 March 1973. U.S. Gov. Print. Office, Wash.,
D.C. 20402.
75
-------
Parrish, Patrick R., John A. Couch, Jerrold Forester, James M. Patrick,
Jr. and Gary H. Cook. 1973. Dieldrin: Effects on several estuarine
organisms. Proc. 27th Annu. Conf. S.E. Assoc. Game Fish Comm. 1973.
p. 427-434.
Parrish, Patrick R., Steven C. Schimmel, David J. Hansen, James M. Patrick,
Jr. and Jerrold Forester. 1975. Chlordane: Effects on several es-
tuarine organisms. J. Toxicol. Environ. Health. (In press).
Schimmel, Steven C. and David J. Hansen. 1974. Sheepshead minnow
(Cyprinodon variegatus): An estuarine fish suitable for chronic
(entire life-cycle) bioassays. Proc. 28th Annu. Conf. S.E. Assoc.
Game Fish. Comm. (In press;.
Schimmel, Steven C., Patrick R. Parrish, David J. Hansen, James M. Patrick,
Jr. and Jerrold Forester. 1974a. Endrin: Effects on several es-
tuarine organisms. Proc. 28th Annu. Conf. S.E. Assoc. Game Fish.
Comm. (In press).
Schimmel, Steven C., David J. Hansen and Jerrold Forester. 1974b.
Effects of Aroclor(B)l254 on laboratory-reared embryos and fry of
sheepshead minnows (Cypri nodon van' e gat us) - Trans. Am. Fish. Soc.
103(3): 582-586.
76
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The Effect of Subtle Temperature Changes on
Individual Species and Community Diversity
William C. Johnson II and
Eric D. Schneider*
ABSTRACT
Decisions about the regulation of thermal additions
in aquatic systems have often been based on acute
or chronic high temperature effects on individuals
from various life stages of selected species, but
rarely on populations or communities (Mihursky, 1969).
This paper offers a synthesis from a variety of data
bases that demonstrated consistent response of bio-
logical systems to long-term, low level temperature
change. Profound effects of subtle prolonged temper-
ature displacement are demonstrated at the species,
and community levels of biological organization.
INTRODUCTION
For the past decade there has been a controversy as to whether or not
artificial temperature alteration should be considered a type of pollution.
It is of interest that Congress (SR 92-500) considers artificial thermal
perturbations as pollution and requires the United States Environmental
Protection Agency to develop regulations which assure maintenance of bal-
anced and indigenous populations within the area of thermal discharge.
Temperature is known to be a key physical controlling parameter of biolo-
gical systems. Latitudinal distributions of organisms is usually a direct
reflection of the temperature gradient that exists between the poles and
the equator. Therefore, one might hypothesize that in any locale, a per-
sistent temperature change may lead to a biotic change with the possible
loss of desirable species and replacement by others. This paper identifies
the data base which substantiates this hypothesis. These data have been
generated previously by studies related to abundance of commercially valu-
able species, power plant impact studies, and pelagic foraminiferal studies
METHODS
In order to demonstrate consistent response of biological systems to
long-term, low level temperature change, it is necessary only to look at
existing models derived from commercial species abundances and power plant
impact studies with a new point of view. However, utilization of foramin-
iferal data requires some manipulation.
*Environmental Research Laboratory, U.S. Environmental Protection Agency,
Narragansett, RI 02882
77
-------
Analyses of the foramini feral data were designed to determine the
relationship between species diversity of trigger core tops and piston
core tops and the surface water temperatures overlying those cores.
These core top samples represent recent sediments. A history of marine
geology has established the fact that the conditions during deposition
generally correlate well with the present environmental conditions.
Foraminiferal species diversity is calculated from relative abun-
dance for the data sets of Kennett (1967) and Imbrie and Kipp (1971).
Using these data the Shannon-Wiener index, fl, was calculated according
to the formula:
fi = -£ pi log pi
where s = the number of species, pi = the proportion of the i species
in the total number of species, and N = the total number of individuals.
The diversity values were taken directly from Williams and Johnson (1975).
The summer-winter average temperatures presented by Williams and
Johnson (1975) were used. A simple arithmetic mean was calculated from
the summer and winter temperatures presented by Imbrie and Kipp (1971).
For each of Kennett's (1967) sediment core locations the same temperature
term was calculated by averaging summer and winter values from the ocean-
ographic atlas of Schott (1935). Regressions of temperature and diversity
values were made by the least squares method of linear regression.
DISCUSSION
Figure 1 depicts a hypothetical relationship between success or abun-
dance of an individual, a population, or community assemblage, and temper-
ature. The determination of the upper and lower thermal limits, of the
"OPTIMUM CONDITION
LOWER RESPONSE RATE
TO TEMPERATURE THAN
. AT EXTREME
Q
_l
LU
DECLINE
WITH SMALLTEMPERATURE
INCREASE
LfRAPID INCREASE
^U/ITU <
WITH SMALL
TEMPERATURE INCREASE
COLD
HOT
TEMPERATU RE (DURING A CRITICAL PERIOD)
Figure 1. The hypothetical relationship between temperature and an indi-
vidual's, a species', or community success. Different rates of
response for the different levels of biological organization
within their tolerated range are suggested.
78
-------
optimum, and of the temperature range that can be tolerated by the indivi-
dual, a species, or the community are of ecological importance. Bradshaw
(1961), through experiments on foraminifera, suggested that such a curve
is skewed toward the warmer temperatures at the species level. Indeed,
data presented herein indicate a similar tendency toward skewing at the
community level.
Temperature may have intensified effects during particular life history
stages of different species. Such effects are often a result of direct re-
lationships between critical biological processes and physiological effects
of temperature. Rates of processes or timing of critical events are two
types of biological variables affected (Jeffries and Johnson, 1974).
Another important consideration with respect to temperature change is
the alteration of community structure and composition. Consider a hypothet-
ical assemblage of five species which have different ranges, optima, upper
and lower limits, and rates of response to temperature (Figure 2). Commun-
ity structure is predicted at two points within the temperature range by
measuring the relative heights of the individual curves at these tempera-
tures, with species interaction not taken into account. The resulting pre-
diction shows very different community structure at the two temperatures.
The magnitude of temperature change required to cause such dramatic biotic
shifts is the question which will be treated here.
UJ
o
o
z
13
CD
UJ
I
UJ
a:
IUU
80
60
40
20
o
-
_
—
-
\ \
C
B
D
C
E
T, T2
HIGH
/^N.
LOW
TEMPERATURE
Figure 2. Hypothetical temperature ranges for five species and their
relative abundances throughout their respective ranges.
Theoretical community composition at two temperatures is
predicted by the relative species' abundance at the two
temperatures.
79
-------
Direct prediction of community composition from this type of figure
would not be warranted if significant species interaction were to occur.
Species interaction is nearly always present, although the amount and kinds
of such interaction have been difficult to predict in natural situations.
An example of this difficulty is the effect that predation by the green crab,
Carcinus maenas, had on soft clams, Mya ajenaria, in Maine (Glude, 1954).
Here, a slight climatic warming of approximately 2°C led to the introduction
of the green crab, which eliminated entire year classes of the soft clam.
COMMERCIALLY VALUABLE SPECIES DATA
A valuable source of information concerning temperature-biotic relation-
ships may be found in the analysis of fishery catch data. Table 1 (Jeffries
and Johnson, 1975) summarizes the responses of eight marine species to tem-
perature change when univariate models (Dow, 1973) and a bivariate model
(Flowers and Saila, 1972) are applied to catch statistics for these species.
Figure 3 demonstrates the extreme thermal sensitivity of Hawaiian corals'
reproductive success (Jokiel et ^1_., 1974). A generalized two compartment
temperature model is presentecPin Figure 4 (Jeffries and Johnson, 1975).
Their specific winter flounder model (Figure 5) demonstrates how tempera-
ture may affect a species abundance through two independent mechanisms.
1.0
2 .8
0
z
w
OT
Ul
o
o .6
W
Ul
H
O
^^
o A
o
o:
Q.
Ill
UJ
tr
ui 2
z
0
J
•
\
. ^/ \
• i^M
18 22 26 30
MEAN TEMPERATURE °C
Figure 3. Net reproductive success of Hawaiian corals over an experimental
temperature range. Note the limited temperature range for the
optimal reproductive success. (Figure from Jokiel, e_t al_., 1974)
80
-------
TABLE 1. CHANGES IN ABUNDANCE OF SEVERAL SPECIES PREDICTED FROM A 1°C INCREASE IN THE GRAND MEAN,
ANNUAL SEA-SURFACE TEMPERATURE (MAINE). (TABLE FROM JEFFRIES AND JOHNSON, 1975)
00
Author
Dow
(1973)
Flowers
and
Saila
(1972)
Species Change
hard clam +73.4
Mercenaria mercenaria
oyster +51.0
Crassostrea virginica
lobster +15.2
Homarus americanus
shrimp -75.0
Panda! us boreal is
scallop -37.5
Placopecten magellanicus
soft clam -36.8
Mya arenaria
sand worm -32.5
Nereis virens
bloodworm -22.4
Glycera dibranciata
lobster +14.6
Homarus americanus
Basis of annual
comparison; total Average
yield relative to catch
mean temperature (metric T)
- UNIVARIATE MODELS -
Same year 83.3
3 yr later 1.2
3
same year 8.3 x 10
4 yr later 139
6 yr later 135
5 yr later 1.9 x 103
same year 239.2
same year 179.7
- BIVARIATE MODEL -
TQ & T_678{ 9.4 x 103t
Observation
period
1939-1967
1951-1967
1939-1967
1939-1949
1954-1967
1941-1965
1940-1966
1949-1967
1949-1967
1947-1967
rf
.770**
.822**
.627**
-.505*
-.743**
-.643**
-.812**
-.669**
.889**
Effect
Winter
survival ,
near N limit
molting,
recruitment
spawning,
early survival ,
near S limit
predation
on spat
spawning,
early
survival
TQ: molting,
recruitment;
/""TO • Go i I y
winter
mortality
*significant at the 95% level
**significant at the 99% level
of probability; tcalculated from data presented by Dow (1973);
of probability t"T is the mean annua1 sea-surface temperature
y' present year;
T-678 ^s the sum of t^ie mean annua^ sea-surface
of the
temperatures for 6,7, and 8 years previous to T .
-------
TEMPERATURE EFFECTS ON ENTIRE LIFE OF A SPECIES
A SUNDANCE
EARLY PERIOD REMAINDER OF LIFE
DIRECT TEMR EFFECTS DIRECT a INDIRECT TEMP EFFECTS
A.IFAT LIMIT OF RANGE
I SPAWNING ACTIVITY I.MAX OR MIN TEMPERATURES
BEYOND ACCEPTABLE LEVELS
2.EGG VIABILITY
2.MATURATION (REPRODUCTIVE)
3. DEVELOPMENT
B. AND IN GENERAL
4. SUCCESS OF METAMORPHOSIS
3.SUCCESS IN COMPETITION
5. CRITICALTIMING OF MET.
4. GROWTH RATE 8 FEEDING RATE
5. PREDATOR RELATIONS
(CRAB 8 CLAM)
6. MATURATION TO ADULT (LOBSTER)
7. LONGEVITY
8. MIGRATIONAL PATTERNS
9. EFFECT ON HABITAT
Figure 4. Biological evaluation of a generalized bivariate temperature
model of species abundance. Climatic temperature change is
the basis for this model. (After Jeffries and Johnson, 1975)
Taylor et_ _al_. (1957) noted a close relationship between shifts of the
geographic ranges in marine fishes and other species in the Northwestern
Atlantic and climatic warming. Dow (1964, 1967, 1969, 1971, 1973) has been
a key contributor on this topic by observing a relationship between temper-
ature and catch statistics of eight commercially valuable marine species.
Considering the response model of Figure 1, the logical question is, at
what rate does species abundance respond to a given prolonged temperature
change. Jeffries and Johnson (1974, 1975) have attempted to quantify the
abundance of several single species with respect to temperature change
(Table 1). These data suggest that as little as a one degree centigrade
positive displacement from the long-term mean coastal surface temperature,
may result in significant changes in species abundance, ranging from +73.4%
to -75%. The extreme response rates are for populations near their thermal
limits, while those living closer to their temperature optimum generally
show smaller change with this temperature rise.
Other models of abundance for marine fish species have been developed
which use temperature as an influential term. Sissenwine (1974, 1975), for
instance, developed fisheries models for yellowtail flounder in New England
which incorporated an important temperature effect. His models indicate
extreme sensitivity of recruitment and growth to minor temperature change
(+ 1°C).
82
-------
Temperature can play an exaggerated role during the early life history
of many species. Figure 3 demonstrates this sensitivity in Hawaiian corals
(Jokiel et. dj_., 1974), where an optimum reproductive range of only 2°C was
observed. Baird (1953) suggested that there was a similar sensitivity dur-
ing the spawning of the giant scallop, Placopecten magellanicus. a temperate
species. A similar effect has been identified by Jeffries and Johnson (1975)
in modeling the winter flounder population, Pseudopleuronectes americanus,
in Narragansett Bay, Rhode Island. Thus, data from both temperate and
tropical locations support the hypothesis of ecologically significant early
life history temperature sensitivity for a variety of species.
A quantitative bivariate population model can give more information, in
some instances, than univariate counterparts by offering more insight into
the effects of temperature at different times of a species life history. A
generalized and a specific model are presented in Figures 4 and 5. Flowers
and Saila (1972) employed a similar bivariate temperature model to describe
lobster abundance in Maine. These two different types of models are applied
here only with respect to their value in pointing out apparent temperature
sensitivity of life stages and without bias toward any particular scheme o1
modeling.
of
TEMPERATURE EFFECTS ON WINTER FLOUNDER
(PSEUDQPLEURQNECTES AMERICANUS)
JEFFRIES AND JOHNSON (1975)
ABUNDANCE
'I
APRIL TEMP.
30 MONTHS PRIOR
TO CATCH
DIRECT EFFECTS
I. DEVELOPMENT
RATE
2.SURVIVALRATE
TEMP HISTORY FROM MAY,
29 MONTHS PRIOR TO CATCH
THRU SEPT., FIRST MONTH
OF CATCH
INTERACTIONS
I.COMPETITION
FOR RESOURCES
2.PREDATION
Figure 5. A specific bivariate temperature model for the abundance of
winter flounder (Pseudopleuronectes americanus) developed by
Jeffries and Johnson (1975).Possible biological interactions
are suggested.
83
-------
POWER PLANT STUDIES
A second source of data on the effect of temperature on populations
and communities is that found in the literature concerning power plant im-
pact. Figure 6 depicts the effect of temperature on the relative abundance
of benthic microalgae in Hawaiian waters (Jokiel et. a\_., 1974). Figure 7
(Jokiel et. al_., 1974) shows that in selected species systems, Hawaiian
macroalgae standing crop drops nearly five fold with an increase in tem-
perature of 4.3°C.
HIGH
UJ
o
z
<
o
z
^>
CD
111
UJ
ee
LOW
Figure 6.
A NJTZCH1ATE
DIATOMS
B FILAMENTOUS
GREENS
C NAVICULATE
DIATOMS
D FILAMENTOUS
BLUE-GREENS
30
34
TEMPERATURE °C
The effect of temperature on the abundance of benthic microalgae
on Hawaiian waters. This figure resembles Figure 2 and demon-
strates the theorized community shift. (Figure from Jokiel
et al_., 1974)
From Figure 6 one can see that two of the subtropical algal types have
optimal ranges of less than 5°C. It is also of interest to note that this
figure resembles Figure 2 and that temperature displacement results in com-
munity shifts as hypothesized. Beyond these algal species, the entire
marine communities in Hawaii and Florida have been observed to undergo major
species compositional changes as a result of prolonged local warming from
heated power plant discharge (Jokiel <2t aj_., 1974, Roessler et. aj_., 1974).
Caution should be applied when using a tropical response to thermal stress
as indicative of a general pattern, as this biota typically lives close to
their upper thermal limit. However, these data agree with the temperate
examples already cited (Taylor e_t aj_., 1957, Jeffries and Johnson, 1974).
A marked increase or decrease in net macroalgal production resulting
from persistent temperature displacement could be responsible for some of
the observed fauna! changes (Roessler e_t al_., 1974). Although the direc-
tion of change is probably not typical, limitations of macroalgal production
84
-------
30 DAYS SUMMER AMBIENT ("27"C) 30DAYS STRESS
500
| | OTHERS
H| ACANTHOPHORA
4001- ^H ULVA
300
o
UJ
* 200
100
TANK 10
26.7
TANK 7
28.5
TANK 4
29.6
TANK I
31.0
AVERAGE TEMP.'C DURING STRESS PERIOD
Figure 7,
Hawaiian macroalgae standing crop at four experimental tempera-
tures. Note the dramatic decrease in biomass with only a 4.3°C
shift. (Figure from Jokiel et_ al_. , 1974)
3.5 r
2.7
x
UJ
o
to 1.9
CL
Ul
CL
UJ
z
o
0.3
-0.5
• • •
-5.0
0.0
5.0
10.0
15.0
Figure 8.
TEMPERATURES
Scatter diagram of Shannon-Wiener diversity values, calculated for
the Kennett data set (1967), plotted versus the summer-winter aver-
age temperature. Note the linear trend of increasing diversity
with increasing temperature.
85
-------
with increased ambient temperatures in the tropics provide an example (Fig-
ure 7). It is worth noting from this figure that there is a significant
shift in relative abundances of the three different groups of algae over
the observed temperature range (4.3°C). Thus, in addition to altered pro-
ductivity, a shift in species composition occurs which affects the amount
and quality of algal habitat afforded to a segment of the animal community.
FORAMINIFERAL INVESTIGATIONS
Another data source showing sensitive temperature-fauna! interaction
is the foraminiferal data base. The optimal and total temperature ranges
for eighteen species of foraminifera are presented in Table 2 (Be' and
Tolderlund, 1971). The effect of temperature as an independent variable
on the Shannon-Wiener diversity index values (R) for the foraminiferal
abundance data of Kennett (1967), Imbrie and Kipp (1971), and Williams
and Johnson (1975) are shown in Figures 8, 9, and 10. Figure 10 also shows
the faunal composition of three cores with average temperatures near the
mean of the diversity data set. Figure 11 summarizes those data by drawing
in the individual regression lines.
TABLE 2. AVERAGE TOTAL AND OPTIONAL SURFACE TEMPERATURE
RANGES FOR EIGHTEEN SPECIES OF PLANKTONIC FORA-
MINIFERA IN THE ATLANTIC AND INDIAN OCEANS (BASED
ON DATA FROM BE'AND TOLDERLUND 1971)
Average Range
Species Name
Optimum °C
Total °C
Globigerina pachyderma
G. quinqjjeloba
G. bulloides
Globorotalia inflata
G. truncatulinoides
G. crassaformis
G. hirsuta
G. menardii
Globigerinita glutinata
Globoquadrina dutertrei
Orbulina uni versa
Globigerinella aequi lateral is
Hastigerina pelaqica
Globigerinoides ruber
G.. conglobatus
G. sacculifer
Pulleniatina obliquiloculata
Candeina nitida
8.2
7.9
9.2
5.4
2.5
1.0
8.0
3.4
2.3
2.5
5.7
3.0
7.0
7.5
4.4
3.0
19.4
18.9
26.0
23.5
16.7
8.5
8.2
12.4
25.7
16.7
17.5
15.4
12.2
14.8
11.8
13.4
10.0
7.4
Grand Average
5.1
15.5
86
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3.5
X
UJ
Q
t 3.0
o>
(T
UJ
Q
a:
UJ
2.5
2.0
i
to
1.5
18.0
20.0
22.0
24.0
26.0
28.0
DIVERSITY VS. SUMMER-WINTER AVERAGE TEMPERATURE
IMBIEaKIPP(l97l)
Figure 9. Scatter diagram of Shannon-Wiener diversity values, calculated
for the Imbrie and Kipp data set (1971), plotted against the
summer-winter average temperature. Note that at summer-winter
average temperature greater than 18.5°C, there is a decline in
species diversity.
87
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4.0
'5 3.4
x
U)
Q
Z
FAUNAL COMPOSITION AT 3 TEMPERATURES
HOO
Q
£E
2.2
UJ
z
o
1.6
X
V)
1.0
4.0 8.0 I2JO 16.0
TEMPERATURE °C
20.0
10.4 12.6
145
TEMPERATURE°C
(AVERAGE 12.0 °C)
Figure 10. Scatter diagram of Shannon-Wiener diversity values of Williams
and Johnson (1975) plotted with respect to the summer-winter
average temperature. Here, as in Figure 8, there is an increase
in species diversity with increased temperature. Faunal compo-
sition at three temperatures is presented as well. The faunal
composition at 14.5°C was calculated by averaging data from two
cores. Note the faunal shifts, additions, and deletions of
species as the temperature changes.
88
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-404 8 IE 16 20 24
SUMMER-WINTER AVERAGE TEMPERATURE °C
AUTHOR NO. PTS. OCEAN _r
I. IMBRIE8 KIPP(I97I) 18 N a S ATLANTIC .834
2. IMBRIE»KIPP(I97I) 41 N 8 S ATLANTIC .805
3. KENNETT (1967) 33 SW PACIFIC .928
4. WILLIAMS a JOHNSON 20 S INDIAN .973
(1975)
28
Figure 11. Individual regression lines of diversity versus summer-winter
average temperature. Note their close correspondence despite
different sources of data. These lines approximate the shape
of the hypothetical community response curve demonstrated in
Figure 1.
89
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Investigators in the micropaleontological field pioneered the inves-
tigation of the effects that small scale temperature fluctuations have on
the abundances of individual species. As early as 1935, Schott made the
observation that pelagic foraminifera could be utilized as markers of cli-
matic change. Other early investigators (Ericson, 1959; Ericson and Wollin,
1956) noted fauna! changes in response to climatic trends^ For those spe-
cies examined in Table 2, the range for the optima is 5.1°C and the average
total range is 15.5°C.
Kennett (1970), Ruddiman (1971), Imbrie and Kipp (1971), and Hecht
(1973) are but a few of the investigators who have developed sophisticated
paleotemperature models from relationships observed between temperature and
living foraminiferal species assemblages. Although developed to satisfy
quite another need, these models and their data bases underscore impact on
living foraminiferal populations resulting from minor, but persistent, tem-
perature change.
The foraminiferal data possibly demonstrate community structure sen-
sitivity to long-term temperature changes. This ubiquitous data base was
chosen because it lends itself to global comparisons. Species diversity
is used as an index of community structure. Although species diversity
should not be considered a panacea for community studies, it can serve as
a simple common denominator for comparing the voluminous foraminiferal
data sets.
The scatter diagrams (Figures 8, 9, and 10) show how well diversity
approximates linear trends with respect to temperature. There is a rapid
rate of change for each of the data sets. It is notewrothy that signifi-
cant shifts of relative species abundances (Figure 10) occur even within
the 4.0°C range examined. This type of faunal shift persists throughout
the data sets and is assumed not to be fortuitous in this example.
Figure 11 demonstrates that a break in diversity occurs generally at
a summer-winter average temperature of 18.5°C. Thus, species diversity is
not linearly related to temperature throughout the entire temperature range
studied. The overall shape of these lines approximates the theoretical
shape of temperature's effect on community success (Figure 1) with the mod-
ification of slight skewing toward the warmer temperatures. This W9uld
imply that species at their southern limits and tropical or subtropical com-
munities would have greater sensitivity to warming trends than species well
within their temperature ranges and temperate or polar communities.
In summation, by applying available modeling tools to completely dif-
ferent existing data sets, we demonstrate appreciable effects of long-term,
low level temperature change on species abundance and community structure
and composition. This demonstration is achieved in a manner which is amen-
able to regulatory requirements generated by Public Law 92-500 (1972) in
sections 304 and 316a. These laws direct the United States Environmental
Protection Agency to develop regulations to control thermal discharge with
limitations more stringent than necessary in order to assure the protection
and propagation of a balanced, indigenous population of shellfish, fish,
and wildlife in and on the body of water into which the discharge is made.
90
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Jeffries and Johnson (1974, 1975) have employed natural fluctuations
of temperature as the stressor in their models. As mentioned earlier, part
of the east coast of North America has undergone short-term warming trends,
giving rise to changes in species abundance (Table 1). With the building
of large scale thermal nuclear power plants on estuarine and coastal waters,
we may see regional temperature rises in the ranges discussed in this paper
(± 1°C). Broecker (1975) has projected that, because of atmospheric C0?
increase, there will be a global warming of 1.1°C by the year 2010 A.D.
Such man induced warming coupled with nature's unpredictable fluctuations
should have appreciable effects on indigenous marine populations.
It is apparent that the biota is very sensitive to slight long-term
temperature alteration. Seasonal cycles and daily fluctuations of temper-
ature tend to obscure long-term temperature trends and their biological
effects, which perhaps explains why many investigators have overlooked this
delicate, but dramatic relationship. Data obtained from a global spectrum
form a basis of support for this suggested relationship. Although these
are preliminary findings, the same overbearing theme persists.
In the future, we would hope to achieve predictive models. At present,
the changes which have occurred, due to thermal pollution or climatic changes,
have been somewhat unpredictable. As mentioned earlier, newly introduced
predators may cause significant changes in the indigenous populations.
CONCLUSIONS
Responses due to thermal perturbation, seen at the species level of
organization, include:
1) Species often have average yearly thermal ranges of about 5°C when
in natural surroundings. However, subtle shifts of ±1°C give rise
to marked changes in population abundances.
2) Response to temperature displacement is appreciable in the resident
biota. Some species benefit by temperature change while others are
adversely affected by the same.
3) Early life history stages may be particularly sensitive to tempera-
ture changes.
The community level of organization appears, also, to be very much
affected by temperature change, as shown by:
1) Entire communities have undergone significant change when prolonged
low level warming occurs because of the discharge of heated effluents
2) Prediction of community changes is confounded by unsuspected inter-
action between species, as in the case of the green crab and soft
clam in Maine.
3) Species diversity changes with ambient temperature for certain com-
munity types. Diversity in foraminiferal communities respond ap-
preciably to temperature change with an apparent maximum at 18.5°C
summer-winter average.
91
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ACKNOWLEDGMENTS
We thank Dr. H.P. Jeffries, Dr. D.C. Miller, Dr. J.P. Kennett, and
D.F. Williams who offered advice and graciously made data available.
Dr. O.K. Phelps critically reviewed the manuscript and offered assistance.
F.T. Short wrote the program for the computation of diversity indices.
REFERENCES
Baird, F.T., Jr., "Observations on the Early Life History of the Giant
Scallop (Pecten magellanicus)", Res. Bull. 14 Maine Dept. of Sea and
Shore Fish.. (1953).
Be', A.W. and D.S. Tolderlund, "Distribution and Ecology of Planktonic
Foraminifera." IN: The Micropalaeontology of Oceans, Edts. Funnel,
B.M. and W.R. RieSel, pp. 105-149 (1971).
Bradshaw, J.S., "Laboratory Experiments on the Ecology of Foraminifera",
Cushman Found. Foram. Res. Contr., 12, pt. 3, pp. 87-106 (1961).
Broecker, W.S., "Climatic Change: Are We on the Brink of a Pronounced
Global Warming?", Science, 189, pp. 460-463 (1975).
Dow, R.L., "Changes in Abundance of the Marine Worm, G1ycera dibranchiata,
Associated with Seawater Temperature Fluctuations' , Commercial Fisheries
Review, 26. No. 8 (1964).
Dow, R.L., "Temperature Limitations on the Supply of Northern Shrimp
(Pandalus borealisj in Maine (USA) Waters", IN: Proceedings of the
Symposium on Crustacea, Part IV, pp. 1301-1304 (1967).
Dow, R.L., "Cyclic and Geographic Trends in Seawater Temperature and
Abundance of American Lobster", Science, 164, pp. 1060-1063 (1969).
Dow, R.L., "Periodicity of Sea Scallop Abundance Fluctuations in the Northern
Gulf of Maine", National Fisherman Res. Bull. No. 31, Maine Dept. of
Sea and Shore Fisheries (1971).
Dow, R.L., "Fluctuations in Marine Species Abundance During Climatic Cycles",
MTS Journal. 7. No. 4, pp. 38-42 (1973).
Ericson, D.B., "Coiling Direction of Globierina pachyderma as a Climatic
Index", Science. 130, pp. 219-220 (1959).
Ericson, D.B. and G. Wollin, "Micropaleontological and Isotopic Determinations
of Pleistocene Climates", Micropaleontology. 2. pp. 257-270 (1956).
Flowers, J.M. and S.B. Saila, "An Analysis of Temperature Effects on the
Inshore Lobster Fishery", J. Fish. Res. Bd. Canada, 29. No. 8,
pp. 1221-1225 (1972).
92
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Glude, J.B., "The Effects of Temperature and Predators on the Abundance of
the Soft-Shell Clam, Mya arenaria, in New England", Trans. Am. Fish.
Soc. 84. pp. 13-26 (1954).
Hecht, A.D., "A Model for Determining Pleistocene Paleotemperatures from
Planktonic Foraminifera Assemblages", Micropaleontology 19. pp. 68-77
^ I y / o / •
Imbrie, J. and N.G. Kipp, "A New Micropaleontological Method for Quantitive
Paleoclimatology: Application to a late Pleistocene Caribbean Core",
IN: The Late Cenozoic Glacial Ages, Ed. K.K. Turekian, Yale University,
flew Haven, pp. 71-182 (1971).
Jeffries, H.P. and W.C. Johnson, "Seasonal Distribution of Bottom Fishes
in the Narragansett Bay Area: Seven-Year Variations in the Abundance
of Winter Flounder (Pseudopleuronects americanus)", J.Fish. Res. Bd.
Canada. 31. pp. 1057-1066 (1974).
Jeffries, H.P. and W.C. Johnson, "Petroleum, Temperature and Toxicants:
Examples of Suspected Responses by Plankton and Benthos in the Contin-
ental Shelf", (In Press, 1975).
Jokiel, P.L., S.L. Coles, E.B. Guinther, G.S. Key, S.V. Smith and S.J. Townsley,
"Effects of Thermal Loading on Hawaiian Reef Corals", EPA Project 18050
DON, element 1B1022, (Progress Report, 1974).
Kennett, J.P., "Distribution of Planktonic Foraminifera in Surface Sediments
Southeast of New Zealand", Proc. 1st Int. Conf. Plankt. Microfossils, Geneva,
pp. 307-322 (1967).
Kennett, J.P., "Pleistocene Paleoclimates and Foraminiferal Biostratigraphy
in Subantarctic Deep-Sea Cores", Deep Sea Research, 17, pp. 125-140 (1970).
Mihursky, J.A. (chairman) "Patuxent Thermal Studies—Summary and Recommendations"
NRI Spec. Report. No. 1, pp. 1-20 (1969).
Roessler, M., D. Tabb, R. Rehrer, and J. Garcia, "Studies of Effects of Thermal
Pollution in Biscayne Bay, Florida", Ecological Research Series, EPA-
660/3-74-014, 145 p. (1974).
Ruddiman, W.F., "Pleistocene Sedimentation in the Equatorial Atlantic: Strat-
igraphy and Fauna! Paleoclimatology", Geological Society of American Bui 1 .,
82, pp: 283-302 (1971).
Schott, G., "Geographic des Indischen und Stillen Ozeans", Boysen, Hamburg,
413 p. (1935).
Schott, W., "Die Foraminiferen in dem Aequatorialen Tie! des Atlantischen
Ozens: Deutsch Atlant. Exped., 'Meteor1, 1925-1927", Wiss. Ergebn, 3,
pp. 43-134 (1935).
93
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Sissenwine, M.P., "Variability in Recruitment and Equilibrium Catch of the
Southern New England Yellowtail Flounder Fishery", J. Cons. Int..
Explor. Mer., 36. No. 1, pp. 15-26 (1974).
Sissenwine, M.P., "Yellowtail Flounder Dynamics", University of RI. Ph D
Dissertation, 178 p. (1975). '
Taylor, C.C., H.B. Bigelow, and H.W. Graham, "Climatic Trends and the Distri-
bution of Marine Animals in New England", U.S. Fish Wild!. Serv. Fish-
Bull. 57. No. 115, pp. 293-345 (1957). '—
United States Congress Senate Fed. Water Pollution Control Act Amendments
of 1972. Pub. Law 92-500, 92nd Cong., 2nd Session, S.2770, Sec. 304,
316a, pp. 35-38, 61 (1972).
Williams, D.F. and W.C. Johnson, "Diversity of Recent Planktonic Foraminifera
in the Southern Indian Ocean and Late Pleistocene Paleotemperatures",
Quaternary Research, 5. pp. 237-250 (1975).
94
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FRESHWATER QUALITY CRITERIA RESEARCH
OF THE ENVIRONMENTAL PROTECTION AGENCY
Donald Mount, presiding
Director, Environmental Research Laboratory--Duluth
Models for Transport and Transformation of Malathion in
Aquatic Systems
James W. Falco, Donald L. Brockway, Karen L. Sampson,
Heinz P. Kollig, and James R. Maudsley
Shagawa Lake Recovery Characteristics as Depicted by
Predictive Modeling
D.P. Larsen and H.T. Mercier
A Mathematical Model of Pollutant Cause and Effect in
Saginaw Bay, Lake Huron
William L. Richardson and Victor J. Bierman, Jr.
Mathematical Model of Phytoplankton Growth and Class
Succession in Saginaw Bay, Lake Huron
Victor J. Bierman, Jr. and William L. Richardson
Implications of Resource Development on the North
Slope of Alaska with Regard to Water Quality on the
Sagavanirktok River
Eldor W. Schallock
Lake Eutrophication: Results from the National
Eutrophication Survey
Jack H. Gakstatter, Marvin 0. Allum and
James M. Omernik
95
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Models for Transport and Transformation
Of Malathion in Aquatic Systems
James W. Falco, Donald L. Brockway
Karen L. Sampson, Heinz P. Kollig, and
James R. Maudsley*
ABSTRACT
A mathematical model has been developed for predicting
the fate and transport of malathion in riverine aquatic
ecosystems. Two competing degradation pathways were
modeled—alkaline hydrolysis and microbial breakdown.
Incorporating data obtained from previous laboratory
studies, the model was used to verify proposed degrada-
tion mechanisms by predicting the behavior of malathion
in the AEcoS, a physical system designed to simulate en-
vironmental conditions as closely as possible. Although
in general results were similar for the two systems,
rates measured in the environmental simulator were slow-
er than those measured in laboratory studies.
INTRODUCTION
The fate and transport of toxic substances in aquatic ecosystems has
been the subject of numerous studies over the years. Field studies are un-
dertaken to estimate the persistence of the toxic materials under natural
conditions and laboratory studies are undertaken to study the persistent
pollutants. In recent years mathematical modeling has also become a useful
technique in the study of environmental pollution. As modeling capabilities
have improved, mathematical models have been applied with increasing fre-
quency to interpret the results of both laboratory and field experiments
and to extrapolate results obtained to other ecosystems.
*Environmental Research Laboratory, U.S. Environmental Protection Agency,
Athens, GA 30601
97
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Such a model was developed to describe the fate and transport of
malathion {0,0-dimethyl S-(l ,2-dicarbethoxy)ethylphosphorodithioate}, an
organo-phosphorus pesticide widely used for control of mosquitos and ag-
ricultural insect pests. A number of laboratory studies have been done
to determine the pathways and rates of chemical and biological degradation
of the insecticide. Ferguson (1975) reviewed the available literature con-
cerning malathion including its chemistry, pharmacology, toxicity, fate and
significance in the environment, and production and use.
Koivistoinen and Aalto (1970) have reported pseudo-first-order rate
coefficients for chemical hydrolysis of malathion as a function of pH
ranging from 1 to 9 and of temperature ranging from 20°C to 70°C. Under
alkaline conditions, malathion hydrolyzes to form predominantly dimethyl
phosphorodithioate and dimethyl phosphorodithionate. Wolfe e_tal_. (in
press) have done a detailed study of chemical degradation pathways and
nave estimated rate coefficients for intermediate product degradation as
well as for malathion hydrolysis. They note that at low temperatures mala-
thion monoacid is formed as an intermediate product that is more stable
than the parent compound and consequently has a potential environmental
impact.
Paris ejt a]_. (1975) completed a microbial degradation study in which
mixed bacterial cultures isolated from the field were inoculated into me-
dium containing malathion as a sole carbon source. From the rate data ob-
tained, malathion degradation was modeled using Monod kinetics (Stumm-
Zollinger and Harris, 1971), and maximum degradation rate, half-saturation
constant, and yield factor were obtained by least squares fit.
All laboratory studies are simplifications of the actual phenomena
occurring in nature. They are well-defined and controllable, but suffer
from the fact that they contain only a few compartments and consequently
important systems interactions may be unobserved. In field studies, how-
ever, the extreme variability and uncontrollability make mechanistic stu-
dies of the ecosystem difficult. A physical model, the Aquatic Ecosystem
Simulator (AEcoS), has been developed to bridge the gap between laboratory
and field studies. The facility was designed to simulate the complexity
of natural field systems as closely as possible, thus providing the realism
of a field study with the controllability of a laboratory system.
To test the mathematical model, results obtained by Wolfe e_t al_. (in
press) for chemical hydrolysis of malathion and results obtained by Paris
et^al. (1975) for microbial degradation were incorporated into the model.
Based on simulations using laboratory coefficients a series of AEcoS ex-
periments were run to verify the proposed mechanisms.
MATERIALS AND METHODS
The facility used in the experiments consisted of a channel, 19.5 m
long by 46 cm wide by 46 cm deep, enclosed in an environmentally controlled
chamber. A detailed description of the facility has been presented by
Sanders and Falco (1973).
Bacterial cultures were kindly supplied by Doris F. Paris, Environ-
mental Research Laboratory, U.S. Environmental Protection Agency, Athens,
98
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Georgia. In all experiments, bacteria were continuously inoculated into
the channel inlet from a chemostat as shown in Figure 1. Bacterial cul-
tures were maintained in the feed chemostat on 1/100 strength nutrient
broth to which was added malathion in quantities to sustain a nominal chemo-
stat effluent concentration of 0.5 mg/1.
Figure 1. End view of the channel showing bacteria feed
system.
The water supplied to the channel was once deionized and once distilled,
In microbial degradation experiments, compounds were added in the ratios
shown in Table 1, and in quantities designed to sustain the inlet concen-
tration of o-POit phosphorus at 90 yg/1, N03 nitrogen at 300 yg/1, NH3 nitro-
gen at 450 yg/1, glucose at 6.0 mg/1, and malathion at 1.0 mg/1. Secondary
reagent grade (97% pure) malathion, provided by American Cyanamid Co., was
used. All other compounds were reagent grade.
99
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TABLE 1. RELATIVE AMOUNTS OF COMPOUNDS IN NUTRIENT MEDIA
Compound Ratio to Amount of (NHit)2SOt+ Added
Na2HP04«7H20 0.195 mg/mg
KH2 P04 99.4 yg/mg
KN03 1-02 mg/mg
MgS04-7H20 10.3 yg/mg
Glucose 1.39 mg/mg
Hunter's Trace Solution 0.938 yl/mg
In the alkaline hydrolysis experiments, the same nutrient composition
was used with two exceptions: the solution contained no glucose and the
ratio of nitrate to ammonia was increased from 0.667 to 1. Absolute con-
centrations of nutrients at the inlet to the channel were maintained at
30 yg/1 o-PO^ phosphorus, 150 yg/1 NHq nitrogen, and 150 yg/1 N03 nitrogen
during the hydrolysis experiments. Iris buffer was also added continu-
ously in the channel inlet to maintain desired inlet pH.
The feed system that supplied nutrients and malathion to the channel
consisted of two closed 40-liter carboys, one containing concentrated nu-
trient solution and one containing concentrated malathion solution. For
alkaline hydrolysis experiments a third carboy was used to supply the
channel with tris buffer. These solutions were pumped to the channel in-
let by peristaltic pumps, which regulated the supply of nutrients and
malathion. Nutrient stocks were autoclaved after preparation and malathion
solutions were filter-sterilized after preparation. A bacterial contamin-
ant, however, was found in some malathion stocks that filter-sterilization
did not eliminate. To eliminate this contaminant, an aliquot of acetone
was mixed with the malathion sample and the mixture was allowed to stand
for approximately one hour, during which the bacteria cells were lysed.
Acetone was removed by evaporation.
The nutrient feed system for the chemostat was similar to the channel
feed system with the exception that only one carboy containing both nutri-
ent and malathion was used. Influent nutrient broth concentration was se-
lected to yield approximately 1 x 108 bacteria/ml in the chemostat effluent.
Flow rate through the chemostat was set at 1 ml/min and consequently the
bacteria count in the channel influent was 1.9 x 105 cells/ml.
The procedure for start-up of microbial degradation experiments was as
follows:
1. Filled chemostats were inoculated with bacteria and allowed to
100
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stand for 24 hours to develop the culture.
2. Channel flow was set at 0.525 liters/min and paddle wheel rota-
tion speed was set at 2 rpm.
3. Chemostat flow was started and effluent stream was directed into
channel inlet.
4. Twenty-four hours after the start-up of chemostat flow, nutrient
and malathion flows into the channel inlet were started.
5. The channel was operated at least eight days to allow the system
to come to steady-state.
The procedure for start-up of chemical degradation experiments was the
same as for the microbial degradation experiments with the exception that
steps 1 and 3 were deleted and nominal channel flow was set at 1.31 liter/
min.
In the first microbial degradation experiment, both air and inlet water
temperatures were set at 22°C. In the second experiment, temperatures were
set at 27°C, and in the third and fourth experiments, temperatures were set
at 32°C. All chemical degradation experiments were carried out at 27°C.
During the transient period of all tests, chemical determinations and
bacteria counts were measured at nine equally spaced sampling locations
along the length of the channel at least once a day. During the steady-
state period, four sets of chemical determinations and bacteria counts were
made at each of the nine sampling stations. Determination of o-OPi* phos-
phorous, N03 nitrogen, NH3 nitrogen, and glucose were accomplished by an
automated auto-analyzer system described by Kollig (in preparation). Mala-
thion analysis was accomplished by extracting water samples with 2,2,4-tri-
methylpentane (isooctane) and analyzing the extract by gas-liquid chromatog-
raphy. Determinations were performed using a Tracer MT550 GC with a nickel
63 electron capture detector. A 1.8 meter long by 0.64 cm diameter column
packed with a 3% S.E. 30 on 80/100 mesh Gas Chrom Q was used at a column
oven temperature of 220°C.
Viable bacteria concentrations were estimated by plate counts (Stan-
dard Methods, 1965). Tryptone-glucose-extract agar was used as plating
medium and the cultures were incubated at 32°C. Species existing in water
samples taken during two microbial experiments were confirmed as
• Flavobacterium meningosepticum
• Xanthomonas species
• Comanonas terrigeri
• Pseudomonas cepacia
A low background level of other bacteria (three Bacillus species) was
observed during the experiment.
Water temperatures were recorded at the nine sampling stations at least
once per day during each experiment.
-------
Water flow rate and relative humidity were also recorded daily.
MATHEMATICAL MODEL
The continuity equation describing the movement and transformations
of material is well known. For one dimensional incompressible flow des-
cribing the flow regime in our channel experiments and in many natural
riverine ecosystems, the equation is
9C. D 32C. v 3C.
_L = L - —L + S. ± £ R..
ax n j 1J
where C. = concentration of constituent i
D = dispersion coefficient
R.. = rate of production or elimination of constituent i
1J by pathway j
S. = source strength of component i
t = time
x = distance in the direction of flow
In equation 1, no distinction is made between point source loads and non-
point source loads. A point source load is simply described by a Dirac
delta function.
The major effort in modeling is usually directed toward development of
an adequate representation for R.JJ. In the case of malathion, two competing
processes occur, namely, alkaline hydrolysis and bacterial degradation.
Wolfe et_ aj_. (in press) modeled the degradation of malathion by alka-
line hydrolysis as a second-order reaction, i.e.,
Rhydrolysis = klCOHCM (2)
where k. = second-order rate coefficient
CQM = concentration of hydroxide ion
C^ = concentration of malathion
They note that two competing temperature dependent reactions occur. An
elimination reaction favored at elevated temperatures results in the pro-
duction of diethyl fumarate and 0,0-dimethyl-phosphorodithioic acid. The
102
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second reaction, favored at low temperatures, results in an intermediate
malathion monoacid product that has environmental significance because of
its persistence in the environment.
In modeling alkaline degradation, therefore, we have separated the
rate coefficient into two contributions
kelim + khydrol
Using the data provided by Wolfe et a! . (in press), we fit each of the
coefficients to the following equations:
kel1n, ' A, exp <-i> (4)
k,, , = B, exp {- S2.} (5)
hydrol l H T ^ '
where Als A2, B19 and B2 are fit coefficients and
T = Temperature (°K)
Paris et_ al_. (1975) proposed two models to describe the bacterial
degradation of malathion. For the first they used the standard Monod ex-
pression for growth of an organism and limiting substrate utilization.
The malathion degradation equation for this model is
RMal = " VCB'CM (6)
Y(K + CJ
m M
where Y = yield coefficient
ym = maximum degradation rate
K = half saturation constant
m
CD = bacteria concentration
D
The corresponding equation for bacterial growth on malathion is
n _ ^m'^R'^M (7}
Bacteria -^—-—- ^''
103
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The second model proposed is a simple second-order equation, which assumes
that microbial degradation can be described by the equation
RMal
where k2 = specific microbial degradation rate
In the model we have developed, equation 6 has been generalized to
account for degradation by any number of species i as follows:
R.,
j
and microbial growth has been generalized to account for utilization of j
carbon substrates simultaneously as follows:
y. . C. C.
R.. . s 1J 3 J (10)
" J K1J + CJ
A flow sheet of the computer program developed for the model is shown in
Figure 2. Input data include length of river reach, average velocity, and
longitudinal dispersion coefficient. For the finite difference equation used
to approximate equation 1, the river reach is divided into 56 equally spaced
segments. First- and second-order spatial gradients at all interior points
are approximated by a third- order central difference approximation (Salvadori,
1961 ). Spatial derivitives are approximated in the first segment of the
reach by a sixth-order forward difference equation and in the last segment
of the reach by a sixth-order backwards difference equation. Time integra-
tions are accomplished by a Runge-Kutta technique developed by Shampine and
Watts (1974). Calculations of chemical reaction rates are accomplished in
a separate subroutine as are calculations of microbial growth and degrada-
tion rate. Thus, other mechanisms for kinetic rates can be substituted into
the program with a minimum of effort.
Initial conditions are read in for each of 56 segments of the river reach.
Two different inlet boundary conditions can be applied. The first assumes no
dispersion upstream of the inlet reach. This boundary condition gives reason-
able results when large non-point source or point source loads are simulated.
The second set of boundary conditions assumes that the concentration at the
inlet is a known constant. This second assumption appears to give reasonable
results when small downstream inputs are simulated. No dispersion at the end
of the river reach is always assumed. Source strengths are assumed to be
constant in the current version of the program and are read in at the begin-
ning of the execution for each of the 56 segments.
104
-------
READ INPUT DATA
READ INITIAL AND
BOUNDARY CONDITIONS
READ SOURCE STRENGTHS
CALCULATE RATE OF
CHEMICAL DEGRADATION
CALCULATE RATE OF
MICROBIAL DEGRADATION
CALCULATE RATES
OF TRANSPORT
EXECUTE TIME STEP
INTEGRATION
PRINT CONCENTRATION
PROFILES
NO
STOP
END
T = T + AT
T = T + T.
PRINT PRINT 1
Figure 2. Flow diagram for malathion degradation program.
105
-------
To compare results obtained from AEcoS experiments with laboratory
results, the data had to be fit to the mathematical model. Assuming that
steady-state has been attained, equation 1 can be rewritten as
. +
IJ
To obtain a satisfactory least squares fit of the data, malathion concen-
tration data and bacteria counts were averaged over the steady-state period.
The averaged values were substituted into a finite difference approximation
to equation 11 to calculate a rate for bacterial growth and malathion de-
gradation at each of the seven interior sampling stations. Since only nine
data points were available, a second-order central difference was used to
approximate first and second spatial derivatives at the five most interior
points. A fourth-order forward difference approximation was used at the
second sampling location, and fourth-order backward difference approxima-
tion was used at the eighth sampling location.
The seven rates were then fit to either equation 2 or equation 7 by a
nonlinear least squares fit using a Marquardt-Levenberg iterative curve-
fitting algorithm (Knott, 1972).
RESULTS AND CONCLUSIONS
Typical simulation results are shown in Figures 3, 4, 5, and 6. Fig-
ure 3 is a plot of malathion concentration versus time of travel that would
result under flow conditions in which longitudinal dispersion is small and
degradation is attributable to alkaline hydrolysis. The effect of tempera-
ture variation is quite large. Figure 4 shows the effect of pH on malathion
concentration profiles. A change in pH has a dramatic effect on the rate of
hydrolysis. Figure 5 shows malathion concentration profiles as a function
of the glucose concentration at the inlet of the reach. All of these model
simulations were obtained using coefficients reported by Paris et al_. (1975)
and Wolfe et_ £l_. (in press).
By comparing the rates of malathion degradation by alkaline hydrolysis
and microbial action, combinations of environmental conditions can be de-
fined in which either hydrolysis or microbial degradation is the dominant
pathway of malathion breakdown. Figure 6 illustrates this comparison. At
high pH and low bacteria counts, alkaline hydrolysis is the major degrada-
tion pathway. At low pH and high bacteria counts, microbial degradation is
the major degradation pathway.
Typical results from alkaline degradation experiments (pH 8.25) con-
ducted in AEcoS are shown in Figure 7. Second-order rate coefficients for
malathion degradation calculated from least squares fit of AEcoS data are
compared in Table 2 with coefficients determined in the laboratory- Rate
coefficients calculated from AEcoS experiments are approximately 26% lower
than those obtained in the laboratory.
106
-------
0-0
0
4 8 12 16
TIME OF FLOW, days
Figure 3. Effect of variation in temperature on alkaline degra-
dation of malathion.
20
16-0
TEMP = 27°C
0-0.
-pH = 8-0
pH=7-4 pH = 7-2 pH =7-1
pH=7-C
0
10 20 30 40
TIME OF FLOW, days
50
Figure 4. Effect of variations in pH on alkaline degradation of
malathion.
107
-------
0-095 mg/l GLUCOSE
0-47 mg/l GLUCOSE
0-95 mg/l GLUCOSE
0-0
TIME OF FLOW, days
Figure 5. Effect of variations in utilizable carbon loads on
microbial degradation of malathion.
TEMP - 27°C
pH
8
7
6-
5-
X
/
/
/
/
CHEMICAL DEGRADATION ^—»
MORE THAN 75% OF '
TOTAL x
/ / /
/
X
/
CHEMICAL DEGRADATION/
MORE THAN 50% /
OF TOTAL
X
/_
V
/
/
/
/
/
/ BACTERIAL DEGRA-
/ DATION MORE THAN
A 50% OF TOTAL
' I
/ BACTERIAL DEGRADATION
/ MORE THAN 75% OF TOTAL
10 10 10°
BACTERIA CONG
10 10
, bacteria/ml
10
10'
Figure 6. Comparison of microbial and alkaline hydrolysis
degradation pathways.
108
-------
0-6
pH 8'25
TEMP = 27°C
0-0
I
_L
0'5 1-0 1-5 2-0
TIME OF FLOW, days
Figure 7. Typical steady-state malathion concentration profile
observed in AEcoS during alkaline degradation study.
TABLE 2. COMPARISON OF RATE COEFFICIENTS FOR ALKALINE
DEGRADATION BETWEEN LABORATORY AND AEcoS STUDIES
pH k
7.5
8.25
;i (M«sec~
3.
4.
i
)
86
1
2
(AEcoS Study)
± 0.
± 0.
11
47
ki (M-sec
5
5
-1)
.4 ±
.4 ±
(Laboratory
0.
0.
1
1
Study)
Typical results for microbial degradation are shown in Figures 8 and 9
The decline of bacteria concentration down the length of the channel indi-
cates that the rate of utilization of malathion as an energy source is too
slow to fulfill the metabolic requirements of the organisms. Consequently,
to describe the dynamic behavior of the microbial population, a death term
was added to the model
where
= -k-CB
k = specific death rate
109
(12)
-------
4-0
pH<6-8
TEMP = 27°C
TIME OF FLOW, days
Figure 8. Typical steady-state bacteria concentration profile
observed in AEcoS during microbial degradation study.
pH<6-8
TEMP = 27°C
BACTERIA CONG =
3-0 x I0 cell/ml
23
TIME OF FLOW, days
Figure 9. Typical steady-state malathion concentration profile
observed in AEcoS during microbial degradation study.
110
-------
Values obtained for specific death rates by a least squares fit of bacteria
data from each experiment are shown in Table 3.
TABLE 3. SPECIFIC DEATH RATE FOR BACTERIA CULTURES USED
IN AEcoS MICROBIAL DEGRADATION EXPERIMENTS
Temperature
k (min"1) x 10"1*
22°C
27°C
32°C
0.98
1.86
0.10
0.72
Glucose fed into the channel inlet was effectively utilized at the point
of injection. Consequently, equation 12 described the total change in bac-
teria concentration down the length of the channel. The only apparent effect
of variation in glucose input was in regulating the size of the bacterial
.population at the channel inlet. Second-order rate coefficients for microbial
degradation of malathion calculated from least squares fits of AEcoS data
are compared in Table 4 with coefficients evaluated in the laboratory. The
values calculated from AEcoS experiments are again lower than values ob-
tained in laboratory studies. Reproducibility of results was fair for the
one test replicated at 32°C.
TABLE 4. COMPARISON OF RATE COEFFICIENTS FOR MICROBIAL
DEGRADATION BETWEEN AEcoS AND LABORATORY STUDIES
K2 (AEcoS Study) k2 (Laboratory Study)
Temperature (1 org^hr'1) x 10~12 (1 org^hr"1) x 10"12
22°C
27°C
32°C
0.59 ± 0.31
1.3 ± 0.52
0.77 ± 0.46
1.38 ± 0.58
4.9 ± 2.1, 2.5a
Coefficient calculated from Monod constants reported by Paris et aj_. (1975)
and malathion concentration of 0.8 mg/1.
Flask experiments and background tests conducted in AEcoS, indicated
that bacteria entering the system from chamber air and inflgwing water sup-
ply made no significant contribution to malathion degradation. Their con-
tribution to the total bacterial population was also small (less than 10%
of the total population).
Ill
-------
The difference in degradation rates between chamber experiments and
laboratory experiments cannot be attributed to changes in non-limiting
nutrient concentrations in the case of microbial degradation; in flask ex-
periments varying the amounts of non-limiting nutrients produced no detect-
able effect on the rate of degradation. The difference in rate of degrada-
tion in the case of hydrolysis reaction could be due to differences in the
buffers used.
We suggest an alternative explanation. In laboratory studies reported,
solutions were well-mixed, i.e., systems were characterized by high turbu-
lence levels. In AEcoS experiments, the turbulence levels were low. The
low level of turbulence introduces the possibility of mass transport limi-
tations to the degradation of malathion by bacteria. Hydrolysis of malathion
may be similarly affected, although it is less likely because of the molecular
nature of the reaction. Further experimentation would be required to deter-
mine the source of the differences in these experiments.
A few further tests must be done to gather additional background data
for the microbial system under study. The next major system scheduled for
testing includes an algal component. In this series of experiments, a green
and a blue-green algae will be added to the bacterial system and its effects
on malathion degradation rate will be studied.
The results of the experiments completed in the AEcoS were similar
to those obtained in laboratory studies. However, the rates of the two
processes studied in the AEcoS were significantly lower than those ob-
tained in laboratory studies. For valid extrapolation of laboratory re-
sults to field situations, phenomena causing this reduced rate should be
studied. If it is due to mass transport limitations, the effect will be
even more important for fast processes, e.g., phosphorus cycling.
112
-------
REFERENCES
American Public Health Association, Inc. Standard Method for the Examin-
ation of Water and Wastewater 1965. 1965.
U.S. Environmental Protection Agency. Initial Scientific and Minieconomic
Review of Malathion. EPA Report 540-1-75-005. 1975.
Koivistoinen, P. and H. Aalto. Malathion Residues and Their Fate in Cereals.
Nuclear Techniques for Studying Pesticide Residue Problems. Proceedings
of a panel, Vienna, December 16-20, 1968. International Atomic Energy
Agency, Vienna. 1970.
Knott, G.D. and D.X. Reece. MLAB: A Civilized Curve-Fitting System.
Proceedings of the Online "72 International Conference, Brunei Uni-
versity, England. September, 1972.
Kollig, H.P. An Automated Chemical Analytical System for an Aquatic Eco-
system Simulator. Environmental Research Laboratory, Athens, Georgia.
In preparation.
Paris, D.F., D.L. Lewis, and N.L. Wolfe. Rate of Degradation of Malathion
by Bacteria Isolated From an Aquatic System. Environmental Science
and Technology. 9(2): 135-138, 1975.
Salvador!, M.G. and M.L. Baron. Numerical Methods in Engineering.
Prentice-Hall, Inc., Englewood Cliffs, New Jersey, p. 65-95. 1961.
Sanders, W.M. and J.W. Falco. Ecosystem Simulation for Water Pollution
Research. In: Advances in Water Pollution Research. Pergamon Press,
New York. p. 243-253. 1973.
Shampine, L.F. and H.A. Watts. Global Error Estimation for Ordinary Dif-
ferential Equations. Sandia Laboratories Report SLA-74-0198. Sandia
Laboratories, Albuquerque, New Mexico. 1974.
Stumm-Zol linger, E. and R.H. Harris. In: Organic Compounds in Aquatic
Environments. Faust, S.J. and J.V. Hunter (eds.). Marcel Dekker, Inc.,
New York. Chapter 23.
Wolfe, N.L., R.G. Zepp, J.A. Gordon, and G.L. Baughman. The Kinetics of
Chemical Degradation of Malathion in Water. Environmental Research
Laboratory, Athens, Georgia. In press.
113
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Shagawa Lake Recovery Characteristics
As Depicted by Predictive Modeling
D.P. Larsen and H.T. Mercier*
ABSTRACT
Predictions obtained using several mass balance models
describing changes expected in lake phosphorus concen-
trations resulting from an external phosphorus supply
reduction to Shagawa Lake were compared with observa-
tions. Two of the models predicted a rapid recovery of
the lake and underestimated present wintertime phos-
phorus concentrations by about 50%. A third model which
includes an algal biomass component projected similar
wintertime total phosphorus concentrations but showed
how internal sources of phosphorus can delay the attain-
ment of this level. Two of these models were used to
project lake phosphorus concentrations expected if
wastewater phosphorus concentrations were allowed to
increase from the present 50 yg/1 to 400 yg/1 and 1.0
mg/1. Both suggest that at effluent concentrations of
1.0 mg/1, the lake would exhibit phosphorus concentra-
tions often associated with a eutrophic state.
INTRODUCTION
The use of mathematical models as tools to assist in understanding
the dynamics of aquatic ecosystems as well as to predict their responses
to man induced perturbations has increased dramatically in recent years.
A partial listing of aquatic ecosystem model's includes those developed
and used by 85 investigators responding to a survey inquiry conducted
during 1974 (Parker and Roop, 1974); many others exist. Models which
describe lake trophic state and algal dynamics (reflecting trophic state)
have been developed at several levels of complexity. Vollenweider (1969,
1975) has advocated single compartment phosphorus mass balance .models on
the premise that lake phosphorus concentrations provide an estimate of
trophic state and algal concentrations. Others have expanded and elabo-
rated this type of model with good success.
Slightly more complex phosphorus lake models have dealt with a two
compartment (particulate and dissolved), vertically stratified (epilim-
inion and hypolimnion) system (Snodgrass and O'Melia, 1975; Imboden,
*Corvallis Environmental Research Laboratory, U.S. Environmental Protection
Agency, Corvallis, OR 97330
114
-------
1973, 1974). These models are more effective when considering seasonal
changes in phosphorus levels and might be more effective when predicting
average phosphorus concentrations (usually taken as values observed at
vernal circulation). More complex models describe the interrelationships
between various components of aquatic ecosystems including nutrients,
algae, herbivores, carnivores, and decomposers (Chen, 1970; Park, et
al., 1974; Thomann, et al., 1975; Baca, et al., 1974). These models may
have three spatial dimensions, but are often one or two dimensional.
At Shagawa Lake, Minnesota, an opportunity exists to test the
predictive capabilities of representatives of these models of several
levels of complexity in describing the lake's response to a large-scale,
man induced, environmental perturbation: the phosphorus supply to
Shagawa Lake was reduced to about 20% of its former level by removing
essentially all of the wastewater phosphorus which could enter the lake.
This report compares results of predictions using models developed by
Vollenweider (1969, 1975), Snodgrass and O'Melia (1975), and a simplified
epilimnetic algal model similar to those developed by Thomann, et al.
(1975), and Baca, et al. (1974) with observations in the lake. In
addition, since it is unlikely that the wastewater phosphorus removal
efficiency will continue at its present level because the operation is
expensive, projections using higher wastewater phosphorus concentrations,
up to 1.0 mg/1 (the Minnesota State Standard) are included.
Shagawa Lake, a shallow (mean depth 5.7m) lake located in north-
eastern Minnesota, has received wastewater from the city of Ely since
about 1880 when the development of mining and logging industries attracted
hundreds of settlers. As a result the lake became eutrophic. A tertiary
wastewater treatment plant designed to reduce effluent phosphorus concentra^
tions to 50 yg P/l, (a 99% reduction) became operational in early 1973.
Since wastewater accounted for approximately 80% of the total supply of
phosphorus to Shagawa Lake from surface sources, (the remainder origi-
nating primarily from natural sources), its removal should cause a
dramatic change in lake conditions. Detailed background of the project
and documentation of nutrient loads and limnological characteristics of
Shagawa Lake can be found in Larsen and Malueg (.1975), Larsen, et al.
(1975), Malueg, et al. (1973), Malueg, et al. (1975), and Schults,
Malueg, and Smith (1975).
VOLLENWEIDER MODEL
Based upon earlier work (Biffi, 1963; Piontelli and Tonolli, 1964)
Vollenweider (1969, 1975) developed a mass balance model for total
phosphorus in lakes to include external supplies, loss through the
outflow and sedimentation. He chose to describe sedimentation as a
function of the amount of phosphorus in the lake, proposing the following
equation:
115
-------
' *p - <>w + V
where [P] = total phosphorus concentration in the lake (M L )
si = volumnar phosphorus supply (ML T )
p = hydraulic washout coefficient (T )
w _1
a = sedimentation rate constant (T )
t = time (T).
Assuming constant £ , p , and a , a time dependent solution to equation
(1) can be obtained analytically as
[P(t)] - [Pje -(pw + V* f -^- (1 - e -(pw + V*) (2)
u Mw p
Assumptions are: a well-mixed lake, constant lake volume, outflow
concentration equivalent to lake concentration, equivalent inflow and
outflow rates, and no net supply from the sediments. An important point
is that this model is essentially an accountability statement, i.e.,
material in the lake occurs as a balance between supplies and losses.
The only hypothesis contained in equation (1) is that sedimentation is a
function of the amount of phosphorus in the lake (Vollenweider, 1975).
Models of this general nature have been described often (Dillon, 1974).
The solution, equation (2), to equation (1) depends on an experiment-
ally difficult to determine sedimentation rate coefficient, a . Dillon
and Rigler (1974) and Sonzogni, et al. (1975) have proposed alternative
means to obtain its value. These methods were used to estimate a ; the
results are summarized in Table 1. Other variables and coefficients can
be determined experimentally.
Equation (1) was solved using flow and phosphorus loading data for
1973 and 1974 (see Malueg, et al., (1975) for flow and loading calculation
methods). For projections beyond 1974, projected wastewater flows and
phosphorus concentrations were added to average natural flow and phospho-
rus concentrations (based upon data obtained during 1972 and 1973).
Model response was compared with wintertime total phosphorus values.
Springtime concentrations are usually used for comparisons because a
lake is likely to be well mixed at this time, but in Shagawa Lake,
concentrations change rapidly shortly after ice-out. For example,
during the three weeks subsequent to ice-out in 1973, mean total phospho-
rus concentrations declined from 63 yg/1 to 44 yg/1. During the interval
from mid-December to mid-January each year, mean concentrations changed
116
-------
only slightly, and hence, were taken as better representatives of mean
conditions. These mean values were determined as follows. Each week a
volume-weighted average lake concentration was calculated from vertical
profiles (1.5 m depth interval) located at three stations in the lake.
These weekly values were then averaged over the interval from mid-
December to mid-January.
Figure 1 compares the expected response of Shagawa Lake to reduced
phosphorus input using this model with the wintertime total phosphorus
values. One run displays the expected lake response treating total
phosphorus as a conservative substance (a =0; termed the hydraulic
washout model); the second run includes the deposition term (a =0.852
-1 "
yr ; termed the phosphorus washout model). Both runs suggest that
Shagawa Lake should respond rapidly to the reduced phosphorus supply,
attaining a stable state within two years. The expected time required
to reach 95% of a steady state value can be used as a measure of the
responsiveness of lakes to changed inputs. This is usually given as
three times the phosphorus retention time (T =l/(p +a )) or 1.3 yr for
p w p
Shagawa Lake using the values summarized in Table 1. Hence, this model
suggests that a new stable state should have occurred by mid-1974. The
phosphorus washout model suggests that mean lake concentrations should
be 10-12 yg/1 when a steady state is attained if phosphorus deposition
is similar to that observed during pretreatment years. The hydraulic
washout model suggests a mean lake concentration of 17-20 yg/1, equiva-
lent to the mean influent concentration. Sufficient time has elapsed
for the lake to have achieved a new stable state. Figure 1 suggests that
the lake responded rapidly; however, mean concentrations have not declined
to expected levels. In fact, mean concentrations are approximately
twice as high as those predicted from the phosphorus washout model and
are higher than those suggested by the hydraulic washout model alone.
There is an indication that the lake may have achieved a stable state,
since late 1974 winter time total phosphorus concentrations are similar
to those during late winter 1975 (Figure 1). This lake concentration
is higher than mean influent concentrations, hence must be maintained
by an internal supply of phosphorus.
Although the lake has not responded entirely as predicted by these
washout models, it is instructive to speculate on what can be expected
if the wastewater phosphorus concentration is altered from its present
level of 50 yg/1 to a higher level. The steady state solution to
equation (1) can be written (Larsen and Mercier, 1975) as
[P] = [p] (1 - Rp) (3)
where [p] = average influent phosphorus concentration (M L" ) and
a
R = phosphorus retention coefficient = f—-
P pw p
(Vollenweider, 1975).
117
-------
60
| SOjjO
„ 40
Z)
cr
o
x
CL
c/)
O
X
CL
30
20
O 10
0
°°
0
0
o
o
o
°
o
• 00
oooo o
oooo ooooooo
• •
* *
1973
1974
YEAR
1975
Figure 1. Comparison of response of Vollenweider mass balance model
with lake observations. Solid circles represent model run
with phosphorus deposition; open circles represent no depo-
sition. Lake observations are mean wintertime values (see
text) j^ 1 standard deviation.
118
-------
TABLE 1. ESTIMATES OF PW AND a FOR SHAGAWA LAKE
pw gp
-1 -1
) (yr )
1971 1.85 0.67a
0.88b
1972 1.26 0.82a
0.93b
a) determined from R = Pp_ = 1 - Annual Export ± Lake Change
P p + a Annual P Import
(Dillon and Rigler, 1974).
b) determined from steady state solution of mass balance model
(Equation 1). (Sonzogni, et al., 1975).
TABLE 2. POTENTIAL EFFECT OF DIFFERENT WASTEWATER EFFLUENT
PHOSPHORUS CONCENTRATIONS ON MEAN LAKE CONCENTRATIONS
Wastewater
Effluent Cone.
yg/1
50
200
400
700
1000
% of
Natural Supply
6
22
45
79
113
Mean Influent
Cone.
yg/1
18
20
24
30
35
Expected Mean
Lake Cone.
yg/1
12
Meso-
13
trophic
16
20
Eutrophi<
23
119
-------
This relationship expresses the concept/that the steady state lake
concentration will be equivalent to the mean influent phosphorus concen-
tration in the absence of deposition of phosphorus (i.e. if phosphorus
were a conservative substance). The effect of phosphorus deposition is
to reduce the mean influent concentration and this effect can be expressed
as the phosphorus retention coefficient, that fraction of incoming
phosphorus which is retained by the sediments.
Equation C3) was used to project the effect of alternative waste-
water effluent total phosphorus concentrations upon mean lake concentra-
tions. The results, using total phosphorus effluent concentrations up
to 1.0 mg/1, are summarized in Table 2. This analysis suggests that
wastewater effluent phosphorus concentrations of 700 yg/1 or greater
might produce a eutrophic lake while the present effluent concentrations
of 50 pg/1 might produce a lake of lower mesotrophic classification. A
400 yg/1 effluent concentration, equivalent to about a 90% secondary
wastewater phosphorus reduction, might produce a mid-mesotrophic lake.
It is interesting to note that an effluent concentration of 1.0 mg/1
would provide a supply of phosphorus approximately equivalent to the
supply from all other sources. The classification of lakes into trophic
categories is difficult; here the guidelines suggested by Vollenweider
(1968) and Dillon (1975) have been adopted.
These projections provide an estimate of the state toward which the
lake might stabilize and are based upon the assumption that the sediments
will act as a net sink for phosphorus as they did prior to 1973. If
this basic model is correct the recent wintertime total phosphorus
concentrations might typify lake characteristics at an effluent of 1.0
mg/1 after stable conditions have been attained.
SNODGRASS - O'MELIA MODEL
Snodgrass and O'Melia (1975) proposed a more complex mass balance
model which includes particulate and ortho-phosphorus, divides the lake
into epi- and hypolimnia, each well mixed, and incorporates two seasons,
each 180 days. One represents summer conditions during which stratifica-
tion occurs and the other hypothesizes a well mixed lake. Lake processes
include: conversion of ortho-phosphorus into particulate phosphorus,
sedimentation of particulate phosphorus, decomposition of particulate
phosphorus into ortho-phosphorus, and vertical exchange of material
across the epilimnion-hypolimnion boundary. An effect of flocculation
was developed such that the net sinking velocity in deep lakes was
greater than that in shallow lakes. Although they specifically state
that the model is applicable to lakes whose hypolimnia remain aerobic
throughout the year, it was instructive to apply the model to Shagawa
Lake, in which anaerobic conditions have regularly developed each year
during late winter (before ice-out) and during late summer.
120
-------
The equations and model coefficients are summarized in Table 3.
Snodgrass and O'Melia estimated coefficients from the literature and
calibrated the model using Lake Ontario data. The coefficients they
presented were used for the Shagawa Lake runs except those that were
site specific (e.g., mean depth, volume, etc.). External water and
phosphorus supplies used were those observed for Shagawa Lake.
Predicted year end total phosphorus concentrations were compared
with lake observations obtained during mid-December to mid-January as
before (Figure 2). This interval was also selected to minimize the
effects of sediment phosphorus supply during anaerobic periods thereby
potentially minimizing the model constraint of aerobic conditions through-
out the year. For the years prior to treatment (1968-1972), the model
results are quite close to observed values with the exception of the
single value for 1967-1968 when the model projects total phosphorus
concentrations of about 46 vg/1 and the lake mean during the only week
for which data were obtained was 28 yg/1.
It is particularly encouraging that the lake observations and model
response for the last three pretreatment years (1970-1971, 1971-1972,
1972-1973) are in close agreement when sampling frequency had been
increased. This agreement might occur because the phosphorus pulses
which occur in the lake during anaerobic intervals deposit rapidly
subsequent to circulation periods, hence, their effect on wintertime
averages is minimized. However, model projections subsequent to treat-
ment are similar to those predicted by the Vollenweider model, suggesting
total phosphorus concentrations near 10 yg/1, substantially below observed
values.
EPILIMNION MODEL
A three compartment model was constructed to describe the seasonal
changes in total phosphorus and algal phosphorus within the well mixed
epilimnion (5.25 m) of Shagawa Lake and to predict recovery of the lake
subsequent to treatment. The model is a simplified version of those
developed by Thomann, et al. (1975) and Baca, et al. (1974). The following
is a general description of the model structure; equations and coefficient
values are summarized in Table 4.
The specific rate of growth of algae was related to solar radiation,
temperature, and soluble reactive phosphate; loss rates included the
effects of sinking, conversion into non-algal particulate phosphorus
(lumping the effects of zooplankton grazing and cell death) and washout.
A specific growth rate reduction factor, as a function of total daily
radiation, was generated by averaging Vollenweider's (1964) expression,
relating photosynthesis to light intensity, over a 24 hour day and the
mixed zone as elaborated by Fee (1973). This integral can be used to
evaluate relative photosynthesis over the euphotic zone for particular
values of physiological "constants" and various amounts of total daily
121
-------
TABLE 3. EQUATIONS AND COEFFICIENTS FOR SNODGRASS-0'MELIA MODEL (1975) AS APPLIED TO SHAGAWA LAKE
Summer Epilimnion ortho- and participate phosphorus
d[OPL k.. k..
- Q[OP] - pV[OP]
e eee
zt
d[PPl
QEPPJe + PeVe[OP]e - geAth[PPJe
Hypolimnion ortho- and particulate phosphorus
th th
d[PP]h k,. k.
- 9hAs[PP]h - rhVh[PP]h + J* Ath[PP]e - -
Hh £t
Winter Ortho- and particulate phosphorus
V -= zQjE'OPlj - Q[OP] - PeuVeuEOPj + rV[PP]
V = zQjlPP]j - QEPP3 + PeuVeuEOP:i " rVlPPj ' g
(Continued)
-------
TABLE 3. EQUATIONS AND COEFFICIENTS FOR SNODGRASS-O'MELIA
MODEL (1975) AS APPLIED TO SHAGAWA LAKE (continued)
Summer Stratification Winter Circulation
ge = 0.1 rag/day 9 = gjl + f(7 - 2^)3
9h a 900 + fZ~h} go = 0.05 ro/day
9Q * 0.05 m/day f = 0.05/m
f = 0.05/m I , = 10 ID for I > 10 m
eu —
kth = k = 0.005 I 7 = I for I < 10 m
I~
Sh
peu = 0.06/day
pe = 2.0/day r = 0.03/day
Coefficients Specific to Shagawa Lake
Vg = 39 x 106 m3 As = 5.5 km2
Ath = 5.5 km2 V = 53 x 106 m3
Vh = 14 x 106 m3 Veu = 39 x 106 m3
Ih = 2.54 m Z = 5.25 m
List of Symbols
2
A Surface area of the sediment water interface (L )
2
Ath Horizontal cross-sectional area of a lake at the thermocllne (L )
f Flocculation coefficient (L~ )
g Sedimentation coefficient of entire lake (L/T)
gh Sedimentation coefficient in the hypolimnion (L/T)
ge Sedimentation coefficient in the epilimnion CL/T)
grt Sedimentation coefficient in the absence of flocculation CL/T)
o o
kth Vertical transport coefficient in the thermocllne (L /T)
p Production rate coefficient in the epilimnion (T~ )
123
-------
Table 3 (continued), List of Symbols
p Production rate coefficient in the euphotic zone, circulation
model (T~ )
2
Q Volumetric rate of discharge from a lake (L /T)
3
Q. Volumetric rate of inflow to a lake from source j (L /T)
0 1
r Decomposition rate coefficient for entire lake (T )
r. Decomposition rate coefficient for the hypolimnion (T~ )
t time
o
V Volume of entire lake (L )
q
V Volume of the epilimnion (L )
3
V Volume of the euphotic zone in circulation model (L )
3
Volume of the hypolimnion (L )
1 Mean lake depth (L)
Z"eu Mean depth of the euphotic zone, circulation model (L)
Z^ Mean depth of the hypolimnion (L)
Zth Mean depth of thermocline region (L)
[OP] Concentration of orthophosphate (M/L )
«3
[PP] Concentration of particulate phosphorus (M/L )
124
-------
60
50
CD
40
T3
CD
CD
(/>
-Q
O
30
10
1972-73 \
(7)
1971- 72 (7)
1968-69
1974-75 (7)
1970-71
(6)
T/
I
1973-74
(6)
I969-7O
(2)
1968
(I)
0
10
20
30
40
50
60
P (Predicted) , pg/liter
Figure 2,
Comparison of predicted total phosphorus concentrations using
Snodgrass-O'Melia model with lake wintertime concentrations.
Solid line is 1:1 correspondence between predictions and obser-
vations. Dates mean wintertime values +_ 1 standard deviation
and numbers of observations making up the mean are indicated
for lake observations. See text for calculation methods.
125
-------
TABLE 4. EQUATIONS AND COEFFICIENTS FOR SHAGAWA LAKE EPILIMNION
PHOSPHORUS MODEL
Algae:
•CN- + a + P) [A]
Soluble Reactive Phosphorus:
Ke
Ke = 301A] +0.75
Kp = yg SRP/1
N1 = 0.05 day"1
N2 = 0.005 day"1
CTI = 0.01 day"1
a2 = 0.007 day"1
V = 40 x 106 M3
+N2[PP] - PW£SRP]
Non-Algal Particulate Phosphorus:
d£PPj = PPIN + NJA] - (N? + ay + PUI)EPP]
~ar~ "i ' ^ * w
Model Coefficients
=0.59 (1.066T)
CLITE = 0-172 (1 - e-°-Q07TDR) + O.QQQ451TDR
126
(Continued)
-------
TABLE 4. EQUATIONS AND COEFFICIENTS FOR SHAGAWA LAKE EPILMNION
PHOSPHORUS MODEL — LIST OF SYMBOLS
IA]
CLITE
[PP]
PPIN
w
[SRPJ
SRPIN
T
t
TDR
V
Concentration of algal phosphorus (M/L )
Fractional reduction in G^CO in epilimnion due to
availability of light
Maximum specific growth rate as a function of
temperature (T )
Extinction coefficient [L }
Concentration of SRP at which specific growth rate is
o
reduced to 1/2 maximum (M/L )
Conversion rate constant from algal phosphorus into
particulate phosphorus (T )
Conversion rate constant from non-algal particulate
phosphorus into soluble reactive phosphorus (T~ )
o
Concentration of non-algal particulate phosphorus (M/L )
Supply of non-algal particulate phosphorus to epilimnion (M/T)
Hydraulic washout coefficient as a function of time (T )
Settling rate constant for algal phosphorus (corresponding
to a settling velocity of 0.05 m/day) (T~])
Settling rate constant for non-algal particulate phosphorus
(corresponding to a settling velocity of 0.04 m/day) (T~ )
Concentration of soluble reactive phosphate (M/L )
Supply of SRP to epilimnion (M/T)
Temperature as a function of time (°C)
Time (T)
2
Total daily radiation as a function of time (gcal/L /T)
Epilimnion volume (L )
127
-------
radiation distributed realistically throughout a daylight day. Values
used for physiological "constants" were those reported by Fee (1973).
The expression developed for the reduction factor is given in Table 3
and is similar in form to that presented earlier (Larsen, Mercier, and
Malueg, 1973). Eppley's (1972) relationship between maximum specific
growth rate and temperature was used. The hyperbolic expression commonly
used to express the relationship between the rate of uptake of a nutrient
and its concentration in the extracellular medium was used to express
the fractional reduction in specific growth rate related to nutrient
concentration. First order rate kinetics were used to express the loss
of algal phosphorus to non-algal particulate phosphorus and through
sinking out of the epilimnion.
External sources of soluble reactive phosphate were wastewater,
tributaries, and precipitation. An internal supply was added to this,
mimicking the sediment and hypolimnetic supply (this supply is discussed
subsequently). Particulate phosphorus was converted to soluble reactive
phosphorus using first order kinetics. Soluble reactive phosphorus
losses were algal consumption and surface outflow. Particulate phosphorus
originated from tributaries, wastewater and conversion from algae.
Losses were sinking, conversion into soluble reactive phosphorus and
outflow.
An average water year was constructed from flow data (from all
sources) obtained during 1972 and 1973 to provide daily water input for
model runs. Weekly soluble reactive and total phosphorus loads deter-
mined during 1972 and 1973 for all natural sources (all sources excluding
wastewater) were averaged to produce weekly natural loads. Thus the
flow and loading inputs to the model displayed the same cycles each
year, representing average "natural" conditions. During 1972 observed
wastewater loads were added to the natural loads; from 1973 onward,
wastewater loads were calculated as the product of an assumed concentra-
tion in the effluent and the average wastewater flow.
The model was calibrated by manipulating the coefficients N-,, N2,
a-p and cr2 to fit the 1972-1973 average epilimnetic concentrations
observed in the lake. Model coefficients were initially estimated from
those presented in Thomann, et al. (1975) and Baca, et al. (1974) and
references therein.
An initial estimate of the internal supply of phosphorus was avail-
able from mass balance calculations comparing supplies and losses of
total phosphorus with lake changes for the years 1970-1973. This
estimate was necessarily a net supply and thus provided a lower limit to
the supply of phosphorus from internal sources. It was reasoned that
the lake sediments act like a capacitor, accumulating phosphorus during
the year and releasing it during the anaerobic intervals. For the
purposes of this model it was assumed that the total internal supply for
the year was proportional to the previous year's deposition. The supply
128
-------
was evenly distributed during two weeks in late winter and ten weeks
during the summer, intervals during which anaerobic conditions have
developed in the lake. For post-treatment years ft was reasoned that
the sediment phosphorus would washout, hence a constant of proportion-
ality greater than 1 was used. A value of 1.1 with an initial supply of
60 kg/day provided a good fit.
Figures 3 and 4 compare calibrated model runs with lake obser-
vations for total phosphorus and chlorophyll a_. Algal phosphorus was
converted to chlorophyll a^ using a factor chlorophyll a, = algal^phosphorus^
based upon regressions of chlorophyll a_ on particulate phosphorus for
lake observations. This conversion ratio is lower than the more commonly
used 1:1 value (Thomann, et a!., 1975; Baca, et al., 1974). The runs
presented are not necessarily the best fit to the data in any statis-
tical sense, but represent runs which approach the actual pattern
observed. Other combinations of coefficients could be used to produce
approximately the same results. The selection of coefficients used to
produce these "final" runs is therefore somewhat arbitrary, and although
techniques exist which provide parameter estimates based upon minimized
deviation techniques, they are somewhat costly and thus have not yet
been tried.
The model mimics the temporal pattern of total phosphorus and
chlorophyll ^although some differences occur both in timing and magnitude
of pulses (Figures 3 and 4). These differences probably occur because
year to year differences in timing of internal phosphorus pulses, initia-
tion of algal blooms and other factors were not included in the model.
The model is based upon average conditions and produces results which
might be expected in an average year. The gradual decline in phosphorus
concentrations in the lake is mimicked by the model up through mid-1975
as is a decline in the magnitude of the summer algal blooms. It is
interesting that the model projects wintertime phosphorus levels of
about 20 yg/1 during 1974-1975, and appears to stabilize at wintertime
epilimnion levels similar to those projected by the previous models for
the entire lake (Figures 3, 5 and 6), but the time required to reach
this level is considerably longer than that projected by the other
models. Mean values for the epilimnion during the mid-December mid-
January are only slightly lower than the average lake levels used to
compare the previous models.
The calibrated model projections of total phosphorus and algal
concentrations expected if wastewater concentrations were altered are
summarized in Figures 5 and 6. The model was run until a stable pattern
occurred; the results displayed compared the effects of 50, 400, and
1,000 yg/1 total phosphorus wastewater concentrations; (total phosphorus
was supplied in a ratio of 30% soluble reactive phosphorus and 70%
particulate phosphorus, consistent with observations based upon propor-
tions in the present wastewater). Data obtained during 1974 are included
for comparison. The model projections suggest that the total phosphorus
129
-------
100
CO
o
CD
Figure 3. Comparison of predicted epilimnetic total phosphorus concentrations (light line)
with lake observations.
-------
CO
7O
60-
C 50-
40
^ 30
i
O
o:
q 20
10
o|
0
1972
1974
YEAR
1975
1976
Figure 4. Comparison of predicted epilimnetic chlorophyll concentrations (light line)
with lake observations.
-------
0
Figure 5.
50
TIME , weeks
Sensitivity of epilimnion model response (total phosphorus)
to different wastewater total phosphorus concentrations
(50, 400, 1000 yg/1) at model stable state (light lines)
compared with 1974 lake epilimnion observations.
132
-------
o
Figure 6,
50
TIME , weeks
Sensitivity of epilimnion model response (chlorophyll a_)
to different wastewater total phosphorus concentrations
(50, 400, 1000 yg/1) at model stable state (light lines)
compared with 1974 lake epilimnion observations.
133
-------
pattern which might occur if wastewater effluent total phosphorus
concentrations were 1.0 mg/1 would be similar to that observed during
1974 with peak average values of about 60 yg/1 and wintertime concentra-
tions of about 20-25 yg/1. Chlorophyll a_ concentrations might similarly
exist as they did during 1974 with peak average summertime concentrations
of 25-30 yg/1. A wastewater effluent concentration of 400 -pg/1 might
produce conditions where total phosphorus concentrations slightly exceed
40 yg/1 during the summer and wintertime concentrations would be approxi-
mately 15 yg/1. Summertime chlorophyll a_ values might be about 20 yg/1.
These projections rely upon the assumption that the lake will respond in
a fashion similar to the manner in which it has responded in the past.
Of particular significance are the projected phosphorus pulses which
drive the summer blooms. These seem to be related to the generation of
anaerobic conditions in the bottom waters thus, if, at present wastewater
effluent levels, the deeper waters cease becoming anaerobic, a different
pattern might emerge. It is also unknown whether higher wastewater
phosphorus concentrations (for example 400 yg/1) will continue to
promote anaerobic conditions, hence provide stimulation for summer algal
blooms.
DISCUSSION
The Vollenweider and Snodgrass-O'Melia models predict lake phosphorus
concentrations considerably below observed values because these models
do not include mechanisms by which phosphorus is supplied from internal
sources, a feature which apparently controls the recovery of the lake to
some extent. Vollenweider (1975) specifically stated that the mass
balance model he developed is inapplicable to situations in which there
is net annual internal supply of phosphorus; Snodgrass and O'Melia
restrict the utility of their model to lakes whose hypolimnia remain
aerobic throughout the year. Nevertheless, it was instructive to apply
these models to compare their predictions with lake observations to
obtain a measure of the deviation of the Shagawa Lake response from
simple model predictions. Both models suggest that wintertime total
phosphorus concentrations should be about 10 yg/1. This value is a
reasonable expectation since mean influent concentrations are presently
about 18 yg/1, and lake concentrations are expected to be less than this
value after a stable state has been achieved. An empirical expression
which relates the retention of phosphorus by lakes to the hydraulic
washout coefficient (Larsen and Mercier, 1975) suggests that 45% of
Shagawa's influent phosphorus should be sedimented annually; therefore,
steady state concentrations should be approximately 10 yg/1 as suggested
by the above two models. Also, total phosphorus concentrations in the
upper 10m of Burntside Lake, upstream of Shagawa Lake and uninfluenced
by urban activities, were approximately 9 yg/1 during 1974. Thus these
models likely provide good indications of the expected level at which
the lake should stabilize.
134
-------
The effect of the internal source of phosphorus has been to delay
the attainment of these predicted levels. This effect has been incorpor-
ated in the epilimnion model by establishing the construct that the
internal supply of phosphorus is proportional to the previous years
deposition. A proportionality greater than 1 implies leeching from the
sediments, i.e. a store which built over pretreatment years discharges
for sometime subsequent. A value of 1 or less might also represent
sediment leeching since it is likely that only a fraction of deposited
phosphorus is converted into soluble form.
The observed wintertime phosphorus concentrations might be the
result of an equilibrium between sediment phosphorus and lake phosphorus
maintained above expected levels by high concentrations of phosphorus
within the sediments. This store of sedimentary phosphorus might be
depleted slowly, depending upon such characteristics as the size of the
reservoir (both as available phosphorus within a unit volume of sediments,
and the depth within the sediments to which this phosphorus can release
into overlying water), hydrodynamic characteristics of the overlying
water, mixing processes within the sediments themselves, or the extent
to which the sediment-water interface continues to become anaerobic. If
there is indeed a temporary or permanent equilibrium, the epilimnion
model does not include mechanisms to describe it or its effect on lake
dynamics, as indicated by the fact that it predicts wintertime total
phosphorus concentrations similar to those of the other models but the
time taken to reach those levels is longer than simple models suggest.
The epilimnion model does show how, by pulsed inputs of phosphorus from
internal sources, average levels of phosphorus during wintertime can be
higher than those predicted by other models and how pulsed phosphorus
inputs during the summertime influence summertime algal biomass as i.s
expected during phosphorus limited conditions.
There is a precautionary note which must be considered when evalu-
ating models of complex systems which have been tuned by coefficient
manipulation, particularly if the models are to be used for predictive
purposes. The models are abstractions of processes thought to be impor-
tant in controlling the response of ecosystem; however, the model
calibration or tuning process can hide failure to include some of the
important processes because model coefficients are often difficult to
verify in natural systems. The experimental values of coefficients used
often display such ranges that wide latitude is available in selecting
values during calibration to provide an acceptable fit to observed data.
Thus the determination of the validity of projections is a somewhat
subjective process and predictive ability might be poor if important
processes are not included.
Although models of the nature of those presented in this paper are
simple representations of complex systems and often include only estimates
of important rates or coefficients, they provide an indication of some
of the characteristics which can be expected in a lake. Perhaps more
important, they provide a framework against which processes within the
lake can be identified, particularly as deviations from model results.
This can assist in designing or revising experimental approaches to
provide a more complete description of a lake's activity or response.
135
-------
REFERENCES
Baca, R. G., M. W. Lorenzen, R. D. Mudd, and L. V. Kimmel. 1974.
A generalized water quality model for eutrophic lakes and reservoirs.
A report prepared for Office of Research and Monitoring, U.S. EPA,
Washington, DC by Pacific Northwest Laboratories, Battelle, Richland,
Washington.
Biffi, F. 1963. Determining the time factor as a characteristic
trait in the self-purifying power of Laga d'Orta in relation to
continual pollution. Atti 1st. Ven. Sci. Lett. Arti. 121:31-136.
Chen, C. W. 1979. Concepts and utilities o'f ecologic model. J. Sanit.
Eng. Div., Am. Soc. Civil Eng. 96:1085-97.
Dillon, P. J. 1974. A critical review of Vollenweider's nutrient
budget model and other related models. Water Resources Bull.
10:969-989.
Dillon, P. J. 1975. The phosphorus budget of Cameron Lake, Ontario:
The importance of flushing to the degree of eutrophy of lakes.
Limnol. Oceanogr. 20:28-39.
Dillon, P. J., and F. H. Rig!er. 1974. A, test of a simple nutrient
budget model predicting the phosphorus concentration in lake water.
J. Fish. Res. Bd. Can. 31:1771-1778.
Eppley, R. W. 1972. Temperature and phytoplankton growth in the
sea. Fishery Bull. 70:1063-1085.
Fee, E. J. 1973. A numerical model for determining integral primary
production and its application to Lake Michigan. J. Fish. Res.
Board Can. 30:1447-1468.
Imboden, D. M. 1973. Limnologische Transport- und Nahrstoff-modelle.
Schweiz. Z. Hydrol. 35:29-68.
Imboden, D. M. 1974. Phosphorus model for lake eutrophication.
Limnol. Oceanogr. 19:297-304.
Larsen, D. P., and K. W. Malueg. 1975. Limnology of Shagawa Lake,
Minnesota, prior to reduction of phosphorus loading. Hydrobiologia.
(In Press).
Larsen, D. P., K. W. Malueg, D. W. Schults, and R. M. Brice. 1975.
Response of eutrophic Shagawa Lake, Minnesota, U.S.A. to point-
source phosphorus reduction. Verh. Int. Ver. Limnol. 19:884-892.
Larsen, D. P., and H. T. Mercier. 1975. Lake phosphorus loading graphs:
an alternative. Submitted to J. Fish. Res. Board Can.
Larsen, D. P., H. T. Mercier, and K. W. Malueg. 1973. Modeling algal
growth dynamics in Shagawa Lake, Minnesota with comments concerning
projected restoration of the lake. In: Modeling the Eutrophication
Process (E. J. Middlebrooks, D. H. Falkenborg, and T. E. Maloney,
eds.). Proceedings of a Workshop held at Utah State University,
Logan, September 5-7, 1973. Published by the Utah Water Research
Laboratory, Utah State University, Logan, Utah. p. 15-31.
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Malueg, K. W., R. M. Brice, D. W. Schults, and D. P. Larsen. 1973. The
Shagawa Lake Project: Lake restoration by nutrient removal from
wastewater effluent. Washington, D. C., Government Printing Office.
Ecological Research Series, EPA-R3-73-026. 49 p.
Malueg, K. W., D. P. Larsen, D. W. Schults, and H. T. Mercier. 1975. A
six-year water, phosphorus and nitrogen budget for Shagawa Lake,
Minnesota. J. Environ. Quality. 4:236-242.
Park, R. A., et al., 1974. A generalized model for simulating lake
ecosystems. Simulation. 23:33-50.
Parker, R. A., and D. Roop. 1974. Survey of aquatic ecosystem models.
The Institute of Ecology Publication. 131 p.
Piontelli, R., and V. Tonolli. 1964. The time of retention of lacustrine
waters in relation to the phenomena of enrichment in introduced
substances, with particular reference to the Lago Maggiore. Mem.
1st. Ital. Idrobiol. 17:247-266.
Schults, D. W., K. W. Malueg, and P. D. Smith. 1975. Limnological
comparison of culturally eutrophic Shagawa Lake and adjacent oligo-
trophic Burntside Lake, Minnesota. American Midland Naturalist.
(In Press).
Snodgrass, W. J., and C. R. O'Melia. 1975. A predictive model for
phosphorus in lakes. Environmental Science and Technology.
9:937-945.
Sonzogni, W. C., P. D. Uttormark, and G. F. Lee. 1975. The phosphorus
residence time model: theory and application. Submitted to Water
Research.
Thomann, R. V., D. M. DiToro, R. P. Winfield, and D. J. O'Connor. 1975.
Mathematical modeling of phytoplankton in Lake Ontario. 1. Model
development and verification. EPA-660/3-75-005. Ecological
Research Series. Corvallis, Oregon, p. 177.
Vollenweider, R. A. 1965. Calculation models of photosynthesis-depth
curves and .some implications regarding day rate estimates in primary
production estimates. Mem. 1st. Ital. Idrobiol. Suppl. 18:425-
457.
Vollenweider, R. A. 1968. Scientific fundamentals of the eutrophigation
of flowing waters, with particular reference to nitrogen and
phosphorus as factors in eutrophication. OECD Technical Report
DAS/CSI/68.27. Paris, France. 159 p.
Vollenweider, R. A. 1969. Moglichkeiten und Grenzen elementarer Modelle
der Stoffbilanz von Seen. Arch. Hydrobiol. 66:1-36.
Vollenweider, R. A. 1975. Input-output models with special reference
to the phosphorus loading concept in limnology. Schweiz. Z. Hydrol.
37:53-83.
137
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A Mathematical Model of Pollutant Cause
And Effect in Saginaw Bay, Lake Huron
William L. Richardson- and
Victor J. Bierman, Jr.*
INTRODUCTION
As part of the joint U.S.-Canadian Upper Great Lakes Study, the U.S.
Environmental Protection Agency, Large Lakes Research Station at Grosse
lie, has undertaken an intensive evaluation of the water quality process
in Saginaw Bay. The project includes field examination of water quality
and development of cause and effect models for data interpretation. These
models are designed to simulate the effect of nutrients on the growth,
composition, and distribution of phytoplankton biomass and will eventually
be used to simulate the effect of nutrient control alternatives.
The primary emphasis of this paper is the presentation of methodology,
including the practical considerations of applying an existing model struc-
ture to a new physical system. The existing model is the phytoplankton
chlorophyll-nutrient model developed by O'Connor, et al. (1973). This ap-
proach has evolved over the past decade largely through research support
from the U.S. Environmental Protection Agency and has recently culminated
in its application to Lake Ontario (Thomann, et al. 1975). The computer
program, which implements the Lake Ontario model (LAKE-1), has been modi-
fied to represent the physical system of Saginaw Bay. First, model output
using the Lake Ontario biological parameters will be shown. This output
will then be compared to model output using a set of biological parameters
determined for Saginaw Bay. These parameters were based on a series of
numerical experiments. The results of this comparison are preliminary and
do not necessarily represent the exact dynamics or kinetics of the bay.
SAGINAW BAY WATER QUALITY CHARACTERISTICS
Although small in comparison to Lake Huron, Saginaw Bay is an impor-
tant water resource serving as a source of water supply for municipal and
industrial uses, for sport and commercial fishing, recreation, waste dis-
posal, and navigation. The Bay has a surface area of about 2500 square
kilometers and a 21,000 square kilometer drainage basin (Figure 1). The
basin supports a population of 1.2 million (1970 census) and a variety of
land uses including large industrial and urban centers. The basin also
contains extensive agricultural, recreational, and natural areas.
*Environmental Research Laboratory--Duluth, U.S. Environmental Protection
Agency, Large Lakes Research Station, Grosse lie, MI 48138
138
-------
Figure 1. Saginaw Bay Basin and five model segments
139
-------
The hydrodynamics of the bay are such that Ayers et al. (1956) con-
cluded that the bay acts like an estuary. Fish and Wildlife Service drift
bottle studies concluded that circulation is variable and closely related
to meteorlogical conditions. Northeast winds for example, can drive water
into the bay which results in water level fluctuations near Saginaw River
of over one meter (USDI 1956).
The physical, chemical, and biological dynamics in Saginaw Bay are
very complex. Because of the dynamic interaction with Lake Huron, large
water quality gradients extend from the Saginaw River, the primary source
of material input, to the outer extremities of the Bay. Chloride levels
in 1974 varied from 39 mg/1 near the Saginaw River to 2 mg/1 near Lake
Huron. Chlorophyll a. concentrations (an indirect indication of phytoplank-
ton biomass) were recorded in the range of 1 to 82 mg/1 in 1974 (Smith,
1975).
GENERAL MODELING PROCESS
The water quality processes for large, complex natural systems can be
considered to consist of two primary components, 1) physical transport and
2) biological-chemical processes. A common approach is to develop separate
process models independently and merge results into a final water quality
management model (Figure 2). The resolution of either model component de-
pends on the purpose or application of the model. For a general indication
of future water quality trends for planning purposes, simple biological and
chemical models averaged over space and time are appropriate. If more
resolution is desired to answer other specific questions, then the model
components must be sub-divided to provide more precise simulations.
Complex ecosystem models have been structured by several theorists but
these have remained primarily research models with little or no field cali-
bration or verification (Middlebrooks et al., 1973). Factors limiting the
complexity of ecosystem models include computer size and execution time, as
well as data acquisition and analysis. Expansion of the simple modeling
framework to include more physical, chemical, and biological resolution is
continuing as part of the EPA Great Lakes research program. Bierman (1975)
has structured a four class phytoplankton model which contains more detailed
nutrient-phytoplankton interaction kinetics. Thomann, et al. (1975), has
structured a 67 segment model for Lake Ontario. These models represent the
next generation of verified ecological models.
In the case of Saginaw Bay, the primary question concerns the effects
of material load reduction on water quality parameters such as dissolved
solids and biomass. The modeling approach applied herein is a simple eval-
uation tool for making such assessments. The space scale considered is on
the order of 10-30 kilometers and the time scale on the order of seasons.
MODEL PRINCIPLES
The key principle used in the modeling of material transport is mass
balance. The model equations for physical, chemical, and biological sys-
tems conserve mass in both space and time. The computer program which im-
plements these equations accounts for and traces materials from their spatial
140
-------
GROSSE ILE LAB MODELING PROCESS
Nutrients
Sunlight
Temperature
Geometry
Chloride
Wind
Geometry
Temperature
Chemical—Biological Process
Chlorophyll a
Chemical—Biological Process
Multi-Species Biomass
Course Grid
Chloride Transport
Model
I
Fine Grid
Hydrodynamic
Process Model
Water Quality
Management Models
Further Research
Refinements
Figure 2. Mathematical modeling process diagram.
-------
points of input to their final export points. If the material is conser-
vative, the transport mechanisms are totally physical. If the material is
non-conservative, such as nitrogen and phosphorus, the transport mechanisms
also include chemical and biological reactions. A mass balance equation in
finite difference form for a water body divided into n finite completely
mixed segments is given by Thomann et al. (1975):
[V] d(s)k = [A] (s.) ± (S) ± (W) (1)
~dT~ K K K
where
v = n x n diagonal matrix of volumes, (L)
(s,,) = n x 1 vector of material concentration
[A] = n x n matrix of advective and dispersive transport terms
(S,) = n x 1 vector of kinetic interaction terms
(W)k = n x 1 vector of material sk inputs
This system of equations accounts for the mass of a substance k in each
model segment which is equal to the mass entering minus the mass leaving,
plus or minus mass produced or lost within the segment.
PHYSICAL TRANSPORT
The first phase of model development for large, dynamic water systems
is usually devoted to quantifying the circulation. This has been done for
Saginaw Bay by tracing a conservative substance, chloride, through the sys-
tem by adjusting transport parameters in equation 1 until a reasonable com-
parison is obtained between computed and measured chloride concentrations
(Richardson 1975).
BIOLOGICAL-CHEMICAL PROCESSES
The general scheme of nutrient/chlorophyll a dynamics for a single
segment has been adapted from O'Connor, et al. (T973) (Figure 3). The spe-
cific system scheme used for Saginaw Bay adapted from Thomann et al. (1975)
includes eight state variables with their interactions (Figure 4). This
model is a simplification of a complex biological-chemical system where
phytoplankton biomass is represented by chlorophyll a_ which is used primar-
ily because of the ease of measurement and availability of data. Phyto-
plankton carbon is specified using carbon-chlorophyll stoichiometry obtained
by experimental data and is the element which zooplankton consume along with
other nutrients contained in the phytoplankton. The nutrients, phosphorus
and nitrogen, are also accounted for and traced through the phytoplankton and
zooplankton by specifying stoichiometry relationships with carbon. The model
assumes all other nutrients to be in sufficient supply so as not to limit
phytoplankton growth.
142
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Material Sinks
(m/T)
Physical Transport
SS
N
0 PHYTOPLANKTON
CONCENTRATION
SS
NUTRIENT
CONCENTRATION
Zooplankton
(m/L3)
f
Prey
I
\
Grazing
J_
Phytoplankton
(m/L3)
Nutrient ... ,,
Limitation Nutnent Use
Nutrients
(m/L3)
TIME
TEMPERA TURE
TEMPERATURE
SOLAR RADIATION
Material Inputs
(m/T)
TIME
Figure 3. General nutrient-phytoplankton model interactions.
••••
Carnivorous
<
Zooplankton
PNC
Herbivorous Zooplankton
TPNC
Phytoplankton
1
1
N
N
J P
Organic
N itrogen
t"
Orga
Phospt
P 1
1
^ Ammonia
Nitrogen
N
^
nic
lorus
L
Available
Phosphoru;
1
Nitrate
Nitrogen
t-
Figure 4. Specific model biological-chemical interaction
diagram (Thomann 1975).
143
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The biological-chemical equations have been developed and reported
by Thomann et al. (1975) and the details will not be repeated here. How-
ever, in summary the key assumptions of the kinetic expressions will include:
1. Phytoplankton growth rate is a function of temperature, light and
nutrients and uses Michaelis-Menton product kinetics. A maximum
growth rate is computed from the temperature relationship and then
reduced by the product of the nutrient and light limitation terms.
The death rate is a function of zooplankton grazing and endogeneous
respiration. Both of these processes are temperature dependent.
The dead phytoplankton return to the non-living organic nutrient
pools and the grazed phytoplankton become part of the zooplankton
biomass. Phytoplankton can also leave the system by sinking to the
sediment.
2. Herbivorus zooplankton growth rate is a function of the grazing
efficiency, grazing rate, and available phytoplankton, and follows
a Michaelis-Menton form (i.e., the growth rate reaches an asympote
as phytoplankton concentration increases). The death rate is a
function of temperature and the grazing rate of the carnivorous
zooplankton. Grazed zooplankton biomass and nutrients become part
of the carnivorous zooplankton biomass whereas the dead zooplankton
biomass and nutrients return to the non-living organic pools.
3. The growth rate for carnivorous zooplankton is a function of grazing
rate and efficiency and the death rate is a function of temperature
only. Dead zooplankton return to the appropriate material pools.
APPLICATION TO SAGINAW BAY
SEGMENTATION
The bay was divided into five (5) segments shown in Figure 1. This seg-
mentation scheme was chosen after considering such factors as water quality
gradients, morphology, spacial resolution desired, and available research
time.
MATERIAL LOADINGS
The primary source of material input to Saginaw Bay is the Saginaw Ri-
ver. The .loadings provide the forcing functions, (W), to the mathematical
model (Equation 1) and must be defined for each state variable during the
entire period for which the model is run. For this investigation loadings
were computed by the product of river discharge and material concentration.
The U.S. Department of the Interior (1975) provides daily discharge in-
formation for the four major tributaries to the Saginaw River including the
Cass, Tittabawassee, Shiawassee, and Flint Rivers. Meaningful flow measure-
ments can not be made near the mouth of the Saginaw River since the river
behaves like an estuary and reacts hydraulically to the fluctuations in the
bay water levels. This can cause flow stagnation and reversals which make
stage recordings meaningless for flow computations. Therefore, the daily
measured flows from the four major tributaries were added along with a com-
puted flow representing the ungaged portion of the basin (about 24% of the
total basin).
144
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The material concentrations were obtained from the Michigan Water
Resources Commission bimonthly sampling station at Midland Street in Bay
City about five miles upstream from the mouth of the river, the Cranbrook
Institute of Science (1975) sampling stations at the Dow Chemical Co.
water intake (samples about every two or three days) about a mile upstream
from the mouth and from the Cranbrook Station at Midland Street (sampled
once or twice per month). During low flow periods, June through December,
the Dow intake is influenced by the diluting effect of the bay, therefore
these data were used only for the peiod of January through June. For days
when a sample was collected a daily load was computed as the product of
the daily average flow and the grab sample concentration. These points
were connected by straight line segments to provide a continuous time series
for the entire year of 1974. These are shown in Figure 5 for each state
variable including chloride.
Circulation --
The circulation pattern in the bay was obtained by mathematically
tracing the transport of chloride from Saginaw River through the five
segments. The details of this approach have been presented by Richardson
(1975). In summary, having measured the chloride loads from the Saginaw
River and the average segment concentrations, the transport dispersion
and the advection terms in the mass balance equation are adjusted (Figure
6) until computed chloride concentrations match the measured in the five
segments (Figure 7). The measured chloride time series is the cruise by
cruise average of chloride measurements. A simulation is acceptable when
the computed concentration falls within a range of plus and minus one stan-
dard deviation of the mean. Note that this criterion has not yet been met
at all times in each segment. This is especially the case in segments 1
and 3, the smallest and most dynamic. However, these comparisons are suf-
ficiently accurate during this initial phase. The degree of further refine-
ment will depend on the results of biological-chemical modeling and the
resolution obtained.
Initial Results --
As a starting point, an initial simulation was obtained from the modi-
fied Lake Ontario "LAKE-1" computer program. Only those parameters unique
to Saginaw Bay in 1974 were changed. These included segmentation, advective
and dispersive transport terms, boundary conditions, initial conditions,
segment temperatures, segment light extinction, and segment depth. The Sag-
inaw Bay physical characteristics are listed in Table 1. The first run was
made using the verified Lake Ontario biological-chemical parameters (Thomann
et al. 1975). To simplify this presentation only the results for Segment 3
are shown (Figure 8).
To compare the computed concentrations (model output) with those measured
(actual data), the data for all sampling stations in a model segment (Figure
9) were combined for each sampling cruise and the mean and standard devia-
tion computed. This was easily facilitated by the use of the EPA data system,
STORET. The cruise means and standard deviations were inputted to the LAKE-1
graphics subroutine and plotted along with the computed results for each of
comparison.
145
-------
Sag in aw River
MATERIAL LOADS TO SAGINAW BAY
£ . 250
|s 200
f sl150
egS'lOO
o
1974 ORGANIC PHOSPHORUS
LOAD
40 120 200 280 360
1974 ORTHOPHOS-
PHATE LOAD
40 120 200 280 360
40
120 200 280 360
1974 NITRATE LOAD
S 4
5 i
40 120 200 280 360
DAYS
1974 CHLORIDE LOAD
40
120 200 240 360
DAYS
Figure 5. Saginaw River material loads to Saginaw Bay.
MODEL CALIBRATION PROCESS
INFLUENCE FROM
LAKE HURON
SEGMENTATION
Depth D
Area A
Volume V
Length L
BAY CIRCULATION
Advection Q
Dispersion £
BAY CHLORIDE CONCEN-
TRATIONS CM
CHLORIDE LOADINGS Wk
INPUT
A v\D\ L
COMPUTE CHLORIDE
CONCENTRATION
C,
Figure 6. Physical transport model calibration process.
146
-------
25
-h-I-
J'F'M A'M'J'J'A'S
100
50
j 'F'M'A'M'JV/VSO N D
25
J'F'M'AVJ'J^'S'O'N'D
1974SAGINAWBAYCHLORIDE'
I Cruise average ± SD (mg/l
«•— Computed (mg/l)
25
-i-TW-M—•
j T'M'A'M'J'J'A'S oVD
50
j'F'M'AMJ JASON D
Figure 7. Comparison of 1974 computed and measured chloride concentrations in
five model segments.
-------
80
60
40
20
0
J FMAMJ J ASON.D
z-0.6
O en
5E 0.4
Ign2
Q_ OQ U.£
MO 0
£ 0.4
C£
O
IT.
Q_
tn —
it 0.2
Q- fc
I
o
g o
C 2.0
en
g i.o
LU
ct:
i o
-------
TABLE 1. SAGINAW BAY MODEL PHYSICAL PARAMETERS
Segment Interacting
Segment
1
2
3
2
3
4
3
5
4
5
Lake Huron
5
Lake Huron
Segment
Volume
3
894
5890
1270
7880
9390
Average
Depth
m
3.85
7.33
3.74
13.22
15.15
Light Ext.
Coef.
m-1
1.5
1.0
1.0
.5
.5
Intersegment
Area
m2 x 103
123
35
137
133
35
499
323
769
SAGINAW BAY
1974 Sampling Network
Legend o Boat Station
A Water Intake
—I h Model Segments
10 Ml
Saginaw River
Figure 9. Saginaw Bay 1974 sampling network.
149
-------
The initial results (Figure 8) reveal the following:
1. Computed phytoplankton grow too early in the year which results in
bad timing for the spring and fall peaks as compared to the data.
2. As a result of the early phytoplankton growth, computed zooplankton
appear too soon which shifts the entire zooplankton time series.
Too much computed zooplankton biomass could be one of the factors
which reduces the computed phytoplankton levels relative to the
observed values.
3. The reaction rates for conversion of organic nitrogen and ammonia
to nitrate are too low. This is apparent because the computed con-
centrations of total kjeldahl nitrogen and ammonia are higher than
the measured and that for nitrate is too low.
4. The poor timing and insufficient growth of phytoplankton result in
poor timing for orthophosphorus. Computed orthophosphorus appears
to be another limiting factor for the computed phytoplankton peak
during the summer.
5. Total phosphorus appears to be too low in the fall. This could be
caused by incorrect loss rates and/or by an inaccurate total phos-
phorus load.
Calibration --
The calibration process proceeds similarly to that described previously
for chloride except there are 16 biological-chemical parameters to adjust
compared to only only two transport parameters. This process is not a curve
fitting exercise in a statistical sense; rather, it requires an understanding
of the cause and effect relationships inherent in the system and knowledge
of a reasonable range of values for each parameter (O'Connor, et al. 1973).
As each parameter is altered keeping all others constant, perception of the
sensitivity and importance of each is obtained by the analyst which further
increases his insight and intuition. One problem arising, however, is the
practicality of assimilating all of the information generated. Computer
output must be reduced to graphs, and graphs of simulations overlaid to
depict the alterations. The analyst soon becomes overwhelmed having to
perceive eight state variables, for five segment, for numerous calibration
simulations.
To simplify this process the emphasis was given to chlorophyll a^ in
segment 3. This reduces-the number of graphs from forty to just one for
each simulation. A few of the more important initial sensitivity simula-
tions are shown in Figure 10 compared to the base run. The results of
these sensitivity runs revealed the following:
1. No single alteration of any of the principal parameters appears
to have a significant effect on the magnitude of the computed
phytoplankton chlorophyll jj.
2. A reduction of 75% in the phosphorus-chlorophyll a^ ratio (from
.001 to .00025 mg per yg chlorophyll a) has the most significant
single impact primarily in summer chlorophyll peak.
150
-------
80
cn
CL.
O
O
O
o
60
40
20
0
30 60 90 120 150 180 210 240 270 300 330
JULIAN DAY
Figure 10. 1974 phytoplankton chlorophyll ^concentration sensitivity simulations including:
a —^—Base simulation
b —.—.—.—Phosphorus-chlorophyll a^ ratio of .00025
c • • « 'Carbon-chlorophyll ratio of .025 and herbivorous zooplankton grazing rate of .2
d Organic nitrogen and ammonia nitrogen decomposition rates of .01
e x-x-x-xAll above except phos-chlor ratio of .0005 and phytoplankton settling rate of .05
-------
3. Only after altering the phosphorus-chlorophyll a_ ratio, carbon-
chlorophyll a_ ratio, herbivorous zooplankton grazing rate, and
nitrogen decomposition rates concurrently does the chlorophyll
ia increase significantly.
4. To fit the data, the organic nitrogen to ammonia, and ammonia to
nitrate decomposition rates must be increased by an order of mag-
nitude over those used for Lake Ontario.
Once these required alterations were determined, additional sensitivity
simulations were made from this new base. The best comparisons (at the time
of this report) of observed versus computed for the eight state variables
and five segments are shown in Figures 11 through 17. As these show, more
effort remains to obtain an acceptable calibration. The parameter values
for this final run are listed in Table 2. In particular, as Figure 13 shows,
computed levels of orthophosphorus are too high in all segments throughout
the year. Also, it would be desirable to refine the chlorophyll ^simula-
tion in Segment 3.
TABLE 2. BIOLOGICAL-CHEMICAL MODEL PARAMETERS
Lake -j
Ontario
Saginaw
Bay
Units
Nitrogen
Phosphorus
Zooplankton
Phytoplankton
Carbon
Half-saturation constant 0.025 0.025
Organic nitrogen decomposition rate 0.0175 0.007
Ammonia to nitrate nitrification rate 0.002 0.015
Nitrogen-chlorophyll ratio 0.01 0.01
Half-saturation constant
Organic phosphorus decomposition rate
Phosphorus-chlorophyll ratio
Conversion efficiency
Endogenous respiration rate
Herbivorous zooplankton grazing rate
Carnivorous zooplankton grazing rate
Chlorophyll half-saturation constant
Endogenous respiration rate (at 20°)
Settling velocity
Saturated growth rate
Carbon-chlorophyll ratio
0.002
0.007
0.001
day",deg"
day" deg"
a
0.005
0.007
0.00025
day" deg
-1
0.6
0.001
0.06
0.06
0.6
0.001
0.04
0.04
day"
1/mg
1/mg
deg"1
C day deg
C day deg
10
0.1
0.1
0.58
20
0.1
0.05
0.50
vg/i.
day"1
m/day
day"1
0.05 0.025
Ratio mg element to yg chlorophyll.
'Thomann 1975.
152
-------
Figures 11-17. Final comparison of 1974 computed and measured material
concentrations.
50
25
O|JIF'MIA'M'J'J'AIS'O'N'D'\ "VF'M'A'M'JVA'S'O'N'D
25
JF'M AM'j'J'A'S'O'N'D
100
50
J'F'MVM'J'j'A'S'O'N' D
1974 SAG INAW BAY PHYTOPLANKTON
CHLOROPHYLL;
J Cruise average! SD lug/I)
—• Computed (ug/l)
100
50
J'F'M'A'M'J'J'A'S'O'N'D
Figure 11
0.5
j F M A M J J'A'S'O'N'D
1.0
0.5
J'F'M'A'M'J J A s'o N D
0.5
0.5
J'F'M'A'M'J'J'A'S'O'N'D
1974 SAG INAW BAY ZOOPLANKTON
CARBON
J'F'M'A'M'J'J'A'S'O'N'D
I Cruise average + SD (mg/l)
— Computed (mg/l)
i.o
0.5
J'F'M'A M'J'J'A s o N D
Figure 12
153
-------
0.025
j 'F'M'A'M'J'J'A'S'O'N'D
0.025
0.025
0.05 -
0.025 -|
'N1 D
1974 SAGINAW BAY ORTHO-PHOSPHORUS
0.05
u/ftffir>
j'F'M'AWj'j'A'S'O'N'D1
0.025
J'F'M'A'M'J'J'A'S'O'N' D
JCruise average ± SO (mg/l)
- Computed (mg/l)
Figure 13
Ii
I II IT
j 'F'M'A'M'J' J'A'S'O'N'D
J F MA MJ JASON D
J F M A M J J A'S'O'N D
j F M A'M j j A'S o N'D
1974 SAGINAW BAY NITRATE NITROGEN
2.0
1.0
ICruise average + SD (mg/l)
— Computed (mg/l) Q
Figure 14
154
J F'M AM J J A S 0 N D
-------
0.25
J'F'M'A'M'J'J'A'S'O'N'D
0.50
0.25
J JASON D
0.25
J F'M A'M J J A'S O'N D
0.25
n T . . li*** , TA. .
J'F'M'A'M'J'J'A'S'O'N'D
1974 SAGINAW BAY AMMONIA NITROGEN
0.50
0.25
J Cruise average± SO (mg/l)
-^ Computed (mg/l)
Figure 15
•fi"
r?M A M J J A S 0 N D
1.0
I I
J'F'M'A'M'J'J'A'S'O'N D
2.0
1.0
J'F'M'A'M j J A s o N D
1.0
J'F'M'A'M'J'J'A'S'O'N'D
1.0
J'F'M'A'M'J'J'A'S'O'N'D
^1974 SAGINAW BAY TOTAL KJELDAHL
NITROGEN
I Cruise average ± SD (mg/l)
— Computed (mg/l)
Figure 16
155
JVM A'M J J A'S 0 N'D
-------
0.05
J'F'M'A^TJ J A s o N D
0.05
0.05
0.1
0.05
j 'F'M A M j j A s o N D
1974 SAG INAW BAY TOTAL PHOSPHORUS
0.1
J F M A M J J A S 0 N D
I Cruise average ± SD (mg/l)
— Computed (mg/l) o
Figure 17
JT M A M J J A S 0 N D
JTVA M J J A S 0 N D
DISCUSSION AND CONCLUSIONS
Although additional effort remains to be expended, some general conclusions
can be drawn from the work done to date. It is emphasized that the model para-
meter values reported herein (Table 2) have yet to be confirmed and checked and
only represent order of magnitude estimates at this time.
It has been learned that substantial research resources are required to
apply an existing computer program/model to a new physical system. The emphasis
in modeling research is not on the writing of the computer program but rather on
arriving at the basic workable kinetic structure and then calibrating this struc-
ture to a given data set. If a calibration can be obtained, the model must then
be verified by comparing model results to an independent data set-ideally after
the system has been perturbed (i.e., phosphorus loads altered).
Although this effort was initially considered to be an "application" pro-
ject, calibration and verification for a new system require such a substantial
effort, with no guarantee of success, that this work should be considered as
applied research. Only after calibration and verification are obtained can a
model be "turned over" to an engineering staff for use in making management de-
cisions. The time required from project initiation to final transfer to manage-
156
-------
ment of any new system is on the order of one to two years if all required
data is available.
The calibration/verification research, itself, has side benefits which
are useful during the course of the work. The modeling process requires a
systematic approach to data collection and analysis. It helps structure a
surveillance program and preliminary model results can be used to gain in-
sight into the data and help in the management of surveillance and experi-
mentation programs. The preliminary results reveal gaps in our knowledge
for a particular system and are useful to direct new study and research.
As an example, the Saginaw Bay model has revealed possible errors in total
phosphorus loading in the fall. This has led to detailed inquiry of pos-
sible additional sources of phosphorus from diffuse sources, release from
sediments, and from the atmosphere. Inquiry was made to regulatory agencies
on possible seasonal loadings from waste sources. Additional work remains
to be done to quantify the effect of dredging as a possible source of phos-
phorus and more effort needs to be made on the method of sampling the Saginaw
River and accurately computing the loadings to the Bay.
Concerning model results, no specific conclusions can be drawn as yet.
However, the model has revealed Saginaw Bay to be quite a different system
than Lake Ontario which reinforces intuitive conclusions made from observing
the data. Decomposition rates used in the model are higher as expected.
The phosphorus-chlorophyll a_ ratio apparently must be lower, especially in
the fall. This indicates a possible need to alter the model structure. Per-
haps this ratio is time-variable due to a shift in the dominant plankton
forms. Phytoplankton settling velocity appears to be lower than that for
Lake Ontario. This is reasonable because the hydrodynamics of the bay tend
to keep materials in suspension. The initial phytoplankton growth was better
using a saturated growth rate of .5 per day rather than .58 used for Ontario
and the Potomac models (O'Connor, et al. 1973). This may be an artificial
compensation for lack of information on exact levels of solar radiation and
light extinction in the water under ice. Alternatively, it could be accounted
for by species differences between the systems. The zooplankton kinetics are
yet to be confirmed by experimental research so the zooplankton parameters
remain to be adjusted over large ranges.
In conclusion, this effort represents a first EPA in-house attempt to
apply an existing computer program/model which had been developed under
previous and ongoing EPA grants. A close working relationship was estab-
lished between the grantee staff and the EPA research staff which resulted
in better communications and an overall enhanced program for all of the
Great Lakes modeling research.
The average chlorophyll a^ biomass model is a relatively economical
research tool which can aid in analyzing complex, limnological interactions
and has been useful in guiding additional research and data gathering
activities.
ACKNOWLEDGEMENTS
The authors would like to express their gratitude to the many persons
who assisted with this project. In particular, we thank Nelson Thomas and
157
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Tudor Davies for their encouragement and support. The assistance provided
by Richard Winfield at Manhattan College in implementing the computer pro-
gram was invaluable. John Gannon from the University of Michigan Biological
Station, Pellston, Michigan, provided the zooplankton data and Elliott Smith,
Cranbrook Institute of Science, provided the chemical data. We also thank
the staff of the Large Lakes Research Station at Grosse He for reviewing
the manuscript and Amy Torongeau for typing numerous drafts.
REFERENCES
Ayers, J.C., D.V. Anderson, D.C. Chandler, G.H. Lauff. 1956. Currents
and Water Masses of Lake Huron, (1954) Synoptic Surveys. Ontario
Department of Lands and Forests, Division of Research and University
of Michigan, Great Lakes Research Institute.
Bierman, V.J., Jr. 1975. Mathematical Model of the Selective Enhancement
of Blue-Green Algae by Nutrient Enrichment. In press.
Middlebrooks, E.J., D.H. Falkenborg, and T.E. Maloney (EDS.) 1973.
Modeling the Eutrophication Process. Ann Arbor Science.
O'Connor, D.J., R.V. Thomann, and D.M. DiToro. 1973. Dynamic Water
Quality Forecasting and Management. Environmental Protection Agency
Ecological Research Series EPA-660/3-73-009.
Richardson, W.L. 1975. An evaluation of the transport characteristics
of Saginaw Bay using a mathematical model of chloride in Mathematical
Modeling of Biochemical Processes in Aquatic Ecosystems. Ann Arbor
Science Press (In press).
Smith, V. Elliott. 1975. Annual Report—Upper Lakes Reference Study:
A Survey of Chemical and Biological Factors in Saginaw Bay (Lake
Huron). Cranbrook Institute of Science, Bloomfield Hills, Michigan,
for the U.S. Environmental Protection Agency, Grosse He Laboratory.
Thomann, R.V., D.M. DiToro, R.P. Winfield, D.J. O'Connor. 1975. Mathe-
matical Modeling of Phytoplankton in Lake Ontario, 1. Model Develop-
ment and Verification. Environmental Protection Agency Ecological
Research Series EPA-660/3-75-005.
U.S. Department of the Interior, Fish and Wildlife Service. 1956. Surface
Current Studies of Saginaw Bay and Lake Huron.
U.S. Department of the Interior, Geological Survey. 1974. Water Resources
Data for Michigan, Part 1.
158
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Mathematical Model of Phytoplankton Growth and
Class Succession in Saginaw Bay, Lake Huron
Victor J. Bierman, Jr. and
William L. Richardson*
INTRODUCTION
Mathematical modeling techniques can provide a systematic basis for
a research approach to the problem of cultural eutrophication and can greatly
aid in the comparison of various management options. From an applied stand-
point, the general techniques of O'Connor et al. (1973) have been brought to
bear on a variety of different physical systems. In particular, Thomann et
al. (1975) and Canale et al. (1973) have investigated phytoplankton-nutrient
interactions in the Great Lakes. Chen and Orlob (1972; have also developed
such techniques and have used them to investigate, among other cases, the
effects of wastewater diversion from Lake Washington. From a research stand-
point, work is progressing on a number of systems models and component models
for purposes of gaining deeper insight with regard to chemical-biological
processes that occur in natural systems (e.g., Middlebrooks et al. [1973]).
The present work is part of the International Joint Commission's Upper
Lakes Reference Study involving Saginaw Bay, Lake Huron. The ultimate goal
of this work is to develop a mathematical model which can be used both to
describe the physical, chemical and biological processes that occur in Sagi-
naw Bay and to predict the effects of reduced waste loadings. Specifically,
the modeling effort will focus on phosphorus, nitrogen and silicon loadings
to the bay and the resultant production of phytoplankton biomass.
Model development is proceeding along two parallel pathways. The first
of these involves the development of research-oriented models which include
biological and chemical detail but which, for simplicity, do not include any
spatial detail. The second pathway involves the development of engineering-
oriented water quality model which mimics, as closely as practicable, the
actual physical system, including spatial detail. At any given point in time,
the water quality model will contain those chemical and biological processes
which have previously been investigated and developed using the spatially-
simplified model. There is constant feedback between the above two pathways
and constant interaction between the entire modeling effort and the ongoing
sampling effort on Saginaw Bay.
*Environmental Research Laboratory--Duluth, U.S. Environmental Protection
Agency, Large Lakes Research Station, Grosse lie, MI 48138
159
-------
Previous work (Bierman [1976]) involved a preliminary calibration of
a spatially-simplified, multi-class phytoplankton model to data from the
inner portion of Saginaw Bay. Only chlorophyll and dissolved nutrient data
were considered, and sensitivity analyses were presented for several impor-
tant processes affecting the development of blue-green algae.
The present paper involves the use of the same spatially-simplified
model. The purpose of this work is to obtain a more refined calibration
using a data set that has been expanded to include total phosphorus, total
nitrogen, and total zooplankton biomass. Also, finer time scales were used
for external nutrient loads, boundary conditions, and water circulation rates.
Ambiguities that can occur when calibrating an ecosystem model are discussed,
in particular, those that can occur by using chlorophyll concentration as an
indicator of phytoplankton biomass.
SUMMARY
A mathematical model of phytoplankton production has been applied to a
set of physical, chemical, and biological data from Saginaw Bay, Lake Huron.
The model includes four phytoplankton types, two zooplankton types, and three
nutrients: phosphorus, nitrogen, and silicon. The phytoplankton types in-
clude diatoms, greens, and both nitrogen-fixing and non-nitrogen-fixing blue-
greens.
The purpose of this study was to obtain the best possible calibration
between model output and the existing data set. This is one of the many
preliminary tasks which must be performed before such a model can ultimately
be used as a tool for making management decisions.
The model output agreed reasonably well with the data for phytoplankton
chlorophyll, total nitrogen, and dissolved forms of phosphorus, nitrogen,
and silicon. The model output did not agree well with the data for total
phosphorus during the latter part of the year, and the output showed a large
discrepancy with the total zooplankton data.
Ambiguities persisted in the interpretation of the model output because
insufficient data were available. The most serious problem was the lack of
simultaneous measurements of phytoplankton biomass and zooplankton biomass.
The existing phytoplankton data were available only in the form of chlorophyll
concentrations, a lumped parameter which can not be used to distinguish among
various functional groups of phytoplankton. Another problem was the lack of
direct measurements for all of the rate coefficients in the model. Given the
present state of the art, the latter problem is common to most ecosystems models
CONCLUSIONS AND RECOMMENDATIONS
Output from complex ecosystems models is difficult to interpret unless
large amounts of experimental data are available. Frequently, more than one
set of model coefficients will produce output which compares favorably with
experimental data. For this reason, it is difficult to gain insight with
regard to cause-effect mechanisms.
160
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As the number of state variables for which there is comprehensive data
is increased, many coefficients in a model become more tightly constrained.
If an ecosystem model can be calibrated to a large number of simultaneous
and independent variables, its reliability as a tool for drawing cause-
effect inferences can be greatly increased.
Continued development of sophisticated phytoplankton production models
will require detailed cell count and cell volume measurements. Chlorophyll
measurements alone give no information with regard to the partitioning of
phytoplankton biomass among various functional groups. In addition, system
atic, comparative studies of phytoplankton-nutrient dynamics are required
for various functional groups of phytoplankton in order to provide rate
coefficients for such models.
MODEL CONCEPTS
The basic model framework and preliminary simulations appear elsewhere
(Bierman [1976] and DePinto et al. [1976]). The compartments in the model
are four phytoplankton, two zooplankton, higher predators, and three nutri-
ents (Figure 1). The phytoplankton types include diatoms, greens, and blue-
greens, both nitrogen-fixing and non nitrogen-fixing.
1
1 I
i
! I
•4-1 |
I \
• L
,
HIGHER PREDATORS
L
ZOOPLANKTER
1
i t
ZOOPLANKTER
2
A V if A
DIATOMS
J
'
AVAILABLE
SILICON
i
s
NON-AVAILABLE
SILICON
A
i
i
i
i
GREEN ALGAE
1
1 1 1
1 1 1
BLUE-GREENS 1 BLUE-GREENS 1 ' ^,
(NONN-FIXING) \ (N-FIXING) 1 1
i i ]
1 -1
i _. i
i A
V T
AVAILABLE
PHOSPHORUS
AVAILABLE ATMOSPHERIC
NITROGEN NITROGEN
A *
NON-AVAILABLE
PHOSPHORUS
NON-AVAILABLE
NITROGEN
it4 i
Figure 1. Principal compartments of the Saginaw
Bay eutrophication model.
161
-------
The motivation for a multi-class modeling approach is that different
classes of algae have very different nutrient requirements, for example,
diatoms need silicon and certain types of blue-greens can fix atmospheric
nitrogen. In addition, at high concentrations, not all of these classes
have the same nuisance characteristics from a water quality standpoint.
Diatoms and green algae are grazed by zooplankton, but blue-green algae are
not significantly grazed and can form objectionable floating scums.
A unique feature of the model is that cell growth is considered to be
a two-step process involving separate nutrient uptake and cell synthesis
mechanisms. The motivation for this variable stoichiometry approach is that
an increasingly large body of experimental evidence indicates that the mech-
anisms of nutrient uptake and cell growth are quite distinct (e.g. Fuhs
[1969, 1971], Droop [1973], Caperon and Meyer [1972a, 1972b]). The model
includes carrier-mediated uptake of phosphorus and nitrogen using a reaction-
diffusion mechanism, and possible intermediate storage in excess of a cell's
immediate metabolic need's. Specific cell growth rates are assumed to be
dependent on the intracellular levels of these nutrients, in contrast to the
traditional Michaelis-Menten approach which relates growth rates directly to
extracellular nutrient concentrations.
MODEL IMPLEMENTATION
A major problem in attempting to implement a complex chemical-biological
process model is the lack of sufficient experimental data. It is often pos-
sible that more than one set of model coefficients could produce an acceptable
fit between the model output and a given data set. In the transition from
single-class to multi-class models, this problem becomes particularly acute
because it is no longer sufficient to ascertain a range of literature values
for a given coefficient. Multi-class models necessitate the definition of
class distinctions within this range. Given the present state of the art of
ecosystems modeling and associated experimental work, many of the coefficients
in such models must simply be estimated.
The primary operational differences among the phytoplankton types in the
model are summarized in Table 1. Principal phytoplankton coefficients are
summarized in Table 2. The working equations of the model and sensitivity
analyses of some of the most important coefficients have been presented in
Bierman (1976).
One of the assumptions of the model is that cell biomass concentration
is a more accurate indicator of standing crop than is chlorophyll concentra-
tion. In addition, chlorophyll is a lumped parameter and can not be used to
distinguish between different functional groups of phytoplankton. For these
reasons, chlorophyll concentration does not appear in any of the kinetic
equations of the model. However, the only available field data for phyto-
plankton in Saginaw Bay at-this time are chlorophyll concentrations. In
order to relate the model output to these data, the output must be converted
to chlorophyll concentration.
162
-------
TABLE 1. OPERATIONAL DIFFERENCES AMONG PHYTOPLANKTON TYPES
Characteristic
Property Diatoms
Phytoplankton Type
Blue-Greens Blue-Greens
(non N2-fixing) (Np-fixing)
Greens
Nutrient
Requirements
Relative Growth
Rates Under Opti-
mum Conditions at
20°C
Phosphorus Uptake
Phosphorus,
Nitrogen,
Silicon
High
Phosphorus, Phosphorus, Phosphorus
Nitrogen Nitrogen
Moderately
High Low Low
Affinity
Sinking Rate
Grazing Pressure
TABLE 2.
Low
High
High
PRINCIPAL
Low
High
High
PHYTOPLANKTON
High
Low
None
COEFFICIENTS
Phytoplankton
High
Low
None
Type
Parameter
Diatoms Greens
Blue-Greens Blue-Greens
(non Np-fixing) (Np-fixing)
Maximum P Uptake
Rate (day)-'
Maximum N Uptake
Rate (day)-T
Saturation Light In-
tensity (ft. candles)
Maximum Growth Rate
at 20°C (day)'1
Milligrams Dry
Weight per Cell 0.
Sinking Rate
(meters/day)
0.502
0.125
1000
2.0
15xlO'6
0.20
0.502
0.125
1000
1.9
0.27xlO"7
0.10
0.588
0.125
500
1.2
0.25xlO"7
0.05
0.588
0.125
500
1.2
0.41xlO"7
0.05
163
-------
The computer program which actually implements the model is written in
FORTRAN IV and is structured in a form such that any number of phytoplankton
and zooplankton types can be simulated, along with any set of food web_inter-
actions among these groups. The version of the model in Figure 1 consists of
20 simultaneous differential equations. The solutions were obtained using a
fourth-order Runge-Kutta method with a time step of 30 minutes for the nutri-
ent kinetics equations and a time step of 3 hours for the growth equations.
For a 365-day simulation, approximately 4 minutes of CPU time are required on
an IBM 370/158 computer. For the same simulation, approximately 45 minutes
of CPU time are required on the Grosse lie Laboratory's PDP-8/e minicomputer
with floating point hardware.
EXPERIMENTAL DATA
The chemistry and chlorophyll data used were collected by Cranbrook In-
stitute of Science (Smith [1975]). During 1974, 12 cruises were conducted
and samples were collected from 59 stations in Saginaw Bay. Samples were
taken at 1 meter and at all depths from 5 meters to the bottom in 5-meter
intervals. A total of 111 station-depth combinations were sampled on most
of the cruises. Analyses were conducted for 21 chemical parameters, includ-
ing phytoplankton chlorophyll. Since the present modeling study is re-
structed only to the inner portion of Saginaw Bay (Figure 2), only data from
the 33 field stations in this region were used.
Figure 2. Saginaw Bay watershed indicating distinctions
between inner and outer portions of the bay.
164
-------
The zooplankton data used were collected on the above cruises at the
same station-depth combinations as the chemical data. This work was con-
ducted by the University of Michigan, Pellston Biological Station (Gannon
[1975]). Zooplankton species counts were converted directly to dry weight
concentrations and then integrated to the level of the two functional groups
in the model. Work is progressing on phytoplankton species counts and cell
volume measurements.
Nutrient loadings to Saginaw Bay from the Saginaw River, the primary
source, were determined on the basis of a field sampling program. For the
first half of the year, samples were taken at two- to three-day intervals
at the Dow Chemical Company water intake plant at the mouth of the Saginaw
River. From July to December, samples were taken from the Midland Street
Bridge in Bay City every two weeks. During this period, the Dow intake
plant was too strongly influenced by the bay itself. The Midland Street
Bridge is approximately 5 miles upstream from the river mouth and is not
influenced by the bay during this period. Concentrations were obtained for
chloride and total and dissolved forms of phosphorus, nitrogen, and silicon.
Daily flow rates were obtained from the U.S. Geological Survey.
BOUNDARY CONDITIONS AND FORCING FUNCTIONS
Since the physical system under consideration is only part of a larger
physical system, Lake Huron proper, the interaction between Saginaw Bay and
Lake Huron is extremely important. The predominant flow pattern in the bay
is counterclockwise with Lake Huron water flowing in along the north shore
and a mixture of Lake Huron water and Saginaw River water flowing out of the
bay along the south shore (Figure 2). The concentrations of nutrients and
biota in the water which flows across the indicated inner-outer boundary are
examples of boundary conditions which must be specified. These concentrations
were determined using the cruise data from two field stations nearest to the
area of water inflow from the outer bay. Daily concentration values were cal-
culated by linear interpolation between the cruise averages for these stations,
Before the model can be implemented, various quantities known as forcing
functions must be specified. Conceptually, the physical system is described
by a number of quantities called state variables (Figure 1). If, at a given
time, values are specified for each of these state variables, then the com-
plete state of the system is known. It is desired to use the model to calcu-
late the state of the system at some future time. However, in order to do
this for the present system, various quantities such as water circulation
rates, light, temperature, and external nutrient loadings must be specified.
These quantities are forcing functions and they are unique to the physical
system under consideration.
External nutrient loads and water circulation rates are the most important
forcing functions in the present study. Total daily flow from the Saginaw
River was calculated by summing the primary tributary gauges and the estima-
ted flow from the ungauged tributary area. Daily nutrient loading rates were
calculated using the measured nutrient concentrations on that day. These
daily loading rates were then plotted and time-series of loading rates were
generated by linearly interpolating between all of the significant peaks and
troughs. For example, for total phosphorus, a series of 46 loading rates/
165
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time-breaks was generated. For orthophosphorus, a series of 46 loading
rates/time-breaks was generated. Water circulation rates between the inner
and outer bay were determined by modeling chloride concentrations in the
bay and chloride loadings from the Saginaw River (Richardson [1976]). Time-
variable flows were used which corresponded to hydraulic detention times
ranging from 45 to 120 days for the inner bay.
RESULTS OF SAGINAW BAY SIMULATIONS
Calibration results are presented in Figures 3-10. Note that this model
output is an attempt to describe an existing data set and is not intended to
be predictive in nature.
To obtain model output for chlorophyll a, the total biomass concentra-
tion for all four phytoplankton classes in tfie model was converted to chloro-
phyll ^concentration using 20 yg chlorophyll a/mg dry weight biomass. This
conversion factor was based on fresh weight biomass and chlorophyll a data
for Saginaw Bay (Vollenweider et al. [1974]), assuming that dry weigFt biomass
is 20% of fresh weight biomass (Kuenzler and Ketchum [1962]). However, these
data were collected at only one station in the inner part of the bay on eight
separate cruises during the 1971 growing season. These data do not necessar-
ily represent the average condition of the inner bay for 1974.
Preliminary simulations of chlorophyll a in Saginaw Bay (Bierman [1976])
showed that the model output was significantly higher than the data during
the month of June. This problem has been eliminated in the present work (Fig-
ure 3) by using variable water circulation rates. Richardson (1976) has shown
that two distinct flow regimes exist in Saginaw Bay, spearated by a turbulent
transition period in June. Previously, this high June flushing rate was not
modeled and only a constant, annual-average hydraulic detention time of 60
days was used.
^ 50
en
a.
40
re
Q-
o
CC
o
3:
o
o
30
20
o
<-> 10
o
0
MEAN± Vz STANDARD DEVIATION
MODEL OUTPUT
JAN ' FEB I MAR ' APR ' MAY ' JUN \ JUL ' AUG ' SEP ' OCT ' NOV ! DEC
TIME
Figure 3. Corrected chlorophyll ^distribution for 1974 in Saginaw
Bay, inner portion, as compared to model output.
166
-------
The class composition of the model output indicates that the early
phytoplankton crops are dominated by diatoms and green algae and that the
broad Summer-Fall peak is dominated by blue-green algae (Figure 4). A
similar successional pattern was observed in the inner bay by Vollenweider
et al. (1974). Chartrand (1973) reported significant late-Summer crops of
Aphanizomenon, a filamentous, blue-green alga in the outer bay, which also
corresponds to model output. It is not possible at this time to rigorously
calibrate the model at the level of the four functional groups of phyto-
plankton. A more rigorous calibration depends on biomass data for each of
the phytoplankton classes.
o
o
JAN I FEB ' MAR ' APR ' MAY ' JUN ' JUL ' AUG ' SEP ' OCT ' NOV ' DEC
TIME
Figure 4. Phytoplankton class composition of model output in
Figure 3.
The model output for total zooplankton biomass is much lower than the
actual data (Figure 5). This might indicate that the zooplankton kinetics
were modeled incorrectly because the model output for phytoplankton chloro-
phyll closely matches the actual data. However, chlorophyll does not appear
in any of the kinetic equations of the model. The phytoplankton-zooplankton
interaction is parameterized completely in terms of dry weight biomass. Since
the only comprehensive phytoplankton data available at this time are chloro-
phyll data, neither the actual phytoplankton biomass nor the class composition
of this biomass are accurately known. This uncertainty in the determination
of phytoplankton biomass could be an alternative explanation for the discrepancy
between the model output and the total zooplankton data.
167
-------
2.5
2.0
Q_
O
o
M —
1.5
O C71
5
1.0
0.5
o
o
" MEAN ± 1I2 STANDARD DEVIATION
MODEL OUTPUT
JAN I FEB I MAR ! APR ' MAY ' JUN ' JUL ' AUG ' SEP ' OCT ' NOV ' DEC
TIME
Figure 5. Total zooplankton biomass distribution for 1974 in Saginaw Bay,
inner portion, as compared to model output.
Model output for total phosphorus (Figure 6) and total nitrogen (Fig-
ure 7) is reasonable, with the possible exception of the late-fall period
for total phosphorus. Since the only external nutrient sources considered
were the Saginaw River and Lake Huron, the present results must be considered
preliminary in nature. The possible roles of sediments and atmospheric sources
must be considered before a complete picture of the nutrient dynamics in Sagi-
naw Bay can be obtained.
The general
ures 8-10) agree
qualifications
forms.
patterns of the model output for dissolved nutrients (Fig-
reasonably well with the actual data. However, the above
for total nutrients must also be applied to the dissolved
168
-------
100
en
a.
o;
o
80
£60
o
£ 40
o
£
tE 20
o
o
MEAN ± J/2 STANDARD DEVIATION
MODEL OUTPUT
' JAN I FEB I MAR ' APR ' MAY ' JUN ' JUL ' AUG ' SEP ' OCT ' NOV ' DEC
TIME
Figure 6. Total phosphorus distribution for 1974 in Saginaw Bay, inner
portion, as compared to model output.
2500
en
a.
o
o
Cd
<.
o
o
O
Z
O
o
2000
1500
1000
500
0
MEAN ± V2 STANDARD DEVIATION
MODEL OUTPUT
LL
JAN I FEB I MAR ' APR ' MAY ' JUN ' JUL ' AUG ' SEP ' OCT ' NOV ' DEC
TIME
Figure 7. Total nitrogen (TKN plus nitrate/nitrite) distribution for 1974
in Saginaw Bay, inner portion, as compared to model output.
169
-------
X
Q_
cn
n
30
1/1 OO
<=> O
Ll_ X
h^ Q_
I—
LU
O
O
o
10
MEAN ± Vz STANDARD DEVIATION
MODEL OUTPUT
JAN I FEB IMAR ' APR ' MAYUUN ' JUL ' AUG ' SEP ' OCT ' NOV ' DEC
TIME
Figure 8. Dissolved orthophosphate phosphorus distribution for 1974 in
ocnn Saginaw Bay, inner portion, as compared to model output.
cXKJ
0
o
QL
O
2000
> 1500
o
I
h-
UJ
O
o
o
500
0
MEAN ± lh STANDARD DEVIATION
MODEL OUTPUT
JAN I FEB ' MAR ' APR ' MAY ' JUN ' JUL ' AUG ' SEP ' OCT ' NOV ' DEC
TIME
Figure 9. Dissolved inorganic nitrogen (ammonia plus nitrate) distribution
for 1974 in Saginaw Bay, inner portion, as compared to model output
170
-------
1000
en
a.
o 800
600
o
CO
uo
400
£ 200
LU
O
o
MEAN + */2 STANDARD DEVIATION
JAN I FEB I MAR ' APR ' MAY ' JUN ' JUL ' AUG ' SEP ' OCT ' NOV ' DEC
TIME
Figure 10. Dissolved silicon distribution for 1974 in Saginaw Bay, inner
portion, as compared to model output.
DISCUSSION
In general, ambiguities can occur when attempting to calibrate a math-
ematical model if the model contains state variables or coefficients for
which there are no direct measurements. This is usually the case with eco-
systems models because the state of the art is still very primitive. There
have been few comprehensive experimental programs designed to provide field
data and rate coefficients for such models. The present model will be
refined as measurements for additional state variables and rate coefficients
become available.
One of the causes for the discrepancy between the model output and the
total zooplankton data could be the lack of simultaneous measurements of
phytoplankton biomass and zooplankton biomass. This problem is not neces-
sarily unique to multi-class phytoplankton models, but can also occur with
conventional chlorophyll models as well. A multi-class approach merely adds
another dimension to this ambiguity because zooplankton are not considered
to graze all of the phytoplankton classes. There is uncertainty in the
partitioning of the total phytoplankton biomass among the various classes,
as well as uncertainty in the total phytoplankton biomass itself. The only
way to determine if resources should be expended to refine the phytoplankton-
zooplankton interaction kinetics in the model is to first obtain comprehensive
phytoplankton cell counts and cell volumes. Such a determination is being
conducted for Saginaw Bay and the results will be incorporated in subsequent
versions of the model.
171
-------
The lack of direct measurements for all of the independent rate coeffi-
cients in a model can still result in model output which corresponds closely
to the actual data. However, cause-effect inferences can only be made with
great caution in these cases. In such circumstances, a model can be valuable
as a research tool if it is used to conduct sensitivity analyses for the pur-
pose of determining the most important coefficients.
It should be noted that, as the number of state variables for which
there is comprehensive data is increased, many coefficients in a model become
more tightly constrained. For example, phytoplankton sinking rates of 0.15
to 0.40 meters/day were used in preliminary calibration work which involved
only chlorophyll and dissolved nutrients (Bierman [1976]). Using the present
expanded data set, it was found that phytoplankton sinking rates could not
exceed 0.20 meters/day without causing a very significant discrepancy between
the model output and the total phosphorus data. If an ecosystem model can be
calibrated to a large number of simultaneous and independent parameters, its
reliability as a tool for drawing cause-effect inferences can be greatly in-
creased. The present model will eventually be calibrated to at least 12 simul-
taneous and independent parameters and tested against similar comprehensive
data from the 1975 field sampling program on Saginaw Bay.
ACKNOWLEDGEMENTS
The authors would like to thank John E. Gannon, University of Michigan
Biological Station, Pellston, Michigan, for providing unpublished zooplankton
data for Saginaw Bay. Donald C. McNaught, State University of New York, Albany,
provided assistance in the interpretation and reduction of these data. V.
Elliott Smith, Cranbrook Institute of Science, provided the chemical and chloro-
phyll data for Saginaw Bay. David M. Dolan, EPA, Large Lakes Research Station,
contributed many valuable suggestions during this study. The authors would like
to thank J. Kent Crawford, Nelson A. Thomas and David M. Dolan, all from EPA,
Large Lakes Research Station, for reviewing the manuscript.
REFERENCES
Bierman, V.J., Jr. 1976. Mathematical Model of the Selective Enhancement
of Blue-Green Algae by Nutrient Enrichment. In: Modeling Biochemical
Processes in Aquatic Ecosystems. (R.P. Cahale, ecITtor). Ann Arbor
Science Publishers, Inc. Ann Arbor, MI pp. 141-169.
Canale, R.P., Nachiappan, S., Hineman, D.J. and Allen, H.E. 1973. A Dynamic
Model for Phytoplankton Production in Grand Traverse Bay. Proceedings,
Sixteenth Conference on Great Lakes Research, April 16-18, 1973. Inter-
national Association for Great Lakes Research, Huron, Ohio pp. 21-33.
Caperon, J. and Meyer, J. 1972a. Nitrogen-Limited Growth of Marine Phyto-
plankton-I. Changes in Population Characteristics with Steady-State
Growth Rate. Deep Sea Research 19: 601-618.
Caperon, J. and Meyer, J. I972b. Nitrogen-Limited Growth of Marine Phyto-
plankton-II. Uptake Kinetics and Their Role in Nutrient Limited Growth
of Phytoplankton. Deep Sea Research 19: 619-632.
Chartrand, T.A. 1973. A Report on the Taste and Odor in Relation to the
Saginaw-Midland Supply at Whitestone Point in Lake Huron. Saginaw
Water Treatment Plant, Saginaw, Michigan.
172
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Chen, C.W. and Orlob, G.T. 1972. Ecologic Simulation for Aquatic Environments.
Water Resources Engineers, Inc., Walnut Creek, California. Report prepared
for Office of Water Resources Research, U.S. Department of the Interior.
DePinto, J.V., Bierman, V.J., Jr. and Verhoff, F.H. 1976. Seasonal Phyto-
plankton Succession as a Function of Species Competition for Phosphorus
and Nitrogen. In: Modeling Biochemical Processes in Aquatic Ecosystems.
(R.P. Canale, edTtor). Ann Arbor Science Publishers, Inc. Ann Arbor, MI
pp. 141-169.
Droop, M.R. 1973. Some Thoughts on Nutrient Limitation in Algae. Journal
of Phycology 9: 264-272.
Fuhs, 6.W. 1969. Phosphorus Content and Rate of Growth in the Diatoms C.yclotella
nana and Thai assiosira fluviatilis. Journal of Phycology 5: 312-321.
Fuhs, G.W., Demmerle, S.D., Canelli, E. and Chen, M. 1971. Characterization
of Phosphorus-Limited Planktonic Algae. Nutrients and Eutrophication: The
Limiting Nutrient Controversy. Proceedings of a Symposium, February 11-12,
1971. American Society of Limnology and Oceanography and Michigan State
University, East Lansing, Michigan, pp. 113-132.
Gannon, J.J. 1975. Crustacean Zooplankton in Saginaw Bay, Lake Huron. A
report to the International Reference Group on Upper Lakes Pollution,
International Joint Commission, Windsor, Ontario.
Kuenzler, E.J. and Ketchum, B.H. 1962. Rate of Phosphorus Uptake by
Phaeodactylum tricornutum. Biological Bulletin 123: 134-145.
Middlebrooks, E.J., Falkenburg, D.H. and Maloney, I.E. (Eds). 1973.
Modeling the Eutrophication Process. Proceedings of a Workshop,
September 5-7, 1973. Utah Water Research Laboratory and Division
of Environmental Engineers, Utah State University, Logan. National
Eutrophication Research Program, U.S. Environmental Protection Agency,
Corvallis, Oregon.
O'Connor, D.J., Thomann, R.V. and DiToro, D.M. 1973. Dynamic Water Quality
Forecasting and Management. U.S. Environmental Protection Agency,
Corvallis, Oregon, Ecological Research Series EPA-660/3-73-009.
Richardson, W.L. 1976. An Evaluation of the Transport Characteristics of
Saginaw Bay Using a Mathematical Model of Chloride. I_n_: Mathematical
Modeling of Biochemical Processes in Aquatic Ecosystems. (R.P- Canale,
editor). Ann Arbor Science Publishers, Inc., Ann Arbor, MI. pp. 113-139.
Smith, V.E. 1975. Saginaw Bay (Lake Huron): Survey of Physical and Chemi-
cal Parameters. A report to the International Reference Group on Upper
Lakes Pollution, International Joint Commission, Windsor, Ontario.
Thomann, R.V., DiToro, D.M., Winfield, R.P. and O'Connor, D.J. 1975.
Mathematical Modeling of Phytoplankton in Lake Ontario. I. Model
Development and Verification. U.S. Environmental Protection Agency,
Corvallis Oregon, Ecological Research Series EPA-660/3-75-005.
Vollenweider,, R.A., Munawar, M. and Stadelmann, P. 1974. A Comparative
Review of Phytoplankton and Primary Production in the Laurentian Great
Lakes. Journal of the Fisheries Research Board of Canada 31: 739-762.
173
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Implications of Resource Development on the
North Slope of Alaska with Regard to Water
Quality on the Sagavanirktok River
Eldor W. Schallock*
ABSTRACT
The Sagavanirktok (Sag) River, located on the North
Slope of Alaska is undergoing a rapid transition from
an isolated undisturbed river system to an accessible
impacted watershed. Impact is caused by: 1. The de-
mands of rapidly expanding industry drawing heavily
upon some of the available resources such as water and
gravel; 2. The extended arctic conditions that affect
the environment which include 8 months of winter (Octo-
ber through May), permafrost a few centimeters beneath
the surface, and annual"precipitation in the coastal
province of about 14 cm; and 3. The specific water
quality characteristics of the river that are sometimes
limited and critical. During winter, stream discharge
virtually ceases, dissolved oxygen concentrations are
low (1.2 mg/1), specific conductance may be high (1700
umhos), and nutrients may be high (0.76 mg/1 nitrate
as nitrogen and 12.5 mg/1 silica). The impact of in-
dustry on these water quality characteristics may affect
indigenous aquatic biota.
INTRODUCTION
The Sagavanirktok (Sag) River, located on the North Slope of Alaska,
(Figure 1) is undergoing a rapid transition from an isolated, undisturbed
stream system to an accessible impacted drainage. This transition started
in 1968 with discovery of oil at Prudhoe Bay near the mouth of the river.
It is continuing with the construction of the Trans-Alaska pipeline (Alyeska
Pipeline) which traverses approximately 200 km (125 miles) through the heart
of the Sag River Basin. Further transition is a promise for the future with
the continued search and development of oil and as the quest for other natural
resources begins. The anticipated completion 'of the bridge crossing the Yukon
River will complete the all-weather road connecting the Alaskan Arctic to the
existing State road system and will enable additional access to the Sag River.
*Corvallis Environmental Research Laboratory, U.S. Environmental Protection
Agency, Arctic Environmental Research Station, College, AK 99701
174
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ARCTIC OCEAN
POINT
BARROW
Ln
FIGURE I. MAP OF ALASKA SHOWING
LOCATIONS OF STREAM SYSTEMS
-------
This new prominence in the world of resources, and the linked acces-
sibility, is impacting a set of terrestrial and aquatic environmental con-
ditions that are substantially different from those in the contiguous 48
states. It is impossible to discuss all of them with the interrelationships
here, but a few pertinent factors will set the environmental scene and re-
late to implications of resource development.
CLIMATE
The Arctic climate may best be described as harsh when compared to
climates of most other areas. Ambient air temperature, one of the dom-
inant features, is characterized by a mean annual range of -12°C (10°F)
to -7°C (19°F) (Watson, 1969) and by severe winter temperatures as low
as -54°C (-65°F) in localized interior areas. As little as 10 cm (4 inches)
of precipitation may collect along the coast (Johnson et al., 1969), and
winter precipitation as snow is often redeposited by strong and persistent
winds (Watson, 1969).
Another factor that contributes to the harshness is lack of solar ra-
diation during the winter. Prudhoe Bay, which is near 70°M latitude, has
no direct sunlight from the middle of November to mid-January, but during
the summer has continuous daylight from the middle of May to early August.
The above factors affect permafrost which becomes unstable when the
temperature equilibrium of the system is disturbed (Proceedings of the First
International Conference on Permafrost, 1963). All of the Arctic is in the
continuous permafrost zone (Ferrians, 1969) with the thickness ranging up
to 396 meters (1300 feet).
Climate, permafrost, and geology of the area all affect the soil that
has developed, and vegetation types that have adapted to this ecosystem.
Several soil types have been described (Tedrow and Cant!on, 1958 and Tedrow
et al., 1958). These soil types support different vegetative communities
which may have as many as 300 different species of plants (Johnson et al.,
1964). These plant communities in turn dampen the soil temperature extremes,
retard heat penetration, reduce the rate of soil and frost erosion and there-
by maintain the permafrost integrity (Johnson, 1963).
WATER QUALITY
A baseline water quality survey of the Sag River basin was conducted
by the Arctic Environmental Research Laboratory in 1969-70. The water qual-
ity of the river at that time could be characterized as good during the
summer open water period. Breakup and summer precipitation in the form of
spates can cause temporary changes and deterioration in the quality of some
parameters. This, however, is usually a short-lived phenomenon and does not
create serious problems for the indigenous biota, or for the domestic and
industrial users. Breakup occurs during early June although the timing, mag-
nitude and related problems such as flooding are dependent upon a combination
of factors and may vary widely from year to year.
176
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Sag River water quality during winter differs substantially from sum-
mer. The differences are caused by a complex interaction of geologic and
climatological factors that affect numerous water quality parameters. Se-
lected summer and winter data are presented in Table 1. These data are
similar to results obtained from the Colvilie, Kuparuk and Sag Rivers by
the U.S. Geological Survey in 1964 and 1972.
TABLE i. SELECTED WINTER AND SUMMER WATER QUALITY DATA
FROM SAMPLES COLLECTED FROM THE SAGAVANIRKTOK
RIVER 1969-1970
Parameter
Silica (mg/1)
Total phosphate (mg/1)
Nitrate (mg/1)
Ammonia (mg/1)
Calcium (mg/1)
Potassium (mg/1)
Sodium (mg/1)
pH
Specific conductance
(umhos)
Range During
0.6 -
0.01 -
0.05 -
0.02 -
Summer
2.7
0.05
0.15
0.09
10.0 - 42.0
0.15 -
0.40 -
7.6 -
80 -
0.75
1.3
8.1
240
Range During
3.6 -
0.01
0.09 -
0.01 -
89.0
0.7 -
2.6 -
7.2 -
660 -
Winter
12.5
0.76
0.18
95.0
1.97
9.0
7.7
1700
Comparison of summer data to winter data in Table 1 shows that several
water quality parameters deterioriated appreciably in winter. In many in-
stances, the ranges found during the winter were significantly higher than
those found during the summer. These trends are supported by dissolved
solids and hardness data from the Colvilie and Sag Rivers by Feulner, et al.,
1971. These higher concentrations during winter are probably caused by a
combination of extrusion of salts during the freezing process, accumulating
metabolites and salt water intrusion in coastal areas.
Dissolved oxygen (DO) is often considered to be one of the critical
parameters affecting aquatic life but until recently was not considered to
be a problem in Alaska. The data presented in Figure 2 indicate that DO
ranges from 9.9 mg/1 (91.7 percent saturation) to 13.3 mg/1 (95 percent
saturation) during the summer but extremely low concentrations may be found
during the winter (1.2 mg/1; 8.2 percent saturation). These concentrations
are comparable to other DO data on the rivers of the North Slope and to the
general spatial and seasonal DO patterns in the Chena, Chatanilca, Tanana,
and Yukon Rivers of interior Alaska (Schallock and Lotspeich, 1974).
177
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u
o
UJ
at
QC
LU
Q.
14
12
10
8
6
4
2
14
01 12
£ 10
(D
X
O 8
o
UJ
i/i
Q 4
JUNE
13X)0
1100
?6o'700560
(mouth)
300 ido
JUNE
APRIL
/\
O
o
(moulh)
S-1300 1100 900 700 500 300 100
1200 1000 800 600 400 200
Figure 2. Dissolved oxygen and water temperature data from
13 stations on the Sagavanirktok River (1969-1970).
A consistent inverse relationship was found between water temperature
and concentrations of DO. During the June study interval, the water tem-
perature reached the recorded maximum of about 12°C (54°F). Water tempera-
tures were generally highest in the foothill province near Sagwon (Station
700) about 104 km (65 miles) inland from Prudhoe Bay.
178
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WATER QUANTITY
Stream discharges and seasonal timing of these discharges are char-
acteristics causing great concern both in industry and in the agencies
charged with resource management. Long periods of low discharge during
winter followed by an increase during spring and rapid decline after break-
up are the "normal" discharge patterns of the Sag and Patuliqayak (Put)
Rivers (Figure 3). '
(Both scales in thousands)
(ft3/sec)
Figure 3. Seasonal discharge in the Sagavanirktok (Sag) and
Putuligayak (Put) Rivers from October 1973 through
September 1974 (U.S.G.S. Data).
Spring breakup accounts for the majority of the annual discharge. In
the Put River, over 71 percent of the annual volume is discharged between
June 1st and June 15th and approximately 94 percent of the annual volume by
June 30th. In the Sag River, the results are not as dramatic if the same
comparisons are made because the seasonal high discharge pattern is bimodal.
Between June 1st and June 15th, over 16 percent of the total is discharged
while 10 percent is discharged between June 16th and June 30th. A higher
peak with shorter duration occurred between August 16th and August 31st and
accounts for 20 percent of the total. The short term, high volume discharges
of both the Put and Sag Rivers rapidly decrease to smaller volumes that ap-
parently originate from ground water sources. This smaller volume is reached
in July and October, respectively. The Colville River, the largest drainage
on the North Slope, may discharge as much as 43 percent of the annual runoff
in a 3-week period (Alexander, et al., 1974).
179
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Some debate has focused on whether the discharge of North Slope ri-
vers during the winter is just extremely low or whether discharge actually
ceases in some instances. In either case, a limited volume of water is
available for man's use.
AQUATIC BIOLOGY
Prior to discovery of oil at Prudhoe Bay, little information has been
published on the life history patterns of the indigenous fish and only oc-
casional data were available on microbiological or macrobenthic communities.
Since 1968, moderate effort has gone into examining various characteristics
of specific biological populations and communities.
Spawning areas, migration patterns and overwintering areas were chosen
as possible limiting factors to any population of fish inhabiting the Sag
River. The two most important fish in the Sag River are the Arctic Char,
Salvelinus alpinus, and the Arctic Grayling, Thymallus arcticus. Lake Trout,
Salvelinus namaycush, are also found in the river basin.
Arctic char and grayling are generally distributed throughout the Sag
drainage but may be concentrated in localized areas of the Sag main stem or
its tributaries during the summer. Some migrations such as the upstream mi-
gration of adult and subadult char during early summer are well documented
while other migrations are suspected but are not well documented.
Spawning areas utilized by char and grayling are difficult to locate
because the char spawn in autumn when ice cover is beginning and grayling
spawn during the high water of spring. Both fish establish redds in gravels
with specific characteristics.
Overwintering of young-of-the-year and juveniles are also difficult to
determine. Studies by the Alaska Department of Fish and Game and the U.S.
Fish and Wildlife Service documented overwintering populations in spring
areas that are primarily located in the upper Sag drainage and tributaries.
However, Furniss (1975) recently demonstrated that young-of-the-year and/or
juvenile fish are also utilizing the deep pools of the Sag River from Franklin
Bluffs to the coast.
Members of the macrobenthic community are ubiquitously distributed along
the length of the river and are sensitive to water quality parameters because
the organisms are relatively immobile. This community consists principally
of Plecoptera (stoneflies), Ephemeroptera (mayflies), Chironomidae (midges),
Trichoptera (cadisflies), and oligochaetes. These organisms may number as
high as 400 per square meter (330 per square yard) during summer conditions
(Schallock and Mueller, 1970) and are the primary food items for most life
stages of fish.
Unpublished data collected by Gordon during 1969-1970 showed that the
number of local coliforms present in the Sag River was low at that time.
However, his recent work (Gordon, 1975) on the survival of enteric micro-
organisms in the Tanana River near Fairbanks reveals that the addition of
these microorganisms into a limited and closed system such as the Sag River
could be extremely dangerous to users. This danger appears where numerous
activities are located along the drainage and the river is utilized as a
source of potable water.
180
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RESOURCE AND MANAGEMENT IMPLICATIONS
WATER APPROPRIATIONS
Quantity and water quality are problems on Alaska's North Slope. Sea-
sonal discharge patterns and changes in the water quality of North Slope
rivers have been presented during early discussion. The demands for water
for both domestic and specific industrial activities are stressing the limited
supplies. This is causing industry to travel some distance to obtain water
and also to consider alternative sources and techniques for obtaining water.
The customary method for collecting water has been to set up a pump sta-
tion at the river or to drill a hole in the ice and utilize a tanker truck to
collect and transport the water. Pipeline construction activity has resulted
in the demand for water to increase drastically; consequently some "watering
holes" have been pumped dry.
Initially water was collected from the river immediately adjacent to the
particular use. However, this past winter water was hauled as far as 72 km
(45 miles) when it was not available in the immediate area. In one instance,
holes were drilled through the river ice on a grid system with holes as close
as 15 meters (50 feet). Whenever water was found, radio communications to
identify the specific location were made to the water truck. The water truck
then came to the water hole and pumped until that particular hole was dry and
then the search for another water hole began.
Tundra lakes have been considered as an additional source but these lakes
generally have small volumes of poor quality water that must be treated to be
potable.
GRAVEL MINING
Gravel mining is presently causing tremendous concern, for this activity
has not been given adequate regulation. One has only to examine the variety
of endeavors utilizing gravel and the magnitude of its use is brought into fo-
cus. All roads, airstrips, pads for building, drill rigs and pipelines require
large amounts of gravel. It is of particular important in those areas where
a permanent gravel foundation is needed to maintain permafrost integrity.
An accurate estimate of the gravel requirements for pipeline construction
was impossible before construction began. Early estimates, however, placed the
amount near 4.6 million cubic meters (6 million cubic yards) for the entire
pipeline while some Federal resource managers at the time estimated the amount
to be closer to 7.6 million cubic meters (10 million cubic yards). At the pre-
sent time, nearly twice this amount has been extracted from the lands adminis-
tered by the Bureau of Land Management on the North Slope alone (Dean, 1975).
It is now estimated that as much as 161 million cubic meters (210 million cubic
yards) have been used along the pipeline and about 18 months of construction are
still remaining.
Adequate amounts of suitable gravel are not readily available in many areas
of the North Slope. However, the Sag River and its tributaries provide a perma-
frost free "thaw bulb" wherein it is economically feasible to borrow gravel. As
a result, gravel used for airstrips, roads, pipeline pads and drill pads is being
removed only from these frost-free areas.
181
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Limited availability and resultant overhauls have increased the cost of
gravel. Gravel purchased from the material site for approximately 18 cents
per cubic meter (14 cents/cubic yard) may cost as much as $52 per cubic me-
ter ($40/cubic yard) by the time it is purchased, loaded, hauled, deposited
and spread at the deposition site. The shortage of gravel and its increasing
cost is causing industry to consider other methods and materials as a substitute
Gravel mining can adversely affect the water quality. It may cause ex-
cessive suspended sediment and increased stream bed load if the mining opera-
tion is improperly placed and an effective settling pond is not provided. The
addition of sediment to a stream system may in turn prevent primary production,
cover the macrobenthic community and cuase the fish to outmigrate or smother
the developing eggs and young-of-the-year. Continued gravel mining from the
flood plain of the Sag River may also cause hydrologic instability that would
require a long time to equilibrate. During this period of reestablishment,
suspended sediment with its attendant problems would likely continue at above
normal levels.
WASTE DISCHARGE
Waste discharges from permanent camps, temporary camps and drilling sites
are a serious threat to water quality of the Sag River. The addition of
oxygen demanding substances to an already severely depressed dissolved oxygen
system would be disastrous. Toxic substances, such as residual chlorine or
chloramines, heavy metals and hydrocarbons, if released into the water course,
particularly during the low flow periods, could rapidly eliminate desirable
fish and macrobenthic organisms. In addition, effluent containing enteric
microorganisms from human waste that enters the river may well become a part
of a downstream user's water supply.
BIOLOGICAL IMPLICATIONS
The Sag River is now experiencing the full impact of man's development
of a nonrenewable resource in arctic Alaska. What are the implications of
water appropriations, gravel mining and waste discharge on the river system?
The most obvious implication is an adverse effect on the aquatic resources of
the river. Overwintering populations of fish and macrobenthos will be impacted
by water use in the Franklin Bluffs to Prudhoe Bay area. Reliable reports have
been received of fish being found in water holding tanks. Other reports des-
cribe instances where the operator pumping water into the tank truck would have
to stop and clean the intake screen of fish carcasses. Still other reports des-
cribe invertebrates contained in water supplies.
Low discharge during winter also magnifies the impact on the stream when
it is used as the receiving water for domestic or industrial effluent. Small
volumes of water, low winter dissolved oxygen and high concentrations of dis-
solved constituents all combine to create a system that is highly sensitive to
an effluent containing oxygen demanding materials and/or toxic substances such
as residual chlorine. With the addition of enteric microorganisms into this
limited and closed system, public health becomes a serious consideration where
activities are located along the drainage and utilize the river as a source of
potable water.
182
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Finally, aquatic organisms within the river system may be deleteriously
affected by gravel mining operations either through the direct addition of
sediment to the system or indirectly by hydro!ogical instability.
REFERENCES
Alexander, V., D.C. Burrell, J. Chang, T.R. Cooney, C. Coulon, J.J. Crane,
J.S. Dygas, G.E. Hall, P.J. Kinney, D. Kogl , J.C. Mowatt, A.S. Naidu,
I.E. Osterkamp, D.M. Schell , R.D. Seifert, and R.W. Tucker, 1974.
"Environmental Studies of an Arctic Estuarine System." Final Report,
Institute of Marine Science, University of Alaska, College, Alaska,
Report R-74-1. U.S. Environmental Protection Agency Research
Grant No. 16100EOM.
Dean, T. , 1975. Personal communications concerning gravel mining on
Alaskan North Slope lands administered by the Bureau of Land Management.
U.S. Department of Interior, Bureau of Land Management, Fairbanks,
Alaska.
Ferrians, O.J., 1969.
Inv. Map #1-445.
Permafrost Map of Alaska. U.S. Geological Survey
Feulner, A.J., J.M. Childers, and V.W. Norman, 1971. "Water Resources Data
for Alaska, 1971." U.S. Geological Survey, 1972. Publ . Anchorage, AK.
Furniss, R. , 1975. Personal communication concerning fish overwintering areas
and water appropriation in the Sagavanirktok River. Alaska Department
of Fish and Game, Fairbanks, Alaska.
Gordon, R.C., 1970. Unpublished water quality data from the Sagavanirktok
River, Environmental Protection Agency, Alaska Water Laboratory,
College, Alaska.
Gordon, R.C., 1975. Personal communication concerning the survival of enteric
microorganisms in a subarctic river near Fairbanks, Alaska. Environmen-
tal Protection Agency, Arctic Environmental Research Laboratory, College,
Alaska.
Johnson, A.W., 1963. "Ecology in Permafrost Areas." Proceedings of First
International Conference on Permafrost, presented by Building Research
Advisory Board, National Academy of Science, Washington, D.C., pp. 25-30.
Johnson, A.W., L. Viereck, R. Johnson and R. Melchoir, 1964. "The Vegetation
and Flora of the Ogotoruk Creek—Cape Thompson Area, Alaska," Environment
of the Cape Thompson Region, Alaska, Norman J. Wilimovsky, Editor. U.S.
Atomic Energy Commission, Division of Technical Information, Oak Ridge,
Tennessee, pp. 277-363.
183
-------
Johnson, P.R., and C.W. Hartman, 1969. "Environmental Atlas of Alaska."
Institute of Water Resources, University of Alaska, College, Alaska, 111 p.
Netsch, N., 1975. Personal communication regarding water quality, water
quantity and related problems in the Sag River. Anchorage, Alaska.
U.S. Fish and Wildlife Service, Anchorage, Alaska.
Proceedings of the First International Conference on Permafrost, 1963.
Presented by Building Research Advisory Board, National Academy
of Science, Washington, D.C.
Schallock, E.W. and E.W. Meuller, 1970. Unpublished data on the Sagavan-
irktok River. Environmental Protection Agency, Arctic Environmental
Research Laboratory, College, Alaska.
Shallock, E.W. and F.B. Lotspeich, 1974. "Low Winter Dissolved Oxygen in
Some Alaskan Rivers." U.S. Environmental Protection Agency, National
Environmental Research Center, Corvallis, Oregon, 33 p.
Tedrow, J.C.F. and J.E. Cantlon, 1958. "Concepts of Soil Formation and
Classification in Arctic Regions." Arctic, Vol. 11, pp. 166-179.
Tedrow, J.C.F., J.V. Drew, D.E. Hill, and L.A. Douglas, 1958. "Major Genetic
Soils of the Arctic Slope of Alaska." J. Soil Science. Vol. 9,
pp. 33-45.
U.S. Geological Survey, 1964. Compilation of Records of Surface Waters of
Alaska, October 1950 to September 1960. Anchorage, Alaska.
U.S. Geological Survey, 1972. Water Resources Data for Alaska, 1971.
Anchorage, Alaska.
Watson, C.D. 1969. Climates of the States: Alaska. U.S. Weather Bureau,
Climatography of the U.S., No. 60-49, 24 p.
184
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Lake Eutrophication: Results from
The National Eutrophication Survey
Jack H. Gakstatter, Marvin 0. Allum
and James M. Omernik*
INTRODUCTION
In early 1972, the U.S. Environmental Protection Agency (EPA) initiated
the National Eutrophication Survey (NES) program to: (1) identify those
lakes and reservoirs in the contiguous United States that receive nutrients
from the discharges of municipal sewage treatment facilities, and (2) deter-
mine the significance of these point-source nutrient inputs to the nutrient
levels and the primary productivity of each system. After the program began,
additional federal legislation was passed (Public Law 92-500), and NES ob-
jectives were broadened to include an assessment of the relationships of non-
point sources; e.g., land use, to lake nutrient levels and also to assist in
establishing water-quality criteria for nutrients.
SELECTION CRITERIA
Freshwater lakes and impoundments in the Survey were selected through
consultation with EPA Regional Offices and state pollution control agencies,
as well as related state agencies managing fisheries, water resources, or
public health. EPA established selection criteria to limit the type and
number of candidate water bodies, consistent with existing Agency water goals
and strategies. For 27 states of the eastern United States where lakes were
selected prior to passage of P.L. 92-500, strongest emphasis was placed on
lakes faced with actual or potential accelerated eutrophication problems; i.e.
an artificially increased rate of algal and/or aquatic plant production. As
a result, the selected lakes:
1. were impacted by one or more municipal sewage treatment plants,
either directly or by discharge to an inlet tributary within
approximately 25 miles of the lake;
2. were 100 acres or larger in size; and
3. had mean hydraulic retention times of at least 30 days.
*Corvallis Environmental Research Laboratory, U.S. Environmental Protection
Agency, Corvallis, OR 97330
185
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However, these criteria were waived for a number of lakes of particular
interest to the states.
In the western states, these criteria were modified to reflect revised
water-research mandates, as well as to address more prevalent non-point
source problems in agricultural or undeveloped areas. Thus each state was
requested to submit a list of candidate lakes for the Survey that:
1. were representative of the full range of water quality (from
oligotrophic* to eutrophic*);
2. were in the recreational, water supply, and/or fish and wildlife
propagation use-categories; and
3. were representative of the full scope of nutrient pollution prob-
lems or sources (from municipal waste and/or nutrient-rich indus-
trial discharges, as well as from non-point sources).
The size and retention time constraints applied in the eastern states
were retained as was the waiver provision.
In all cases, listings of potential candidate lakes or reservoirs, pre-
pared with the cooperation of the EPA Regional Offices, were made available
to the states to initiate the selection process.
In total, the Survey includes 812 lakes and reservoirs across the con-
tiguous 48 United States. Figure 1 shows the distribution of the lakes and
reservoirs by state and the year during which each water body was sampled.
GENERAL SURVEY METHODS
Several kinds of information are required as a basis for management
decisions regarding the need for point or non-point source control of phos-
phorus and perhaps other nutrients as well. The Survey purpose is to collect
the type of data which will provide a basis for such decisions or at least to
provide a data base which can be supplemented with more detail, if required.
First, an annual nutrient budget is estimated for each water body, differen-
tiating between inputs originating from point and non-point sources; second,
the existing trophic condition of the water body is evaluated by sampling;
and third, an algal assay is performed to determine whether phosphorus, ni-
trogen, or some other element is limiting primary productivity of the water
body. The methods used to gather this information are described below.
The operations aspects of the Survey are shared by branches of two EPA
laboratories (46 people) and a small headquarters staff (3 people). The
Environmental Monitoring and Support Laboratory at Las Vegas, Nevada (Las
Vegas-EMSL) is responsible for sampling each lake, doing the associated an-
alyses, evaluating a portion of the data, and reporting results. The Cor-
vallis Environmental Research Laboratory (CERL) at Con/all is, Oregon is
responsible for coordinating the sampling of streams and sewage treatment
plants, analyzing the samples, and performing the algal assay on lake samples
Oligotrophic--low nutrient concentrations and primary productivity.
Eutrophic--high nutrient concentrations and primary productivity.
186
-------
Oo
1975-152
GRAND TOTAL- 812
Figure 1. Number of lakes and reservoirs sampled in each state and year of
sampling by the National Eutrophication Survey.
-------
CERL also has major responsibility for evaluating the lake, stream, and
point-source data and incorporating these data into a report on each lake.
The headquarters staff (Washington, D.C.) makes the initial contact with
each state water pollution control agency to explain the function of the
Survey and to cooperatively determine which lakes and reservoirs will be
included. They also contact each State National Guard to explain the
function of the Survey and to request their assistance in meeting Survey
objectives by collecting monthly samples from selected tributaries to sur-
veyed lakes. In addition, the headquarters staff provides general coordin-
ation and guidance to the operational aspects of the program.
Because the Survey has to cover a large geographical area in a rela-
tively short period of time, pontoon-equipped UH-1H Bell helicopters with
automated and manually-operated instruments are used to measure the water
quality of each lake. Two helicopters - carrying a limnologist and a
technician - are operated simultaneously, and a third helicopter is used
for ferrying parts, equipment, and people. The sampling teams from the
Las Vegas-EMSL are supported by a mobile analytical laboratory, chemistry
technicians, electronic specialists, and other staff involved with heli-
copter maintenance or program coordination. The total staff in the field
usually ranges from 12 to 14 people.
Operating procedures involve establishing a work center at an airport
and then sampling all lakes within a 100-mile radius. When all of the water
bodies within the area are sampled, the support staff moves to a new central
location, and sampling begins on a different set of lakes. In this manner,
150 to 250 lakes have been sampled three times each year, and the sampling
will be completed on all of the 812 lakes in a four-year period.
Table 1 depicts the routine water-quality parameters which were selected
to characterize each lake and assess its trophic condition. Parameter selec-
tion was based on the relevance of each parameter as a measure of potential
and existing primary production. Both the number and the type of parameters
measured were also limited to a certain extent by the operational aspects of
the Survey.
TABLE 1. WATER-QUALITY CHARACTERISTICS MEASURED
Physical-Chemical
Alkalinity Nitrogen:
Conductivity* Ammonia
pH* Kjeldahl
Dissolved oxygen Nitrate
Phosphorus: Secchi depth
Ortho Temperature*
Total
Biological
Algal assay Algal count and
identification
Chlorophyll £
*Determined on-site with electronic probes.
188
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Concurrent with the lake sampling, the significant tributaries and
outlet(s) of each lake are sampled monthly, totaling about 4,200 sampling
sites nationwide. Volunteer National Guardsmen of each state, trained on-
site by EPA or state agency staff, collect and preserve the samples at sites
pre-selected by EPA personnel. The samples are shipped to CERL for analysis
of the various forms of nitrogen and phosphorus (see Table 1).
Through an interagency agreement, the U.S. Geological Survey estimates
flows for each sampled stream. These data are used in conjunction with con-
centration values to determine nutrient loadings.
A voluntary sampling program was established through the respective
state water pollution control agencies to have plant operators collect
effluent samples from those municipal sewage treatment plants which impact
Survey lakes—about 1,000 treatment plants. The effluent samples are col-
lected monthly, preserved, and shipped to the Con/all is laboratory for
nitrogen and phosphorus analyses.
Specific procedures used in collecting, preserving, shipping, and anal-
yzing the various kinds of samples collected by the Survey are described in
National Eutrophication Survey Working Papers No. 1 (1974) and 175 (1975).
Presently, the field portion of the Survey is almost completed with the
last samples scheduled for collection in November, 1975. Data analysis is
scheduled for completion in December, 1976.
RESULTS AND DISCUSSION
LIMITING NUTRIENTS
For each of the surveyed lakes, an algal assay is performed on a sample
of lake water and, to supplement the assay findings, inorganic nitrogen to
dissolved orthophorus ratios are determined from the lake sampling results.
For the 623 surveyed lakes in states east of the Rocky Mountains, the assay
demonstrated that with respect to algal growth requirements, 67% were phos-
phorus-limited, 30% were nitrogen-limited and 3% were either limited by an
element other than phosphorus or nitrogen or the results were not conclusive
(Table 2).
TABLE 2. SUMMARY OF ALGAL ASSAY RESULTS FOR SURVEYED WATER
BODIES IN THE 37 STATES EAST OF THE ROCKY MOUNTAINS
Limiting Nutrient Number of Lakes % of all Lakes
Phosphorus
Nitrogen
Other
417
186
20
67
30
3
Total 623 100%
189
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A higher percentage of phosphorus limited lakes would probably have
been found had the Survey not been mostly concerned with lakes which were
impacted by municipal wastes. The algal assay results should, therefore,
be evaluated with some caution because they reflect existing conditions
which often include man's impact on the nutrient regime.
Municipal waste treatment plant effluents, for example, have an average
total nitrogen to total phosphorus ratio of about 2.5 to 1 whereas natural
waters usually have a ratio in excess of 15 to 1. The relative abundance of
phosphorus provided in municipal effluents could change a lake from phosphorus-
limited to nitrogen-limited. Such a lake could theoretically be changed back
to phosphorus-limited by reducing phosphorus inputs.
Figure 2 is an indication of the significance of municipal wastes to
the total annual phosphorus load to some of the eastern lakes and reservoirs.
Of the 234 water bodies included in the frequency histogram, 135 receive
more than 20% of their annual total phosphorus load from municipal sources.
200
175
u 150
-1 125
u-
° 100
ce
LU ....
CD 75
2
ID
z 50
25
0-10 11-20 21-30 31-40 41-50 51-60 61-70 71-80 81-90 91-100
PERCENT OF TOTAL PHOSPHORUS LOAD FROM MUNICIPAL POINT SOURCES
Figure 2. A frequency histogram representing the percent of total
annual phosphorus load attributable to municipal wastes
for a number of eastern U.S. lakes and reservoirs.
190
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If 80% of the phosphorus were removed from these discharges bv treatment,
only 9 of the lakes would still receive more than 20% of their total phos-
phorus load from municipal wastes as shown in Figure 3.
200
175
IS 150
125
100
75
50
25
2
0-10 11-20 21-30 31-40 41-50 51-60 61-70 71-80 81-90 91-100
PERCENT OF TOTAL PHOSPHORUS LOAD FROM MUNICIPAL POINT SOURCES
FOLLOWING 80% REMOVAL
Figure 3. A frequency histogram representing the percent of
total annual phosphorus load attributable to municipal
wastes after 80% effluent phosphorus reduction for a
number of eastern U.S. lakes and reservoirs.
The reduction or removal of phosphorus originating from municipal sources
does not guarantee that the trophic status of the receiving lake will be sig-
nificantly improved. That determination can only be made on a case-by-case
basis in which many factors, such as background phosphorus levels, the limit-
ing nutrient, lake morphometry, etc., are considered. It is apparent, however,
that in many cases, eutrophic conditions are either the direct result of phos-
phorus from municipal wastes or at least are worsened by phosphorus inputs from
these sources which could be readily controlled.
19]
-------
TROPHIC CONDITION OF SURVEY LAKES
About 80% of the lakes and reservoirs included in the first two years
of the Survey in the eastern United States were eutrophic. This was not
unexpected since a large number of these water bodies were impacted by
municipal wastes.
The classical terms, oligotrophic, mesotrophic, and eutrophic were used
to describe the trophic condition of each water body. Based partly on ob-
servations during the first year of the Survey and partly on literature val-
ues, some general guidelines were developed for each of four key parameters
to assist us in assigning a trophic classification to each lake. These
values are listed in Table 3.
TABLE 3. KEY PARAMETER VALUES ASSOCIATED HITH
THREE LAKE TROPHIC CONDITIONS
Parameter
Oligotrophic Mesotrophic Eutrophic
Total Phosphorus (yg/1)
Chlorophyll a (yg/1)
Secchi depth (meters)
Hypolimnetic Dissolved Oxygen
(% saturation)
<10
<4
>3.7
>80
10-20
4-10
2.0-3.7
10-80
>20-25
>10
<2.0
<10
If each of the four parameters from a given lake were within the range
of a specific trophic condition (e.g., oligotrophic) then it was fairly cer-
tain that the indicated trophic condition appropriately described the lake.
Unfortunately, in many cases, all the parameter values did not neatly fall
within one trophic classification; therefore, a relative index or ranking
system was also used. This index included the four parameters shown in Table
3 (except that minimum dissolved oxygen concentrations were used) plus inor-
ganic nitrogen and dissolved orthophosphorus concentrations. The index was
based on percentile rankings for each of the six parameters which were then
added together to produce a single index number. Using this system, a large
number of lakes could be ranked in general order from most oligotrophic to
most eutrophic. There were enough well-studied lakes included in the Survey
to allow us to determine approximately where the transition from oligotrophic
to mesotrophic and from mesotrophic to eutrophic occurred in the ordered list
of lakes. This system was not without exception but did prove useful. The
index is discussed in detail in National Eutrophication Survey Working Paper
No. 24 (1974).
PHOSPHORUS LOADING - TROPHIC CONDITION RELATIONSHIPS
Another of the Survey objectives was to estimate annual phosphorus and
nitrogen loadings for each of the study lakes and to examine relationships
between these nutrient inputs and the resulting trophic conditions. Such
relationships are needed by lake managers to predict trophic responses which
would result from either increasing or decreasing phosphorus loads. They
192
-------
would also give regulatory agencies a firmer basis for allocating total
phosphorus loads from point or non-point sources so that the desired tro-
phic condition of a lake or reservoir could be maintained or achieved.
The Survey has not developed any original nutrient loading-lake res-
ponse relationships. However, the data have been applied to models recently
developed by other investigators.
Prior to 1968 there were no models of general applicability which re-
lated total phosphorus load to trophic condition in the receiving lake. Now,
however, there are at least three which seem very promising. These models
are presented and compared using data collected by the Survey from twenty-
three lakes and reservoirs. These twenty-three water bodies represent a
cross-section of trophic conditions, mean depths, and mean hydraulic reten-
tion times. All are located in northeastern and north-central states except
for two reservoirs in Georgia and two in South Carolina. In this group of
lakes, six are oligotrophic, nine are mesotrophic, and eight are eutrophic.
The three relationships (or models) which will be compared were devel-
oped by Vollenweider and Dillon (1974), Dillon (1975), and Larsen and Mercier
(1975), respectively.
Vollenweider (1968), using existing data from a number of European and
North American lakes, was the first to relate total phosphorus loading to
lake trophic condition. He plotted annual total phosphorus loadings (g/mVyr)
against lake mean depths and empirically determined the transition between
oligotrophic, mesotrophic, and eutrophic loadings.
Although this approach worked reasonably well for lakes with detention
times of several months or longer, it did not account for the fact that two
lakes with identical mean depths could have quite different hydraulic reten-
tion times and therefore different trophic responses to the same loading rate.
Subsequently, Vollenweider modified his initial relationship and based his
revised model on considerations of a mass balance equation for phosphorus.
The application of Vollenweider's revised model to the Survey lakes is illus-
trated in Figure 4.
The observed loadings and trophic conditions of the 23 Survey lakes did
not fit the Vollenweider relationship very well. Phosphorus loadings for
five of the eutrophic lakes plotted clearly within the eutrophic zone of the
Vollenweider relationship while loadings of two eutrophic lakes plotted with-
in the mesotrophic zone and one within the oligotrophic zone. Loadings for
five of the mesotrophic lakes fell within the oligotrophic zone while the
remainder were within the mesotrophic portion of the Vollenweider relationship,
Vollenweider's work was extremely important not only because he was
the first to investigate the loading-response relationship but also because
his original ideas interested others in this type of approach.
193
-------
— 10.0
k.
^
CM
E
J9
UJ
i—
cc
o 1.0
z
Q
3
__l
CO
ID
ir
o
X
n_
cn O.I
CL
_|
o
n ni
1 1 1 1 1 INI 1 1 M II III I 1 1 1 1 Mil 1 1 1 II
D
4515
— —
— —
~ ..DANGEROUS ~
"EUTROPHIC" /
- 4512 S -
D // ^PERMISSIBLE
/
D 27A8 .S' /
// /
// /
- /X^A1316 S -
~ ^^ xA^seoe ~
^ / s ' —
- 3639 ^^^701 x^- O 2314
_ 2750 __^-^ ^ D /AI3I8 -
- L(o)=0-30 0. — -"-^^^- ' ^^/
i - n ?n • — — " — ' D 2618 -^" ^*" ^
, ,° m= """^ ^-A27B4 "OLIGOTROPHIC"
L(°)'°-15 ^--A A 2306
—3303 O A 2313
- L(0) = O.IO - 27B2A A36I7 =
- 2694 O D 5539 ^\\
^1 A 2696 O Oligotrophic Lakes
O 2695 ^ Mesotrophic Lakes
_ D Eutrophic Lakes
i i i i i i I n I i i i i i 1 1 1 i i i i i i i n I i i i I
O.I 1.0 10.0 100.0
MEAN DEPTH (m) / MEAN HYDRAULIC RETENTION TIME (yrs.)
Figure 4. The Vollenweider relationship applied to a number of eastern U.S.
lakes and reservoirs sampled by the Survey.
-------
Stimulated by Vollenweider's earlier work, Dillon (1975) used the
mass balance modeling approach to derive the relationship illustrated in
Figure 5. Dillon's approach relates lake mean depth to a factor which
includes total annual phosphorus loading, the phosphorus retention coef-
ficient, and hydraulic flushing time. The 23 Survey lakes fit the Dillon
relationship quite well as illustrated in Figure 5. Two oligotrophic lakes
plotted in the mesotrophic zone and two mesotrophic lakes plotted in the
oligotrophic zone; however, observed conditions for the other lakes were
as predicted by the Dillon relationship.
"EUTROPHIC"
D 27A8
D 2750
E
\
o>
tr
i
D 2618
D3639
O.I -
D27CI
045I2/ A 2313^ °23"
02695
A'27B4 / °23°9
^2762 / 03303
A 2306
AI3I8
"OLIGOTROPHIC"
0.01
O Oligotrophic Lakes
A Mesotrophic Lakes
D Eutrophic Lakes
J_
10.0
MEAN DEPTH ( METERS)
100.0
Figure 5. The Dillon relationship applied to a number of eastern U.S.
lakes and reservoirs sampled by the Survey.
195
-------
Larsen and Mercier (1975), working independently of Dillon, also
solved a mass balance equation for phosphorus to develop a relationship
between the average incoming phosphorus concentration and the phosphorus
retention coefficient. The average incoming phosphorus concentration is
defined as the total annual phosphorus load divided by the total hydraulic
inflow which is also equivalent to:
where, L = annual total phosphorus
P = hydraulic flushing time
2 = mean depth (meters)
areal load (g/m /yr)
(exchange/year)
The Larsen and Mercier relationship therefore incorporates the same
variables as the Dillon relationship although the graphical solution of
the mass balance model for phosphorus is different. Figure 6 depicts the
23 Survey lakes plotted against the Larson-Mercier relationship. The fit
is very good and the relative location of each point on the graph is very
similar to Figure 5, the Dillon relationship.
o
oc
UJ
o
o IOO.O
CO
z>
oc
o
X ~
ft-
O o»
10.0
o
o
1.0
D 2750
"EUTROPHIC"
D 27A8
"MESOTROPHIC"
O 2314
D 4515
D 27CI^ A 2696
"AO-7Q/1 2313
A27B4 A |3i6
1318
2309
A
2306^02311 "OLIGOTROPHIC"
O Oligotrophic Lakes
A Mesotrophic Lakes
D Eutrophic Lakes
J_
_L
_L
_L
_L
_L
_L
0.0 O.I 0.2 0.3 0.4 0.5 0.6 0.7 0.8 0.9 1.0
PHOSPHORUS RETENTION COEFFICIENT ( RFyp)
EXPJ
Figure 6.
The Larsen-Mercier relationship applied to a number of
eastern U.S. lakes and reservoirs sampled by the Survey.
196
-------
Since both the latter two models predict in-lake concentrations of
total phosphorus, the vertical distance from an observed point, repre-
senting a lake, to one of the transitional lines is at least a semi-
quantitative measure of the degree of oligotrophy or eutrophy.
The vertical distance from a given point to a transitional line in
the Vollenweider relationship has less meaning in terms of the degree
of oligotrophy or eutrophy because the model does not directly relate
total phosphorus loading to in-lake phosphorus concentrations.
In summary, the models developed by Dillon and Larsen-Mercier, which
relate total phosphorus loads to lake phosphorus concentrations, should
prove to be useful lake management tools. The Vollenweider model, at this
time, is probably less precise because it considers only total phosphorus
loading without regard to in-lake processes which reduce the effective
phosphorus concentration; however, the model can be used to determine
approximate acceptable total phosphorus loads.
THE RELATIONSHIPS OF LAND USE TO NUTRIENT LEVELS
Another of the Survey objectives is to examine, on a National scale,
the relationships of land use and other draining area characteristics to
stream nutrient levels and subsequently lake trophic status.
Of the 4,200 sub-drainage areas sampled by the Survey across the United
States, about 1,000 were selected for a detailed study of land use and other
drainage area characteristics (see Figure 7). Criteria for selecting the
1,000 stream sampling sites and associated drainage areas were:
1. Absence of identifiable point sources.
2. Availability of usable aerial photography (scale 1:40,000 to
1:80,000) or exising land-use data.
3. Availability of accurate topographic maps for drainage area
delineation.
4. Sufficient land relief for clear delineation of drainage area
1 i mi ts.
5. The need to encompass a variety of geographic and climatic areas.
Note that few, if any, of the selected drainage areas were in Florida,
the Atlantic and Gulf coastal plains, or northern Minnesota. These areas
were excluded from consideration because of the difficulty of accurately
defining drainage area boundaries due to low topographic relief, and, in
many cases, because of the strong influence of ground water.
At the present time only the data from the eastern United States (east
of the Mississippi River) have been compiled, but the analysis of these data
is not complete. Therefore only general results are presented.
197
-------
00
DISTRIBUTION OF
N. E. S. LAND USE
STUDY DRAINAGE AREAS
Each ol the 1002 dels
reprrM'nl a Inbuiciry sampling
sile and ili associated
drainage area
72 73 and 74 refer lo
ihe yeari Inbulrjry sampling
began in each group of
Figure 7. The distribution of stream drainage areas selected by the
National Eutrophication Survey for land use studies.
-------
Figure 8 summarizes the data collected from 473 eastern U.S. drainage
areas for total phosphorus and total nitrogen concentrations originating
from different land use categories. The categories are defined as follows:
1. Forest; other types negligible
a. >75% forested (including forested wetland)
b. < 7% agriculture
c. < 2% urban
2. Mostly forest; other types present
a. >50% forest
b. not included in the forest category
3. Mostly agriculture; other types present
a. >50% agriculture
b. not included in the agriculture category
4. Agriculture; other types negligible
a. >75% agriculture
b < 7% urban
5. Urban
>39% urban
6. Mixed; not included in any of the other categories
Streams draining predominately agricultural areas have total phosphorus
concentrations averaging about 10 times higher than those draining forested
areas (Figure 8). The difference between total nitrogen concentrations was
not as marked. Streams in agricultural areas averaged nearly 5 times higher
total nitrogen concentrations than those draining forested areas. It is in-
teresting to note that, based on the mean concentration values, phosphorus
would be expected to be limiting in surface waters draining either forested
or agricultural areas. The total nitrogen to total phosphorus ratio changes
from 60 to 1 for forested areas to 31 to 1 for agricultural areas. Gen-
erally phosphorus is the limiting nutrient when the N:P ratio exceeds 14:1.
The nutrient loads per unit area of drainage for total phosphorus and
total nitrogen are shown in Figure 9. The differences in exports for the
different land use categories are not as pronounced as the nutrient concen-
trations were. Total phosphorus export from agricultural lands was only
317 times greater than from forested lands and total nitrogen export only
2.2 times greater. The differences in magnitude between stream loads and
stream concentrations are due to the differences in stream flows resulting
from the two types of land use. The data suggest that stream flow per unit
of drainage area is somewhat higher for forested than for agricultural areas
This seems logical since forested areas frequently are those which are un-
suitable for agricultural purposes because of steeper slopes and relatively
thin soils.
-------
NUMBER
OF SUBS
53 FOREST
eihur ryp0si
170 MOSTLY FOREST
oihtr tfpfi pntfl
52 MIXED
11 MOSTLY URBAN
96 MOSTLY AGRIC.
orttfr ryp« freifnt
91 AGRICULTURE
olh»r rypn rmfligibt*
MEAN TOTAL PHOSPHORUS CONCENTRATIONS
vs
LAND USE
DATA ON 473 SUBDRAINAGE AREAS IN
EASTERN UNITED STATES
0.014
0.035
0.040
0.066
°-135
0.05
MILLIGRAMS PER LITER
0.10
0.15
ro
o
o
NUMBER
OF SUBS
53 FOREST
olhei lypei negligible
170 MOSTLY FOREST
52 MIXED
11 MOSTLY URBAN
other types present
96 MOSTLY AGRIC.
'
91 AGRICULTURE
MEAN TOTAL NITROGEN CONCENTRATIONS
vs
LAND USE
DATA ON 473 SUBDRAINAGE AREAS IN
EASTERN UNITED STATES
1.O
2.0
MIILIGRAMS PER LITER
4.170
3.0
4.0
Figure 8. The relationship between total phosphorus and total nitrogen
concentrations in streams and land use in the eastern U.S.
-------
NUMBER
OF SUBS
53 FOREST
OTAvr OPMfi
170 MOSTLY FOREST
52 MIXED
11 MOSTLY URBAN
M/Mr npn pruww
96 MOSTLY AGRIC.
Otter trt*3 pnsmm
91 AGRICULTURE
TOTAL PHOSPHORUS EXPORT
vs
LAND USE
DATA ON 473 SUBDRAINAGE AREAS IN
EASTERN UNITED STATES
8.3
30.8
10 20
KILOGRAMS PER SQUARE KILOMETER PER YEAR
30
40
ro
o
NUMBER
OF SUBS
53 FOREST
170 MOSTLY FOREST
otbmt type* prm**nt
52 MIXED
11 MOSTLY URBAN
0f/w tyr** pmmr
96 MOSTLY AGRIC.
or/Mi- W»» prnmt
91 AGRICULTURE
«A«Y n^P«i rmgltgibl*
TOTAL NITROGEN EXPORT
vs
LAND USE
DATA ON 473 SUBDRAINAGE AREAS IN
EASTERN UNITED STATES
788.6
630.5
J 982-3
500
KILOGRAMS PER SQUARE KILOMETER PER YEAR
1000
Figure 9. The relationship between total phosphorus and total nitrogen export
in streams and land use in the eastern U.S.
-------
The pattern for orthophosphorus concentrations was very similar to
that for total phosphorus as shown in Figure 10. Except with predomin-
ately urban drainage areas, of which there were only eleven, mean ortho-
phosphorus concentrations represented 40 to 43% of the total phosphorus
concentrations regardless of overall land use. Orthophosphorus concen-
trations in streams draining agricultural areas were nearly 10 times the
concentrations in streams draining forested areas.
Inorganic nitrogen exhibited quite a different pattern from total
nitrogen in that substantially higher (13.7X) concentrations were observed
in streams draining agricultural lands than in forested lands (Figure 10).
In streams draining forested areas, inorganic nitrogen constituted
about 27% of the total nitrogen, however, this increased to 76% in streams
draining predominately agricultural areas. Although the sample size (11
drainage areas) was relatively small, inorganic nitrogen made up about 98%
of the total nitrogen in streams draining mostly urban drainage areas. In-
organic nitrogen export was also significantly higher (5.6X) from agricul-
tural areas than from forested areas as shown in Figure 11. The difference
probably reflects the use of inorganic nitrogen fertilizers and the high
water solubility of inorganic nitrogen compounds.
What conclusions can be drawn from these general results? First,
these data suggest that streams draining agricultural watersheds have
higher nutrient levels and therefore would be expected to be more produc-
tive than those draining forested watersheds. The increase in nutrient
levels is generally proportional to the increasing percent of the land in
agriculture.
Second, the data indicate that the inorganic portion (orthophosphorus)
of the total phosphorus component stays roughly at the 40% level regardless
of land use type, whereas, the inorganic portion of the total nitrogen com-
ponent increases markedly from 27% for forested areas to 75% for agricul-
tural areas. Inorganic nitrogen in streams draining mostly urban areas
represented a substantially larger fraction of the total nitrogen (98%),
however, the number of test areas was relatively small (11).
Lastly, what uses can be made of the data derived from this segment of
the survey? Other than elucidating the land use-nutrient level-eutrophi-
cation relationships, probably the two most important uses will be: (1) to
provide a basis for a quick and relatively accurate method of determining
nitrogen and phosphorus concentrations and loadings based on land use and
other non-point source types of geographical characteristics, and (2) to
provide a large nationwide collection of watershed data for testing other
methods of estimating nitrogen and phosphorus levels in streams from non-
point sources.
SUMMARY
The National Eutrophication Survey, which was initiated in 1972 by the
U.S. Environmental Protection Agency, is in the first stage of collecting
data from over 800 lakes and reservoirs in the contiguous United States.
In the eastern U.S., a large percentage of the surveyed water bodies are
impacted by municipal sewage treatment plant effluent and are in various
202
-------
NUMBER
OF SUBS
53 FOREST
170 MOSTLY FOREST
Other types present
52 MIXED
11 MOSTLY URBAN
other tvoe* present
96 MOSTLY A6RIC.
other ffpes. pie tent
91 AGRICULTURE
Other types negligible
MEAN ORTHOPHOSPHORUS CONCENTRATIONS
vs
LAND USE
o.oi
DATA ON 473 SUBDRAINAGE AREAS IN
EASTERN UNITED STATES
0.033
0.027
O.OS8
0.02
0.03 0.04
MILLIGRAMS PER LITER
0.05
0.06
no
o
oo
NUMBER
OF SUBS
53 FOREST
170 MOSTLY FOREST
other types present
52 MIXED
11 MOSTLY URBAN
96 MOSTLY AGRIC.
othei fype* present
91 AGRICULTURE
other iype$ negligible
MEAN INORGANIC NITROGEN CONCENTRATIONS
vs
LAND USE
DATA ON 473 SUBDRAINAGE AREAS IN
EASTERN UNITED STATES
J 0.67»
dfaZl J 3-190
0.50
1.00
1.50 2.00
MILLIGRAMS PER LITER
2.50
3.00
Figure 10. The relationship between orthophosphorus and inorganic nitrogen
concentrations in streams and land use in the eastern U.S.
-------
NUMBER
OF SUBS
53 FOREST
other types negligible
170 MOSTLY FOREST
52 MIXED
11 MOSTLY URBAN
96 MOSTLY AGRIC.
91 AGRICULTURE
other rrpes negligible
ORTHOPHOSPHORUS EXPORT
vs
LAND USE
DATA ON 473 SUBDRAINAGE AREAS IN
EASTERN UNITED STATES
T ™".
S 10
KILOGRAMS PER SQUARE KILOMETER PER YEAR
15
ro
o
NUMBER
OF SUBS
53 FOREST
other types negligible
170 MOSTLY FOREST
52 MIXED
11 MOSTLY URBAN
other types present
96 MOSTLY AGRIC.
other types present
91 AGRICULTURE
other types legligible
INORGANIC NITROGEN EXPORT
vs
LAND USE
DATA ON 473 SUBDRAINAGE AREAS IN
EASTERN UNITED STATES
200
400 600
KILOGRAMS PER SQUARE KILOMETER PER YEAR
800
Figure 11. The relationship between orthorphosphorus and inorganic nitrogen
export in streams and land use in the eastern U.S.
-------
states of enrichment. Phosphorus loads to a significant number of these
impacted lakes and reservoirs could be substantially reduced by controlling
phosphorus inputs from municipal sources.
Primary production in 67% of the water bodies surveyed east of the Rocky
Mountains was phosphorus-limited and 30% were nitrogen-limited according to
algal assay results. It is believed that the apparent nitrogen-limited con-
dition was frequently the result of excessive phosphorus inputs from munici-
pal sources.
Land use in the watershed was shown to be a significant factor in de-
termining levels of phosphorus and nitrogen in streams in selected areas
studied in the eastern United States. Average total phosphorus concentra-
tions were about 10 times greater in streams draining agricultural areas
than in streams draining forested areas; total nitrogen concentrations were
about 5 times greater. The percentage of total nitrogen in the inorganic
form was substantially higher in streams draining agricultural lands than in
those streams draining forested lands.
Phosphorus loading data for 23 selected survey lakes were applied to
three general models relating annual total phosphorus loading rates to lake
trophic conditions. The "fit" of observed conditions to predictions made by
each model was compared and discussed.
REFERENCES
Dillon, P. J. 1975. The phosphorus budget of Cameron Lake, Ontario:
the importance of flushing rate to the degree of eutrophy of lakes.
Limnol. Oceanogr. 20:28-39.
Larsen, D. P. and H. T. Mercier. 1975. Lake phosphorus loading graphs:
an alternative. National Eutrophication Survey Working Paper No.
174.
U.S. Environmental Protection Agency. 1974. Survey methods for lakes
sampled in 1972. National Eutrophication Survey Working Paper No. 1.
40 pp.
U.S. Environmental Protection Agency. 1975. Survey methods, 1973-1976.
National Eutrophication Survey Working Paper No. 175. 91 pp.
U.S. Environmental Protection Agency. 1974. An approach to a relative
trophic index system for classifying lakes and reservoirs. National
Eutrophication Survey Working Paper No. 24. 44 pp.
Vollenweider, R. A. 1968. The scientific basis of lake and stream
eutrophication with particular reference to phosphorus and nitrogen
as factors in eutrophication. OECD, DAS/CSI/68-27. 159 pp.
Vollenweider, R. A. and P. J. Dillon. 1974. The application of the
phosphorus loading concept to eutrophication research. National
Research Council of Canada No. 13690. 42 pp.
205
-------
PAPERS PRESENTED BUT NOT AVAILABLE FOR PUBLICATION
A Method of Predicting Bioaccumulation Potential of Chemicals
Oilman Veith. Environmental Research Laboratory,
U.S. Environmental Protection Agency, Duluth, MN
Adverse Effects of Chlorine Disinfection on Aquatic Organisms
William A. Brungs. Environmental Research Laboratory,
U.S. Environmental Protection Agency, Duluth, MN
Z06
-------
TECHNICAL REPORT DATA
(Please read Instructions on the reverse before completing)
REPORT NO.
EPA-600/3-76-079
3. RECIPIENT'S ACCESSION-NO.
.TITLE AND SUBTITLE WA I h R QUALITY CRITERIA RESEARCH OF THE
.S. ENVIRONMENTAL PROTECTION AGENCY Proceedings of an
PA-sponsored Symposium on Marine, Estuarine and Fresh-
5. REPORT DATE
July 1976
6. PERFORMING ORGANIZATION CODE
?g^ented at the 26th annua1 meeting of
. AUTHOR(S)
8. PERFORMING ORGANIZATION REPORT NO.
nvironmental Protection Agency
. PERFORMING ORGANIZATION NAME AND ADDRESS
Corvallis Environmental Research Laboratory
U.S. Environmental Protection Agency
200 S. W. 35th Street
Con/all is, OR 97330
10. PROGRAM ELEMENT NO.
1BA608
1 1. CONTRACT/GRANT NO.
2. SPONSORING AGENCY NAME AND ADDRESS
Corvallis Environmental Research Laboratory
Dffice of Research and Development
U.S. Environmental Protection Agency
Con/all is, OR 97330
13. TYPE OF REPORT AND PERIOD COVERED
14. SPONSORING AGENCY CODE
EPA-ORD
15. SUPPLEMENTARY NOTES
16. ABSTRACT
These proceedings include a cross-sectional representation of the broad base
ecological effects research programs conducted by research laboratories of the EPA
Office of Health and Ecological Effects. The presentations focus on microbial and
abiotic degredation processes, the problem of trace metals, the effects of toxic or-
ganics, and the feasibility of new stress-measuring methodologies in the marine environ
ment. The freshwater segment of the symposium addresses the transport and biological
modeling capabilities of the laboratories, cold climate aquatic biology, lake trophic
states in the eastern United States, and the impact of toxic substances on freshwater
systems.
7.
KEY WORDS AND DOCUMENT ANALYSIS
DESCRIPTORS
Bioassimilatior
b.IDENTIFIERS/OPEN ENDED TERMS
COSATl Field/Group
:reshwater
Marine
Estuarine
hosphorus
Nitrogen
race Metals
Cold Climate
Stream Flow
Toxicity
Bioassay
Communities
Microbiota
Chlorine
Ecosystem Mode
Malathion
Lake Restorati
Advanced Waste
Treatment
Great Lakes
Water Quality Criteria
06/F
08/A,H,J
Ecosystem Models Phytoplankton
18. DISTRIBUTION STATEMENT
Release to Public
19. SECURITY CLASS (This Report)
Unclassified
21. NO. OF PAGES
20. SECURITY CLASS (This page)
Unclassified
22. PRICE
EPA Form 2220-1 (9-73)
207
it U.S. GOVERNMENT PRINTING OFFICE: 1976—697-627)104 REGION 10
------- |