&EPA
United States
Environmental Protection
Agency
Environmental Research
Laboratory
Corvallis OR 97333
EPA 600/3-82-034
April 1982
Re-search and Development
Second US/USSR
Symposium:
Biological Aspects of
Pollutant Effects on
Marine Organisms
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EPA-600/3-82-034
SECOND US/USSR SYMPOSIUM
BIOLOGICAL ASPECTS OF POLLUTANT EFFECTS ON MARINE ORGANISMS
Terskol, USSR
June 4 to 9, 1979
Symposium sponsored as part of the US/USSR
Agreement on Protection of the Environment
compiled by
D. J. Baumgartner
Chairman, US Delegation
and
A. I. Simonov
Chairman, USSR Delegation
ENVIRONMENTAL RESEARCH LABORATORY
OFFICE OF RESEARCH AND DEVELOPMENT
U.S. ENVIRONMENTAL PROTECTION AGENCY
CORVALLIS, OREGON 97333
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DISCLAIMER
This report has been reviewed by the Con/all is Environmental Research
Laboratory, U.S. Environmental Protection Agency, and approved for publica-
tion. Approval does not signify that the contents necessarily reflect the
views and policies of the U.S. Environmental Protection Agency, nor does
mention of trade names or commercial products constitute endorsement or recom-
mendation for use.
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ABSTRACT
This Symposium was conducted under a US/USSR Environmental Agreement,
Project 02.06-21 titled "Effect of Pollutants on Marine Organisms." Papers by
American and Soviet specialists present advances in hydrobiological analysis
of basic structural components of marine ecosystems and the influence of
various pollutants on these components. Results of laboratory research as
well as field observations on the influence of pollution on the marine
environment are presented.
Participants attending the Symposium discussed problems related to
methods for modeling the influence of pollutants on the marine environment,
long-term forecasting and determination of permissible loads of pollutants,
and the unification and intercalibration of methods for determining biological
responses and chemical contamination.
The Protocol of the Symposium (the official summary) which follows refers
to an introduction to this report. The introductory material to provide
background information about technical discussions which took place at the
Symposium but which did not result in published reports is included in this
abstract.
Extensive discussions were held on the scope of the technical topics to
be included in the exchange. We recognized that EPA's interests were
primarily related to the scientific basis for regulating waste discharges,
whereas the Soviet scientists tended to be less involved in this activity and
more interested in global or at least large-scale marine water quality for its
own sake. The United States co-chairman proposed to invite greater participa-
tion from the National Oceanic and Atmospheric Administration (NOSS) in future
project activities to balance the interest of the Soviet scientists. In
another way the scope of the exchange was consolidated rather than expanded.
This was in relation to the development of two series of symposia — one for
biological effects, the other for "Chemical Pollution of the Marine Environ-
ment." There was one proceedings published in the latter series. The
participants embraced the concept that biological effects could not be
discussed realistically without a consideration of chemical and physical
factors, and indeed the papers presented in this biological symposium are not
exclusively biological in content. Consequently, it was agreed to abandon the
chemical symposium series in favor of one series devoted to an appropriately
broad range of scientific disciplines. The next symposium, therefore, to be
held in the United States will be labeled simply as the Fourth American-Soviet
Symposium on the Effects of Pollution on Marine Organisms.
Proceedings are published in English and Russian in compliance with the
Memorandum from the 4th Session of the Joint US-USSR Committee on Cooperation
in the Field of Environmental Research.
iii
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CONTENTS
Page
Modeling Effects of Environmental Perturbations
on Marine Plankton Dynamics 1
Bruce W. Frost
A Quantitative Method for Evaluating External Effects on Ecosystems .... 15
V. D. Fedorov, V. N. Maksimov and V. B. Sakharov
Petroleum Pollution and Microflora in Marine Ecosystems 26
M. V. Gusev, T. V. Koronelli and V. V. Il'inskiy
Development of Plankton Algae in Conditions of Petroleum Pollution 43
0. G. Mironov
Long-Term Biological Variability and Stress in Coastal Systems 52
Robert J. Livingston
Some Features of the Biologic Effects of Pollutants on Marine Organisms . . 67
S. A. Patin
Fish as an Object for Monitoring Petroleum Pollution
of the Marine Environment 71
N. D. Mazmanidi, G. I. Kovaleva and A. M. Kotov
Improved Assessment of Ecological Effects by Incorporating Physical
Simulations in Bioassay Procedures 82
D. J. Baumgartner
A Program for Investigating Pollution of Marine Macrobenthos by Heavy
Metals and Its Position in the System for Monitoring the Environment. ... 98
K. S. Burdin and I. B. Savel'yev
Effects of Liquid Industrial Wastes on Estuarine Algae, Plants,
Crustaceans, and Fishes 112
Gerald E. Walsh and Richard L. Garnas
Ecologic Aspects of Using Chemical Agents for Eliminating the Results
of Oil Spills in the Ocean 123
M. P. Nesterova
Chemical Pollution of the Film Layer of the Pacific Ocean 131
A. I. Simonov and V. I. Mikhaylov
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CONTENTS (continued)
Studies on the Biological Transport of Materials From Surface to Deep
Ocean Waters: I. Fluxes of Carbon, Nitrogen, and Phosphorus
II. Fluxes of Trace Elements 145
G. A. Knauer and J. H. Martin
Correlation Between Dose Loads in Fish and the Biogeochemistry
of Artificial Radionuclides in a Marine Environment 166
I. A. Shekhanova and A. P. Panarin
The Consequences of Chemical Pollution of the "Waterbottom Sediment"
Contact Zone in the Sea 180
A. Bronfman and Z. B. Aleksandrova
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PARTICIPANTS
American Side
Dr. Donald J. Baumgartner
Co-chairman of the Symposium
Environmental Research Laboratory
U.S. Environmental Protection Agency
Corvallis, Oregon
Dr. Bruce W. Frost
Department of Oceanography
University of Washington
Seattle, Washington
Dr. George A. Knauer
Moss Landing Marine Laboratories
Moss Landing, California
Dr. Robert J. Livingston
Department of Biological Science
Florida State University
Tallahassee, Florida
Dr. Gerald E. Walsh
U.S. Environmental Protection Agency
Environmental Research Laboratory
Gulf Breeze, Florida
Soviet Side
Dr. Anatoliy I. Simonov
Co-chairman of the Symposium
State Oceanographic Institute
Moscow
M. P. Nesterova
Deputy Co-chairman of the Symposium
P.P. Shershov Oceanology Institute
USSR Academy of Sciences
Dr. Vadim D. Fedorov
Head of the Chair of Hydrobiology
Moscow State University
Dr. Mikhail V. Gusev
Dean, Department of Biology
Moscow State University
0. G. Mironov
Institute of Biology
for Southern Seas
AN USSR, Sevastopol'
S. A. Patin
All-Union Fisheries and Oceanograpy
Scientific Research Institute
Moscow
N. D. Mazmanidi
Georgian Branch of the VNIRO
K. S. Burdin, Decent
Department of Biology
Moscow State University
V. I. Mikhaylov
State Oceanographic Institute
Odessa Branch
Irina A. Shekhanova
Head of the Radiation
Biology Laboratory
Al 1-Union Scientific-Research
Institute of Marine Fisheries
and Oceanography
Moscow
A. I. Bronfman
Economics Institute of the
Ukrainian Academy of Sciences
Odessa
Vll
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Dr. Ralfrid A. Gasanov
Institute of Botany
Academy of Sciences of
Azerbaijan S.S.R.
Baku
Dr. Gasanov presented a paper entitled "Biophys-
ical Parameters of Photosynthesizing Organisms in
a System for Monitoring Marine Pollution." His
co-authors on the paper were N. M. Aliyev, Z. Sh.
Aliyev, and N. M. Karayeva. The paper was not
available for publication in this volume of the
proceedings.
vm
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PROTOCOL OF THE SECOND US/USSR SYMPOSIUM
BIOLOGICAL ASPECTS OF POLLUTANT EFFECTS ON MARINE ORGANISMS
In accordance with the principles outlined in the Memorandum of the VII
Session of the Joint US/USSR Committee Cooperation in the Field of Environ-
mental Protection (January 29 - February 2, 1971, Moscow, USSR), the Second
US/USSR Symposium on Biological Aspects of Pollutant Effects on Marine Organ-
isms was held in settlement Terskol, June 4-9, 1979.
Dr. D. J. Baumgartner, Director, Marine Division, Corvallis Environmental
Research Laboratory, U.S. Environmental Protection Agency, and Prof. A. I.
Simonov, Chief of Department of the State Oceanographic Institute (USSR), were
co-chairmen of the Symposium.
The participants of the Symposium presented reports on modeling the
effects of pollutants on ecological systems, marine plankton dynamics influ-
enced by natural and anthropogenic factors, microflora and planktonic algae of
marine ecosystems containing petroleum hydrocarbons, marine toxicological
bioassays, the use of models to incorporate physical conditions in bioassays,
relationships of dose loads in fish with radionuclide biogeochemistry in the
marine water, ecological aspects of utilization of chemical means for the
elimination of oil spills in the sea, oceanic surface layer chemical pollu-
tion, significance of chemical interactions in the seawater-seabed interface,
biological consequences of marine environmental chemical pollution including
effects on marine macrobenthos pollution by heavy metals, fish contamination
by petroleum hydrocarbons, and application of biophysical parameters of photo-
synthesizing organisms while monitoring.
The participants expressed their wish to hold the Third Symposium on
biological aspects of pollution effects on marine organisms and marine pollu-
tion dynamics in 1-2 years in the United States after conducting joint and
independent studies. Problems concerned with estuaries and the continental
shelf in the selected regions of the world's oceans would receive attention in
accordance with the program of long-term research designed at the joint
US/USSR meeting in September 1978 (Gulf Breeze, Moss Landing, Corvallis, USA).
The number of participants of the USSR delegation to the Symposium, date and
place will be agreed upon by Co-chairmen of the Project by correspondence by
October 1979 and submitted for the consideration to the VII Session of the
US/USSR Joint Committee on Cooperation in the Field of Environmental
Protection.
-------
Plans were made for the next step in metal intercalibration studies and a
schedule was developed for two USSR scientists to visit the EPA marine
research laboratory at Narragansett, Rhode Island. This 2-week working visit
will begin July 1, 1979 for the purpose of exchanging information on methods
to analyze oil pollutants in seawater, sediments, and marine animal tissues.
Both sides agreed to publish the Symposium proceedings during 1980.
Co-chairmen agreed to write the introduction to the proceedings of the Second
Symposium by correspondence by October 1, 1979. They also agreed to send to
each side final texts of reports by July 1979. The USSR side will present its
reports in English and the United States reports will be presented in Russian.
Both sides consider it important to study biological effects of chemical
pollution in estuaries as well as to develop techniques of investigations and
design suitable predictive models. While developing techniques, one should
proceed both from the purposes and tasks of research. Apart from the tasks,
however, the sides believe it is significant to use known hydrodynamic models
to carry out complex observations that include both abiotic and biotic compo-
nents. The principal factors that should be included are those that influence
the ionic form of pollutants in rivers and those responsible for precipita-
tion, dissolution, and ionic speciation in seawater (salinity and other
chemical properties of seawater). Each specific estuary is likely to have a
characteristic species composition which may dictate the factors that need to
be included in a measurement program.
If the task is to study long-term changes of hydrobiological conditions,
then multiyear observations should be performed, similar to those demonstrated
in Dr. Livingston's paper, with subsequent statistical processing with the
help of systems analysis. If the task is to track the short-term changes due
to pollutant effects, then limited but systematic observations are required
when suitable biological monitors are selected, as it was shown in the paper
by Dr. K. Burdin and Dr. Savel'jev. In developing models to describe hydro-
biological condition changes caused by both natural and anthropogenic factors,
long-term observations should be combined with experimental laboratory data.
Both sides agreed that, in field assessments, it is recommended that
samples are to be taken both for chemical and biological analysis in the
surface water microlayer, thermocline layer, near bottom horizons, and in
bottom sediments, similar to procedures proposed in papers by Drs. A. Simonov
and V. Mikaylov; G. Knauer and J. Martin; M. Gusev, T. Koronelli and V.
Il'inskiy. Both sides consider it reasonable to conduct investigations in
coastal estuaries both in the USA and USSR and exchange results of investi-
gations. The estuaries for study, as well as methods for exchanging and
discussing the investigation results, will be agreed upon by Co-chairmen by
correspondence before the end of 1979. Research on the continued improvement
of bioassays, as discussed by Drs. Walsh and Baumgartner, was considered in
relationship to the importance of protecting estuarine and coastal ecosystems
from pollution stress.
-------
During the visit to the USSR, the American delegation visited Elbrus
scientific station of the Geographical Department of the Moscow State
University (settlement Terskol), Georgian All-Union Scientific-Research
Institute of Fishery and Oceanography in Batumi and the Biological Department
of the Tbilissi and Moscow State Universities.
The Symposium was held in the spirit of friendship, cooperation, and
mutual benefit. The United States delegation expresses gratitude to the
Soviet side for the excellent organization of the Symposium and for their
generous hospitality. The delegation thanks interpreters for their excellent
work.
The protocol was signed in settlement Terskol,
and Russian. Both copies are equally valid.
on June 9, 1979 in English
D. J. Baumgartner
Chairman of the Project
from the US side
A. I. Simonov
Chairman of the Project
from the USSR side
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MODELING EFFECTS OF ENVIRONMENTAL PERTURBATIONS
ON MARINE PLANKTON DYNAMICS
by
Bruce W. Frost
Department of Oceanography
University of Washington
Seattle, Washington
INTRODUCTION
Assemblages of marine planktonic organisms with varying degrees of trophic
complexity have been experimentally treated with pollutants such as heavy
metals, chlorinated hydrocarbons, and petroleum hydrocarbons. It has been
repeatedly observed that, because species differ in their tolerance to pol-
lutants, the major effect on assemblages is a shift in the species composition
of one or more trophic types, often without concomitant effects on standing
stock or rate measurements (Menzel 1977). Therefore, simulation models which
attempt to represent the effects of pollutants on plankton dynamics should
describe the structure of the plankton as well as biomass and production.
One such model of plankton dynamics was recently developed by Steele and
Frost (1977). The model simulates the structure and dynamics of phytoplankton
and herbivorous zooplankton during and after a spring bloom in a temperate
ocean. In the theoretical study described below, a reasonably realistic
simulation is characterized, then the model ecosystem is perturbed in a variety
of ways to simulate some effects of pollutants which have been documented in
the literature.
METHODS
A revised version of the simulation model of Steele and Frost (1977) was
used to examine some potential effects of pollutants on the structure and
dynamics of marine plankton communities. Briefly, the model simulates growth
of 20 size classes of phytoplankton and the population dynamics of two species
of filter-feeding copepods.
Except for change in the value of one parameter, the submodel specifying
phytoplankton growth remains as described by Steele and Frost (1977). That
is, growth of size classes of phytoplankton is determined by size-dependent
functions of nutrient uptake, respiration, sinking, and grazing. Variations
in size-specific patterns of growth of the phytoplankton were produced by
changing the value of the parameter a in the equation
R = aD~1/3 (1)
-------
where R is specific respiration rate (mg C/mg plant C per day) and D is dia-
meter (pro) of a phytoplankton cell. The parameter a was 0.36 in the standard
simulation.
Substantial revisions were made in the submodel of population growth of
planktonic grazers. The two species of grazers in the model are planktonic
copepods similar to Calanus pacificus (a close relative of C. helgolandicus)
and Pseudocalanus sp. These species have fundamentally different patterns of
optimal growth (Frost 1979). Therefore, separate submodels were used for the
two grazer species. All parameters of population growth for the two species
are identical to those used by Steele and Frost (1977) except the maximum
specific growth rate, G , which in the standard simulation takes the value
rn3.x
of 0.4 for Calanus and 0.2 for Pseudocalanus. The rationale for this change
is discussed by Frost (1979).
A second major change concerns the procedure of Steele and Frost (1977)
for describing the growth of copepods as passage of individuals through dis-
crete size (weight) classes. This procedure causes a severe problem of
"numerical dispersion" (Evans et al. 1977) and was therefore replaced by a
multiple cohort reproduction scheme (see Landry 1976) in which each day's
reproductive products are followed as a cohort until all members of the cohort
die. This gives a more realistic pattern of growth in the copepods.
Finally, the description of predation rate on the zooplankton species was
greatly simplified from that used by Steele and Frost (1977). It is assumed
that predation rate depends only on abundance of a grazer species
mZ
Predation Rate = (2)
H + Z
where m is the maximum predation rate (% per day), Z is population size
(numbers per m2) of a grazer species, and H is the population size at which
predation rate is half the maximum rate. A specific population of predators
is not simulated. The population dynamics of the two grazer species are
described separately in the model and the parameter H takes the value 100,000
per m2 for Calanus and 200,000 per m2 for Pseudocalanus. This is equivalent
to assuming that the two zooplankton species have different predators. The
parameter m is 0.1 in the standard simulation.
In all other respects, the format and parameter values of the simulation
model are identical to those described by Steele and Frost (1977).
RESULTS
Standard Simulation
A reasonably realistic simulation, the standard simulation, was selected
through trial and error by varying the coefficient for algal respiration, a
(eq. 1). The criterion for realism was that both Calanus and Pseudocalanus
maintained breeding populations for 100 days.
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The standard simulation (Figure 1A) predicts changes in concentration of
nitrate and standing stock of phytoplankton which are quite similar to those
of the basic run of Steele and Frost (1977: Figure 17). The major difference
is the relative abundance of the species of copepods (Figure IB). In this
revised model, Calanus dominates the zooplankton biomass during and after the
phytoplankton bloom, primarily because of its higher maximum specific growth
rate.
10
15
10
en
£
fO r-
£ 5
O
CP
0
A
o
CHLOROPHYLL
NITRATE u ^
40
80
C\J
E 2
o
o>
0
B
Pseudocalanus
0
40
80
DAYS
Figure 1. Simulated dynamics of plankton during and after a spring bloom in a
temperate ocean. This is the standard simulation. A, changes in
time of chlorophyll and nitrate. B, changes in time of biomass of
populations of Calanus and Pseudocalanus.
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The phytoplankton size composition (Figure 2) is strongly affected by grazing
and develops a bimodal distribution during the last 50 days of the simulation.
The larger mode in Figure 2 represents the sizes of cells with optimum intrin-
sic growth rates at the ambient nitrate concentrations, while the smaller mode
represents cells which have low intrinsic growth rates but are also very
inefficiently utilized by grazers.
100
4 8 16 32 64
CELL DIAMETER (pm)
128
Figure 2. Size composition of phytoplankton from day 20 to day 100 in the
standard simulation. Each point gives the concentration in a dis-
crete size category.
Population structures indicate that both species of copepods produce two
generations in the 100-day simulation (Figure 3). Adults of the first genera-
tion of Calanus recruit about 15 days before Pseudocalanus, which is due to
the
predominance
simulation.
of large phytoplankton cells during the first 60 days of the
Some very generalized effects of pollutants will now be examined by
altering, in turn, parameters which affect size composition of the phytoplank-
ton, growth and mortality in copepods, and abundance of higher trophic levels.
The simulated perturbations could, of course, occur at any time in the seasonal
cycle, yet to illustrate the basic patterns it is assumed that induced changes
in parameters persist for an entire simulation. Moreover, possible adaptations
of species to pollutants (Stockner and Antia 1976) are ignored.
Changes in Size Composition of Phytoplankton
A variety of pollutants cause shifts in the structure of the phyto-
plankton, generally toward predominance by very small cells (Menzel 1977;
-------
o
Q_
^
O
o
UJ
CD
CO
100
Co/onus
UJ
O
or
UJ
Q_
g
Q_
O
Q_
U_
O
z
O
50
0
0
50
100
Pseudoco/onus
100
50
0
EGGS-NH
0
50
DAYS
t
100
Figure 3. Percentage composition of populations of Calanus and Pseudocalanus
in the standard simulation. After day 15 the category copepods
includes adults (ad). The category nauplii includes naupliar stages
III-VI. Arrows indicate time of recruitment of adults of the first
generation.
-------
Greve and Parsons 1977). This effect may be simulated by changing the value
of the parameter a (eq. 1) which specifies the size dependence of respiration.
When the paramenter is decreased from 0.36 to 0.24, the phytoplankton size
composition rapidly shifts toward small cells (Figure 4). The cycle of stand-
ing stock of the phytoplankton (Figure 5A) is also severely disrupted in
comparison with the standard simulation. Phytoplankton become very abundant
late in the simulation because the predominant small cells are not efficiently
grazed by either species of grazer.
4 8 16 32 64
CELL DIAMETER (jjm)
128
20
Figure 4. Size composition of the phytoplankton from day 20 to day 80 when the
phytoplankton respiration coefficient a (eq. 1) is 0.24. ^
-------
15
ro
o I0
o>
E
i_
o
O
en
A
0
0
6r
40
80
CO
\
o 2
CP
0
B
Calanus
^ Pseudocalanus
40
80
DAYS
Figure 5. Changes in time of chlorophyll and nitrate (A) and zooplankton
biomass (B) when the phytoplankton respiration coefficient is 0.24.
The shift in phytoplankton composition to small cells tends to favor the
growth of the smaller species of grazer after the spring bloom (Figure 5B).
In this simulation, Calanus does not reproduce after day 40 and, although a
large population of copepods is produced in the first generation, they never
recruit to the adult stage. On the other hand, first generation adults of
Pseudocalanus recruit between days 35 and 40, and a second complete generation
is produced before the end of the 100-day simulation.
Should a pollutant favor growth of some other type or size range of
phytoplankton cells, the relative abundance of grazer species would change
depending upon the abilities of the grazers to utilize the predominant phyto-
plankton size classes. In turn, the growth rates of predators of the grazers
would also be affected (e.g., Koeller and Parsons 1977).
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Lethal and Sublethal Effects on Zooplankton
As shown previously, pollutants may indirectly but strongly affect the
zooplankton by causing changes in the composition of the phytoplankton. With
the model it is also possible to investigate more direct lethal and sublethal
effects of pollutants on zooplankton.
Some pollutants may cause total mortality of zooplankton species (e.g.,
Federle et al. 1979). If either Calanus or Pseudocalanus is eliminated from
the simulation there are concomitant large effects on the dynamics and
structure of the phytoplankton. With Calanus eliminated, the predicted phyto-
plank ton standing stock resembles Figure 5A, and the size composition of the
phytoplankton is dominated by small cells which are less effeciently grazed by
Calanus (Figure 6B).
Sublethal effects of pollutants on zooplankton may be manifested in
several ways. Observed sublethal effects include decreased rates of growth
and fecundity of copepods (Reeve et ah 1977; Moraitou-Apostolopoulou and
Verriopoulos 1979). This may be simulated by decreasing the maximum growth
rate of a species. Assume that Calanus is more sensitive than Pseudocalanus
to a particular pollutant. If the maximum growth rate, G_, . of Calanus is
max ^~^~~~~~
decreased from 0.4 to 0.2, that is, the same value as for Pseudocalanus, then
Pseudocalanus dominates during the phytoplankton bloom but subsequently the
composition of the zooplankton shifts twice in response to change in the size
composition of the phytoplankton (Figure 7). Early in the simulation (day
20), an abundance of all sizes of cells (Figure 7C) permits optimal feeding
and growth of both species. However, because Calanus matures more slowly than
Pseudocalanus with the assumed value of G_~~ ([Frost 1979), Pseudocalanus
predominates. Near the middle of the simulation (day 60), large cells dominate
the phytoplankton (Figure 7C), which promotes good growth conditions for
Calanus but not for Pseudocalanus. Thus, the composition of the zooplankton
shifts toward Calanus. However, near the end of the simulation (day 90),
nitrate has declined to such a low level (<0.8 ug of N/liter) that growth of
smaller cells is favored (Figure 7C) and feeding and growth conditions improve
once again for Pseudocalanus.
Effects on Higher Trophic Levels
The final application of the simulation model is to examine some possible
effects of pollutants which impact primarily the higher trophic levels such as
the predators of the herbivorous zooplankton. The effects may be of two
general types: An overall reduction in predation rate on both species,
reflecting general decline in abundance of all predators, and a selective
reduction in predation rate, reflecting qualitative shifts in the composition
of the predator assemblage.
For the first case the simulation model was run with the predation coef-
ficient m (eq. 2) decreased to 0.05, that is, half the value used in the
standard simulation. Because of decreased predation, the zooplankton initially
increase at a faster rate than in the standard simulation and the Rhytoplankton
bloom is terminated sooner (Figure 8A,B; cf. Figure 1). Nitrate concentration
8
-------
GRAZER=
Pseudocalanus
4 8 16 32 64 128
B
GRAZER = Calanus
80
4 8 16 32 64
CELL DIAMETER (jjm)
Figure 6. Size composition of the phytoplankton from day 20 to day 80 when
Calanus is omitted from the standard simulation (A) and when
Pseudocalanus is omitted from the standard simulation (B).
-------
o
E
to
6
o>
E
CJ
E
\
o
15
10
A
°o
2r
0
0
40
to
,
E
20
0
CHLOROPHYLL
—,
NITRATE
40
Pseudocalanus
80 DAYS
40
4 8 16 32 64
CELL DIAMETER (jum)
128
Figure 7. Simulated dynamics of plankton when the maximum specific growth
rate, G , of Calanus is decreased to 0.2. A, changes in time of
fllclX
chlorophyll and nitrate. B, changes in time of zooplankton biomass.
C, size composition of the phytoplankton at days 20, 60, and 90.
10
-------
15
0 10
o>
E
i_
o
•°E 5
\
IE
O
o>
s 0
A
CHLOROPHYLL
NITRATE
r
0
, 1 1
40
i
80
i
DAYS
(M
0
0
B
Calanus
Pseudocalanus
40
80 DAYS
TIME= 100 DAYS
4 8 16 32 64
CELL DIAMETER (jjm)
Figure 8. Simulated dynamics of plankton when the predation coefficient m (eq.
2) is decreased to 0.05. A, changes in time of chlorophyll and
nitrate. B. changes in time of zooplankton biomass. C, size
composition of the phytoplankton at day 100.
11
-------
declines more slowly than in the standard simulation and this favors the
growth of large cells which Pseudocalanus cannot utilize efficiently. Con-
sequently, the population of Pseudocalanus declines drastically after the
phytoplankton bloom. The increased abundance of zooplankton during the bloom
seems to accelerate the process, evident in the late stages of the standard
simulation (Figure 2), leading to strongly bimodal size frequency distribution
of the phytoplankton (Figure 8C). The result of this is that in the later
stages of the simulation both copepod populations consist of reproducing
adults whose offspring starve in the naupliar stages because of the scarcity
of phytoplankton cells in the size range of 4 to 32 urn.
Qualitative change in predation rate on the zooplankton was simulated by
alternately varying the predation coefficient m (eg. 2) for one of the two
grazer species. When m for Calanus is reduced to 0.05, while maintaining m
for Pseudocalanus at 0.1, the predicted plankton dynamics are similar to those
of the standard simulation except that Calanus dominates the zooplankton even
more strongly throughout the simulation and is much more abundant than Pseudo-
calanus at 100 days (Figure 9A). By contrast, when m for Pseudocalanus is
reduced to 0.05, it is favored over Calanus which, as in an earlier simulation
(Figure 6A), results in a shift in the phytoplankton composition toward very
large cells (Figure 9B). The predominant large cells cannot be efficiently
grazed by the young developmental stages of either species and therefore the
total zooplankton biomass gradually declines until at day 100 it is only 18%
of that predicted for the standard simulation (Figure 9B).
In conclusion, quantitative and qualitative changes in predation rate
strongly affect the growth dynamics of both the zooplankton prey and the
phytoplankton. This reinforces the conclusions of Landry (1976), Steele
(1976), and Steele and Frost (1977) that in simulation models of plankton
dynamics the pattern of predation on zooplankton is as important as any other
process in structuring the plankton assemblage.
DISCUSSION
The model of Steele and Frost (1977) provides one method for explicitly
describing the population dynamics of species in a plankton assemblage. A
model of this type will be required to simulate the effects of pollutants on
the dynamics of plankton because species differ in their tolerance to pol-
lutants. Yet even the present model is an extreme simplification. The assump-
tion of Steele and Frost (1977) that size of cell determines growth rate of
phytoplankton species, while in general correct, may misrepresent significant
variations among species within a narrow range of size (see, for example,
Figure 1 in Banse 1976). Indeed, the model has already been revised, as
described above, to account for differences in the growth rate of species of
copepods which are unpredictable from consideration of body size (Frost 1979).
The predictions of the model would also be very different if greater
complexity were introduced into the herbivore and predator components. For
example, inclusion of a microzooplankton component, grazing ^,on the very
12
-------
smallest sizes of cells, would probably stabilize the structure of the zoo-
plankton after the bloom. On the other hand, explicit description of popula-
tion dynamics of a predator, such as a species of chaetognath or ctenophore,
would very likely change the predictions in a material way, particularly if
the predator were size—selective its feeding.
Finally, it is highly unlikely that pollutants would affect only the
members of one trophic type. In reality, one could expect stress of several
modeling of effects of pollutants will therefore require extensive knowledge
on different trophic types.
40
20
0
ZBIOM*I339
C/P * 30.9
N = 2.20
A
10
E
\
o
6
60
40
20
ZBIOM= 212
C/P = 0.8
N = 0.66
B
4 8 16 32 64
CELL DIAMETER (pm)
128
Figure 9. Size composition of the phytoplankton from day 20 to day 80 when
Calanus is omitted from the standard simulation (A) and when Pseudo-
calanus' is omitted from the standard simulation (B).
13
-------
REFERENCES
Banse, K. 1976. Rates of growth, respiration and photosynthesis of unicellu-
lar algae as related to cell size - a review. J. Phycol. 12:134-140.
Evans, G. T., J. H. Steele, and G. E. B. Kullenberg. 1977. A preliminary
model of shear diffusion and plankton populations. Scottish Fish. Res.
Rept. No. 9.
Federle, T. W. , J. R. Vestal, G. R. Hater, and M. C. Miller- 1979. Effects
of Prudhoe Bay crude oil on primary production and zooplankton in Artie
tundra thaw ponds. Mar. Environ. Res. 2:3-18.
Frost, B. W. 1979. The inadequacy of body size as an indication of niches in
the zooplankton. _In The Evolution and Ecology of Zooplankton Populations,
W. C. Kerfoot (ed.). Amer. Soc. Limnol. Oceanogr. Spec. Sympos. III.
Greve, W. , and T. R. Parsons. 1977. Photosynthesis and fish production:
Hypothetical effects of climatic change and pollution. Helgolander wiss.
Meeresunters 30:666-672.
Koeller, P., and T. R. Parsons. 1977. The growth of young salmonids (Oncor-
hynchus keta): Controlled ecosystem pollution experiment. Bull. Mar.
Sci. 27:114-118.
Landry, M. R. 1976. The structure of marine ecosystems: An alternative.
Mar Biol. 35:1-7.
Menzel, D. W. 1977. Summary of experimental results: Controlled ecosystem
pollution experiment. Bull. Mar. Sci. 27:142-145.
Moraitou-Apostolopoulou, M. , and G. Verriopoulos. 1979. Some effects of
sub-lethal concentrations of copper on a marine copepod. Mar. Pollut.
Bull. 10:88-92.
Reeve, M. R. , M. A. Walter, K. Darcy, and T. Ikeda. 1977. Evolution of
potential indicators of sub-lethal toxic stress on marine zooplankton
(feeding, fecundity, respiration, and excretion): Controlled excosystem
pollution experiment. Bull. Mar. Sci. 27:105-113.
Steele, J. H. 1976. The role of predation in ecosystems models. Mar. Biol.
35:9-11.
Steele, J. H. , and B. W. Frost. 1977. The structure of plankton communities.
Phil. Trans. R. Soc. Lond. 6280:485-534.
Stockner, J. G., and N. J. Antia. 1976. Phytoplankton adaptation to environ-
mental stresses from toxicants, nutrients, and pollutants -- a warning.
J. Fish. Res. Bd. Canada 33:2089-2096.
14
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A QUANTITATIVE METHOD FOR EVALUATING
EXTERNAL EFFECTS ON ECOSYSTEMS
by
V. D. Fedorov, V. N. Maksimov, and V. B. Sakharov
Moscow State University
Different methods for evaluating external effects on the environment, and
ecosystems in particular, are discussed in this paper. However, there are
various difficulties which can only be surmounted by significant intellectual
and material efforts.
Questions concerning methods for evaluating the state of ecosystems,
specifically regarding their productivity, their ability to resist harmful
impacts, the preservation of their characteristics, and the potential for
exploitation are still debated. The situation is aggravated by the need to
overcome these problems soon if the balance between man and nature is to be
preserved. The complexity of the situation has fostered different approaches,
each of which has advantages and disadvantages. We shall discuss one method
for evaluating the state of ecosystems that was recently developed in the
Department of General Ecology and Hydrobiology at Moscow State University
(USSR).
STATEMENT OF THE PROBLEM
We shall examine a model of an ecosystem to test for the effect of pol-
lutants (see diagram). The center square represents an ecosystem, formed by a
set of populations which interact with each other and the biotope, each of
which is sufficiently heterogeneous in internal structure (sex, age, function)
and performs a similar or dissimilar function in biocenosis.
Transformation of energy and metabolism occurs within this ecosystem,
i.e., it is sufficiently autonomous and complex and has, like any spatially
defined entity, a series of integral characteristics which determine its
composition (set of populations), structure (interaction of population) and
changes in function (population shifts, changes in their relationships).
Indices of the resistance (conditioned by homeostasis), stability (conditioned
by maturity) and complexity (conditioned by the species difference and the
difference in the relationships between populations) are included in these
integral characteristics. The potentially harmful substances are indicated by
the broken arrows on the left, as the "input" into the system. It is hypothe-
sized that there, is a small number for each real ecosystem, that the number
can be determined approximately, and that the physical-chemical features,
i.e., the set of pollutants, will differ and vary from ecosystem to ecosystem.
15
-------
x,
V
E
-Y,
Diagram
It is known a priori that some portion of pollutants will "interact" with
each other in their effect on the ecosystem. This means that the effect of
each one individually will be substantially attenuated or potentiated as a
function of the effect of any other pollutant or pollutants.
The so-called indices of the state of the ecosystem, both with respect to
the biotic and abiotic parts, are indicated by the solid arrows to the right,
at the "output" of the ecosystem. Since the number, based on the condition k
» m [translator's note: handwritten notation illegible], we will eliminate a
limited number of indices which describe the ecosystem as "good" or "poor" as
a result of the effect of k pollutants based on some rules (it is assumed that
they have a satisfactory basis). We can then consider such indices of the
type as quality indices if the quality is determined by a set of properties of
the ecosystem which we consider desirable.
The diagram thus generally reflects the multiple effects of the pollut-
ants on the ecosystem whose condition is determined by multiple responses
selected as a function of human interests (e.g., indices of the quality of
drinking water) or in the interests of the ecosystem itself (e.g., preserva-
tion of the productivity characteristic of the given type of ecosystem). In
this case, the problem is formally reduced to studying the following type of
function:
y = f(x)
(1)
16
-------
In developing a strategy for biological monitoring, we previously
suggested specific procedures for establishing priority among the set of
interacting factors and determined the need to conduct a multifactor experi-
ment as a basic method for obtaining information on predicted consequences of
pollutants on the ecosystem (Fedorov 1976, 1977).
Matrices of first and second order plans (Maksimov and Fedorov 1969;
Golikova et al. 1974) were proposed as an experimental design which would
allow independent, simultaneous investigations of a combination of n selected
variables. These schemes are now routinely used in the form of a polynomial.
Thus, we can consider these methods and approaches for studying a set of
independent variables as well developed and tested.
WHAT ARE THE DEPENDENT VARIABLES AND
HOW SHOULD THEY BE INVESTIGATED?
An ecosystem has many features which determine its composition, struc-
ture, function, development, maturity, resistance, etc. Within the framework
of our task, these can be considered dependent variables. Since the number of
dependent variables significantly exceeds the number of active elements (k »
m), the need to limit the number of indices describing the condition of the
ecosystem becomes apparent.
Evaluations related to processes (related to the differences in types and
the force of interaction between organisms) and the results of these processes
(expressed by the relationships of the nature or biomass of the populations
forming the "face" of the biocenosis) are the basic indices of the structure
of ecosystems.
In addition, the need to add an indicative approach to the series of new
principles for describing the structural features of natural ecosystems seems
obvious. In this respect, regardless of whether the different indices are
related to estimates of the rate of the processes or to their results, we
shall examine the recent attempts to improve the estimates of pollution of
natural waters.
The first group of estimates consists of indices which could be expressed
by an integral in time, i.e., as some result of the effect (functions) at the
time they were recorded. This category includes the indices which describe
the size of the biomass, the number of species, the relationships of the
members, etc., and also the different indices for species resources, variety
and equivalence (expression), the relative abundance, domination, etc., calcu-
lated on this basis.
One of the most popular and successful models for evaluating water pollu-
tion based on a series of indices concerning the benthic fauna is the Woodwiss
method developed for the Trent River (Woodwiss 1964). Having significantly
reduced the list of indicative organisms, Woodwiss introduced the concept of
"group"—usually for denoting easily determined forms: in some cases,
species; in others even families. The principle of the indicative value of
such taxa is supplemented by the principle of a decreased variety of fauna in
polluted water.
17
-------
The second group of evaluations of the structure of ecosystems is
composed of indices which can be expressed by a derivative in time, i.e., as
the rate of change in some function. This category includes indices of
productivity, respiration, assimilation of substances, and other features of
the processes which take place in the ecosystem. The number of these indices
is large, thus, in order to monitor or judge the condition of the ecosystem,
it is necessary to select a few representative indices using any discrimina-
tory criterion. Earlier (Federov 1977), we formulated the basic requirements
for determining dependent variables and listed some reference indices
describing the condition of an ecosystem. We examined the possible situation
where certain combinations of "normal" and "pathologic" individual indices are
specific with respect to different pollutants and consequently the character-
istics of the combinations can be considered symptoms of a specific disease in
the ecosystem.
It is possible to examine the opposite situation which includes the
absence of any specific effect of the various pollutants, at least for indices
which describe the flow of energy and substances in the ecosystem. In this
case, the investigation of the principles of behavior of the individual
features permits interpretations of whether they are "good" or "bad" for the
function and existence of the ecosystem. The analysis of curves or a set of
data relating to evaluations of the intensity of a process essentially only
provides information concerning the quality of the index which we consider
demonstrative in judging whether this is "good" or "bad" for the system. The
need then arises to combine indices, some of which indicate the good quality
of the envionment or condition of the ecosystem and others which indicate the
poor condition of the ecosystem. Two successive stages are implied in solving
the problem which arises in this way: The first requires introducing an
evaluation of the quality of each dependent variable d.; the second involves
evaluating the quality D based on the results of evaluations of selected
indices d..
The features associated with d. and y. can be established by "common
sense" which absolutely rejects the presence of a linear correlation between
them, since the region of a sharp change in desirability in the norm implies
the existence of a region of more serious changes in function.
Harrington (in Maksimov 1977) has introduced the nonlinearity of the
formula on a continuous scale:
d = e"e i (2)
where z. is a coded variable which can easily combine a linear function with
any rea1! variables y.. The general form of function 2, which resembles a
logistics curve, is snown in Figure 1.* It is necessary to consider the fact
that the value z = 5; d = 0.98 with z = 4; d = 0.5 with z = 1.5; d = 0.68 with
z = i; d = 0.37 with z = 0; and finally, d = 0.0006 with z = -2 correspond to
desirability of 0.993 (i.e., very close to one). For discovering the correla-
* (Editor's note: figure not reproducible),
18
-------
tion of z with real dependent variables y., it is sufficient to combine the
limits of the normal variability of the dependent variables with the range of
changes in z from -2.2 to 5.
The detailed basis of the approaches used by Harrington's school to
evaluate the quality of biologic systems is given by Maksimov (1977).
However, Harrington's function is not adequate to evaluate the condition of
natural systems due to the monotonic nature of the correlation between d. and
y. which it describes. In reality, the type of function should be unimoaular
in biology in general and in ecology in particular, i.e., the curve for the
function should have one maximum corresponding to the optimum desirability.
If we use "more-frequently-better" as the initial position (instead of
"greater-better"), the desirability of any index can be determined by knowing
its distribution function.
Theoretically, distribution functions should in some way be related to
the concept of a statistical standard, but this question has not been investi-
gated since the concept of statistical standard is based on a hypothesis and
is only intuitively perceived (Fedorov 1977). Regardless of the nature of the
distribution, desirability equal to "1" can always be designated to a model
class value and, in this case, the basic difficulty is reduced to making a
decision concerning the rule for selecting the range to the left or right of
"1."
If we know the limits of the normal variability of the variable, it is
possible to combine each index for the desirability value equal to 0.63 with
any "extreme" normal value which corresponds to the lower repartition of the
concept "good" on Harrington's scale.
Many different methods can be used for determining the "extreme" value;
two are shown below.
The first method is borrowed from medical practice and is based on the
percentile method (Sepetliyev 1968). Any response values within the limits of
the 25th to 75th percentile should be considered normal, and a comparison of
the actual values of these percentiles for each response with a given desir-
ability value will provide the necessary bond between any real index and its
desirability value.
The second method is based on knowledge of the law of the distribution of
response indices in an intact system and the assumption that we can actually
determine the evaluation of the mean ("a") and the evaluation of the deviation
from the mean (a). The assumption of a hypothesis of normality (if necessary,
"normalization" should be obtained by taking the logarithm of the results)
allows isolating the region of "normal" values located within the limits of
"a" - a to "a" + a, corresponding to the evaluation of "good" and including
approximately two-thirds of the response evaluation figures.* The region of
the values for variables assigned by condition "a" ± 2a will also include
"bad" evaluations, indicating an alarming state of affairs in the ecosystem.
* Editor's note: figure not reproducible.
19
-------
In order to judge the "poor" or "good" state of the ecosystem as a whole, we
should use a generalized desirability index which could be calculated as the
mean geometric set of evaluations d-
D = rv/dj, d2 ... dm (3)
where d. are the coded values of the particular desirability values found by
one of the methods described above.
TABLE 1. RECOMMENDED INTERVALS IN THE
DESIRABILITY SCALE IN A 5-POINT SYSTEM FOR
EVALUATING THE WELL-BEING OF ECOSYSTEMS
State Points
1.0 -
0.8 -
0.6 -
0.4 -
0.2 -
0.8
0.6
0.4
0.2
0.0
Excellent
Good
Fair
Poor
Very Poor
5
4
3
2
1
AN EXAMPLE OF EVALUATING THE STATE OF ECOSYSTEMS
USING THE DESIRABILITY FUNCTION
Calculation of the desirability values is based on the hydrochemical data
given in the Woodwiss article (1977). In this study, the results for deter-
mining the biotic index in different sections of the Trent River and a number
of its tributaries are compared with the values of the most important hydro-
chemical indices measured at the same stations.
Based on the graphs in the Woodwiss work, graphs were plotted to convert
the natural values of the selected indices to their desirability values.
The generalized desirabilities were calculated for 49 stations based on
the data cited in Woodwiss1 work. The individual desirability values for all
three indices and the generalized desirability values D calculated on their
basis are shown in Table 2. This table also shows the biotic index value
determined by Woodwiss at these stations for comparison. Based on his state-
ment, the deviations in estimating the index usually do not exceed 1 point on
the scale. Within these limits, the desirability value D, multiplied by 10,
coincides with Woodwiss' index in 34 cases, and the difference between the
desirability and the index exceeds 2 units in only four cases. This agreement
of the data should be considered totally satisfactory.
In the example, the desirability function was used to evaluate the effect
of different pollutants on an ecosystem. The results of an experiment whose
purpose was to determine the effect of two petroleum products—diesel^ fuel and
motor oil—and a dispersion substance—correxite 7664—on White Sea plankton
20
-------
TABLE 2
Station
No.
1
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
37
38
39
40
41
42
43
44
45
46
47
48
49
Total Permanganate
Oxidi zability
( /ing/liter) and
and Its Desirability
2
5.2
6.7
7.9
6.2
7.3
6.2
4.6
8.8
8.9
2.0
3.8
3.2
1.5
7.5
8.0
6.3
4.8
6.4
2.8
6.0
4.8
3.6
4.6
6.0
5.0
5.0
6.0
4.0
3.0
4.0
4.0
3.0
5.0
5.0
4.0
5.0
5.0
6.0
6.0
7.0
9.0
4.0
6.0
5.0
10.0
10.0
10.0
10.0
12.0
3
0.68
0.53
0.43
0.58
0.48
0.58
0.74
0.36
0.36
0.93
0.81
0.85
0.96
0.46
0.42
0.57
0.72
0.56
0.83
0.60
0.72
0.38
0.74
0.60
0.70
0.70
0.60
0.79
0.87
0.79
0.79
0.87
0.70
0.79
0.70
0.79
0.60
0.60
0.50
0.35
0.79
0.60
0.70
0.29
0.29
0.29
0.29
0.29
0.20
Ammonium Nitrogen
( /ing/liter)
and Its Desirability
4
0.4
3.7
2.1
0.4
2.4
0.6
0.2
9.1
1.2
0.0
1.5
0.1
0.1
1.5
5.4
4.3
0.4
7.4
0.0
1.2
0.5
3.8
0.2
1.9
1.5
1.4
2.3
0.2
0.2
0.2
0.3
0.3
0.3
0.8
0.8
1.6
0.9
2.1
1.2
3.0
2.7
0.3
0.4
0.5
6.2
7.7
6.2
5.3
7.1
5
0.34
0.29
0.46
0.84
0.42
0.78
0.92
0.09
0.62
1.00
0.56
0.96
0.96
0.56
0.17
0.24
0.84
0.11
1.00
0.62
0.83
0.28
0.92
0.49
0.50
0.58
0.37
0.93
0.93
0.93
0.83
0.83
0.88
0.72
0.72
0.55
0.69
0.47
0.62
0.67
0.39
0.88
0.85
0.82
0.14
0.10
0.14
0.18
0.12
Dissolved Oxygen
and Its Desirability
( /ing/liter)
6
9.9
7.0
8.1
8.1
6.3
9.2
10.3
6.0
7.2
11.7
6.3
12.5
12.4
9.5
6.4
10.2
11.1
9.8
13.7
9.3
9.4
5.6
9.8
7.7
7.4
7.1
6.3
10.9
10.8
9.6
10.7
18.4
9.2
9.5
9.1
8.3
8.4
6.0
8.1
7.8
6.7
10.5
9.6
10.6
6.3
6.1
7.6
7.0
6.6
7
0.88
0.37
0.46
0.46
0.32
0.64
0.92
0.30
0.39
1.00
0.32
1.00
1.00
0.74
0.32
0.92
0.99
0.86
1.00
0.68
0.74
0.28
0.86
0.43
0.40
0.37
0.32
0.98
0.97
0.79
0.96
1.00
0.64
0.78
0.62
0-48
0.49
0.30
0-46
0.44
0.34
0,94
0.73
0.93
0.32
0.30
0.41
0.37
0.84
Desir-
ability
8
0.80
0.33
0.45
0.61
0.40
0.66
0.86
0.21
0.44
0.98
0.53
0.93
0.98
0.58
0.28
0.50
0.84
0.38
0.96
0.63
0.76
0.31
0.84
0.5d
0.54
0.53
0.41
0.90
0.92
0.88
0.87
0.91
0.73
0.73
0.71
0.57
0.62
0.44
0.56
0.53
0.36
0.87
0.74
0.82
0.24
0.21
0.26
0.27
0.20
Biotic
Index
9
7
4
5
6
5
6
8
2
6
10
6
8
10
4
2
7
8
4
7
6
8
3
6
3
5
4
4
10
10
10
9
9
9
7
7
6
4
4
4
2
2
8
8
8
4
2
i 3
' 3
3
Difference
10
1.0
0.2
0.5
0.1
1.0
0.6
0.6
0.1
1.6
0.2
0.7
1.3
0.2
1.8
0.8
2.0
0.4
0.2
2.6
0.3
0.4
0.1
2.4
2.0
0.4
1.3
0.1
1.0
0.8
1.7
0.3
0.1
1.7
0.3
0.1
0.3
2.2
0.4
1.6
3.3
1.6
0.7
0.6
0.2
1.6
0.1
0.4
0.3
0.1
21
-------
when certain combinations of the substances were added according to a plan for
total factor experiment (TFE) 23 were used. The state of the plankton was
evaluated based on changes in the number of the five species of algae and four
types of zooplankton which dominated at the time the experiments were
conducted (July 1975) and based on changes in production indices in fractions
of the controls, set at 1 for all 11 response functions (Table 3).
TABLE 3
1
Controls
8
Concentration, mg/1
Diesel fuel x4
Motor oil x2
Correxite x3
0
0
0
10
0
0
0
10
0
10
10
0
0
0
10
10
0
10
0
10
10
Ratio of Experimental Values to Control Values
Scelentonema
costatum yx
Small flagellata y2
Chaetoceros
wighamil y3
Cyanobacteria y4
Dynabrion
pellucidum ys
Microzetella
norvegica y6
Oitona similis y7
Acartia longiremis y8
Temora longicornis y9
Primary production y10
Tempo assimi-
lation C02 yn
Overall desirability D
Factor symbols
Regression
coefficients b.
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
ii -I n
0.
0.
10
06
0.12
0.16
0.09
0.12
0.28
0.60
0.08
0.04
0.16
0.12
0.98 0.17 0.07
0.92 0.06 0.63
0.89 0.21 0.29
0.12 0.97 0.10
0.95
0.92
0.14
0.48
0.82
0.25
0.24
xi
0.97
0.57
0.99
0.98
0.82
0.70
0.39
X2
0.86
0.32
0.14
0.41
0.56
0.93
0.22
X1X2
0.98
0.79
1.00
0.99
0.93
0.92
0.81
0.88
0.50
0.14
0.48
0.32
0.95
0.24
XXX3
0.93
0.19
0.99
0.98
0.94
10
10
10
0.07
0.05
0.21
0.16
0.10 0.06 0.11 0.19 0.10 1.00 0.14
0.75
0.10
0.12
0.41
0.10
0.72 0.34
0.49 0.15
X2X3 X1X2X3
0.442 -0.230 -0.130 0.102 -0.020 0.002 0.028 -0.045
The relationship between each of these functions and the three factors
studied generally does not coincide with the rest. If the number of diatoms
decreases under the effect of the petroleum products, the growth of blue-green
algae will be significantly stimulated with certain values of impurities, with
which the increase in primary productivity in those variants of the experi-
ments where diesel fuel was added is apparently related. The different types
of zooplankton significantly differed with respect to sensitivity to the
petroleum products and the correxite.
22
-------
Desirability functions can be used to conduct a complete evaluation of
the effect of the pollutants on the entire set of indices characterizing the
condition of the plankton. In this respect, a basic difficulty exists in
selecting the scale for converting the actual response function values to
their desirability values. It is more or less clear that it would be more
desirable to consider a situation where there is generally no harm in evalu-
ating the damaging effect of external factors on the system. This means that
the value of the response function in the controls should be used for 1 on the
desirability scale. Establishing such a value for the response function in
which the desirability should be considered equal to 0 is less specific. In
particular, for a number of individual species in a community, decreasing them
to 0, i.e., the total death of the population, is naturally an extremely
undesirable phenomenon, but it is also clear that decreasing the number of
species to some critical value, beyond which the population cannot recover
even after the harmful effect has been eliminated, is no less undesirable.
Since the critical values for the number of real species in real ecosystems is
still unknown, we selected some intermediate point whose position can be
determined on the basis of any biologically plausible hypotheses for con-
structing the desirability scales. In our example, we used the rather widely
held opinion that a two-fold decrease in the number of the population with one
harmful effect should be considered acceptable in the sense that this popula-
tion preserves the capacity to return to the initial state after the pollution
has been eliminated (e.g., due to self-purification of a reservoir). We
extended this hypothesis to the productive indices due to the absence of any
other sufficiently plausible hypotheses. Indices which were higher in the
experiments than the controls, i.e., indices for which y. > 1 were doubled in
comparison with the controls.
In this context, "acceptable" means a deviation from the control which
corresponds to an evaluation between "good" and "excellent" on the desir-
ability scale. In our case, we selected the numerical value of the desir-
ability d = 0.85 for quantitatively expressing this evaluation. The value of
the arbitrary variable z = 1.9 corresponds to this value. ^We used z = 5.0 for
the values of y. in the controls; this gives a value of D ~ 1 in using formula
(2). In selecting these "reference points," the values of y. can be converted
to z values by using the linear transform
z = Vis'2 • where
p = u when y ^ 1.
p = 1/y when y > 1.
Table 2 shows the desirability values calculated by this method.
The same table shows the values of the regression coefficients calculated
according to the values for the three factors studied and their interactions
[verification of the significance using Daniel's method [Daniel 1959)] showed
that diesel oil had a more pronounced effect than motor oil on plankton (based
on the total of 11 selected indices) in certain hypotheses concerning the
23
-------
desirability of changes caused by additives. The effect of each of them was
greater separately due to the presence of a positive effect of their inter-
action. This interaction should be considered the sign of a monotypic
nonlinear "dose-effect" function for the two substances; as a result, adding
each of them to a system which is already subject to the inhibiting effect of
the second does not cause an equally great effect. A similar case has been
previously examined (Maksimov 1977).
The effect of correxite on the value of the desirability function, like
the effect of its interaction with the petroleum products, was insignificant;
it is thus possible to conclude that a 10 mg/liter concentration of correxite
does not cause any significant undesirable changes in plankton and does not
alter the toxicity of the petroleum products if "toxicity" is defined as the
ability of a substance to cause undesirable changes in a community or
ecosystem as a whole.
CONCLUSION
Using the existing methods for evaluating standards, and establishing the
extreme limits of "good," etc., it is doubtful whether it is expedient to
attempt to improve a proposed system for evaluating the state of an ecosystem.
It can be accepted or rejected on the basis (again!) of a totally subjective
opinion concerning the fact that it is worse or better than other evaluation
systems. Attempting to combine the different schemes or, on the other hand,
allowing them to exist in parallel and independently if knowing the logical
basis and structure of the design is the only criterion for their existence,
could also be justified.
It is doubtful whether any one system could be developed which would be
better than all of the others and which would consequently be preferred in
analyzing the conditions of ecosystems. As a function of human goals, the
features of the conditions and communities and the methods of evaluation will
be different in each case. However, the principles on which the methods of
evaluation are based cannot be numerous. Based on general considerations, it
would seem that all methods will concern the features of the functional-
spatial structure of ecosystems or their living components to a greater or
lesser degree. The integrity of such structures as ecosystems cannot be based
on the features of their organization. The methods of evaluation will be
easier to vary as a function of the type of ecosystem, its maturity, and
"resistance reserves" (resistance) when they are used by humans. With respect
to toxicologic evaluations of the quality of water, despite their primitivism
and low price, their role in making ecologically based decisions cannot be
decisive.
SUMMARY
The problem of evaluation of the state of the ecosystem on the basis of
many structural and functional indices has been considered. Rules for
choosing these indices during the elaboration of the system of control of the
quality of the environment are suggested. The main difficulty in carrying out
such a control was shown, namely: the numerous response functions and inde-
24
-------
pendent variables. Establishing a generalized criterion for the state of the
ecosystem which unifies individual indices makes it possible to overcome this
difficulty.
The use of the so-called desirability function is proposed as one
approach to the solution of this problem. The principles of the transition of
the concrete indices to the conditional scales of desirability are also
described in the paper. The practical application of the proposed approach is
shown by two specific examples.
REFERENCES
Fedorov, V. D. 1976. Problems in the maximum admissible effects of the
anthropogenic factor with respect to the ecologist. _In Comprehensive
Analysis of the Natural Environment. Gidrometeoizdat Press, Leningrad.
Fedorov, V. D. 1977. The problem of evaluating the normal and pathologic
condition of ecosystems. In Scientific Principles for Controlling the
Quality of Surface Waters Based on Hydrobiologic Indices. Proc.
Sov.-Eng. Seminar, Gidrometeoizdat Press, Leningrad.
Golikova, T. I., L. A. Panchenko, and M. Z. Fridman. 1974. List of second
order plants. Izd-vo MGU.
Maksimov, V. N. , and V. D. Fedorov. 1969. Mathematical planning of biologic
experiments. _In Mathematical Methods in Biology 1968. Izd-vo VINITI,
Moscow.
Maksimov, V. N. 1977. Specific problems in studying the combined effect of
pollutants on biologic systems. Gidrobiol. Zhurn. 13(4).
Sepetliyev, D. A. 1968. Statistical methods in scientific medical studies.
Meditsina Press, Moscow.
25
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PETROLEUM POLLUTION AND MICROFLORA IN MARINE ECOSYSTEMS
by
M. V. Gusev, T. V. Koronelli and V. V. Il'inskiy
Department of Biology
Moscow State University
INTRODUCTION
Petroleum and petroleum products have become integral components of the
marine environment. In undergoing microbiologic oxidation, these compounds
enter the natural cycle and their presence has a definite effect on the
microflora of ecosystems.
Since petroleum products are almost insoluble in water and are less dense
than water, their distribution in the ocean is not uniform. There are
critically polluted areas (petroleum discharges resulting from accidents,
ports, petroleum output regions, outfalls discharging polluted water) where
the surface of the ocean is completely or partially covered by a film of
petroleum; moderately polluted areas where petroleum cannot be visually
detected and there is no film, but the concentration of petroleum products
exceeds the maximum admissible quantity (open waters of the ocean); and
finally, there are sections which are free of petroleum and the concentration
of petroleum products is below the maximum admissible concentration (some of
the world unperturbed by man). The effect of petroleum pollution on micro-
flora is a function of its intensity. For this reason, our studies involved
areas of the ocean with different degrees of pollution: ports and polluted
coastal waters, the open sea and regions totally free of pollution.
The following aspects of the problem were emphasized:
(1) Determining the quantitative mechanisms of the distribution of
petroleum-oxidizing microorganisms and microorganisms from other
groups;
(2) Evaluating the possibility of indicating petroleum pollution using
specific microflora;
(3) Determining and studying the active petroleum-oxidizing micro-
organisms.
The regions investigated were generally located in the northern
latitudes. They included the Finnish Straits of the Baltic Sea, the Arctic
Ocean of the USSR (White, Barents, Kara, Laptev Seas), the coastal waters of
Wrangel Island, the coastal waters of the Komandorskiy Islands* the north-
western part of the Pacific Ocean (Figure 1).
26
-------
100
120
140
160 180
160
140
120
100
60
/ / / 7 7
60
40
20
20
120
Figure 1. Location of the sampling sites in the Pacific Ocean.
We note that similar studies have been conducted previously in the
regions of the Soviet Arctic. However, pollution from petroleum products
represents the greatest hazard in these latitudes, since petroleum bio-
degradation processes are slow at low temperatures. The quantitative
mechanisms in the distribution of petroleum-oxidizing microflora and other
microorganisms were studied in the Arctic Seas, the coastal waters of the
Komandorskiy Islands, and the northwestern section of the Pacific Ocean.
Water samples from all of these regions, and samples from the coastal waters
in the Finnish Straits of the Baltic Sea and the region of Wrangel Island were
used to determine and investigate the active petroleum-oxidizing micro-
organisms.
METHODS
In studying the coastal waters, the samples were taken from the surface
layer with a 100-ml sterile bottle; in open waters the samples were taken with
a Zobell bathometer.
27
-------
In working in the open waters of the Pacific Ocean (October-December
1977), the samples of water were taken from a boat which was 1 km away from
the ship. Samples of the surface film were taken with a sterile capron
screen, and samples were taken from the meter level with Zobell bathometer.
The water temperatures during the investigation period were as follows:
0 to 6° in the Arctic Seas (surface partially covered with ice), 2 to 4° in
the coastal waters of the Mednyy I. (Komandorskiy I.) and 20 to 22° in the
northwest part of the Pacific Ocean.
The total number of bacteria was determined on "synpor" membrane filters
with pore diameter of 0.2 urn. The number of individual groups of micro-
organisms was determined by the cup method in the following media: MPA
(heterotrophs), Tiller's medium (oligocapnophils), Chapek medium with 1%
diesel fuel (hydrocarbon-oxidizing + hydrocarbon-resistant). SGO with diesel
fuel was inoculated in silica gel medium to calculate the hydrocarbon-
oxidizing microflora from the microorganisms separated in the Chapek medium
with diesel fuel (Walker and Colwell 1976). Carbon tetrachloride extraction
with subsequent measurement in an "oil-102" device (Japan) was used to
determine the hydrocarbons in the water samples. Krasil'nikov and Berg
detectors and some original works were used to determine the species affilia-
tion of the hydrocarbon-oxidizing microorganisms.
NUMBER AND DISTRIBUTION OF PETROLEUM-OXIDIZING
AND OTHER MICROORGANISMS IN THE OCEAN
During the Arctic expedition of 1974, the northern seas were examined
along the routes from Arkhangel to the Tiks Bay. The level of pollution could
be described as moderate in the open waters and critical in ports and local
spots in the straits. The number of microorganisms in the surface waters,
counted in agar medium containing a petroleum product, consisted of 1500-7000
cells/ml. In the relatively shallow areas (9-39 m), as the depth increased
and in the subsurface layer, there were one to two orders of magnitude fewer
microorganisms. The number of petroleum-oxidizing microorganisms in the
deep-water regions (129-208 m) on the surface and subsurface layer was 1 cell
in 10 ml of sample. The number of petroleum-oxidizing bacteria in the surface
layer of the port, Dikson, where strong pollution was observed, was 10 times
greater than in the samples from the open sea. However, the same number of
colonies grew in the media with diesel fuel and machine oil. Agar-treated
mineral medium containing no hydrocarbons was used as the control. We found
that a significant number of microorganisms grew in the control medium, often
equaling the number of microorganisms in the medium containing the petroleum
product. In investigating the microorganisms isolated in the medium contain-
ing the petroleum product in the laboratory, we found that only 46.4% of the
strains also oxidized the petroleum product in a liquid medium; the remainder
grew because of the agar and the impurities contained in it.
Using agar-treated medium containing a petroleum product in quantita-
tively calculating petroleum-oxidizing microflora produced results which were
too high (almost two times) due to the oligocapnophilic forms. However,
despite this fact, the number of petroleum-oxidizing microorganisms in the
28
-------
northern seas of the USSR is high, and the number is significantly higher in
the chronically polluted regions. At the same time, a significant number of
microorganisms which cannot actively use petroleum products but which carry
high concentrations of these compounds, also live in the seas.
The results of the Arctic expedition induced us to turn our attention to
the method of calculating the petroleum-oxidizing microorganisms and their
quantitative relationships with microorganisms from other physiologic
groups--heterotrophs and oligocapnophilic organisms. The method of saturated
cultures in a liquid medium containing a petroleum product used for
calculating petroleum-oxidizing microflora is not sufficiently precise and is
inconvenient, particularly in conducting investigations in water with a low
number of bacteria (less than 1 c/ml). As Japanese and American authors have
shown, the most precise results are obtained by inoculations in silica gel
medium containing a petroleum product (Seki 1973; Walker and Colwell 1976).
The disadvantage of this method is based on the fact that it is method-
ologically difficult to prepare silica gel medium and impossible to do so in
the conditions of an expedition. For this reason, in the subsequent studies,
we used agar-treated medium containing a petroleum product, representing the
bacteria which grew in the medium as the sum of hydrocarbon-oxidizing and
hydrocarbon-resistant bacteria, and we then inoculated the microorganisms
separated in the laboratory in silica gel medium with a petroleum product and
calculated the number of hydrocarbon-oxidizing microorganisms. We simul-
taneously developed a method for preparing silica gel which was suitable for
field conditions.
The study of the quantitative relationships between petroleum-oxidizing
and other microorganisms revealed the role and significance of petroleum-
oxidizing microflora in the ecosystem with respect to the level of pollution.
These microbiologic studies were conducted in open waters in the northwestern
part of the Pacific Ocean. We investigated the following: (1) the number and
distribution of saprophytic, oligocapnophilic and hydrocarbon-oxidizing
bacteria between the surface film and the subsurface water; (2) the activity
and resistance of bacterioneustons and bacterioplankton to petroleum products;
and (3) the indicative significance of hydrocarbon-oxidizing bacteria.
The number of bacteria from all of the groups studied was higher in the
surface film than at the meter level (Table 1). Saprophytes predominated in
both levels. The number of saprophytes in the surface microlevel varied from
70 to 5500 cells per 100 ml of water; it did not exceed 500 c/100 ml at five
stations, and was over 1000 c/100 ml at the seven others. The enrichment
factor E* at the six stations did not exceed 10 and was greater than 10 at
only three stations; the maximum value was 59.5. A higher number of
saprophytic bacteria was observed on the meter level at three stations. The
number of saprophytic bacteria on this level varied from 10 to 3100 c/100 ml
and exceeded 1000 c/100 ml at only four stations. A correlation was found
between the number of this group in the surface microlayer and the water layer
(r = 0.58).
The ratio of the concentration of bacteria in the surface microlayer to the
concentration on the meter level.
29
-------
TABLE 1. MEAN NUMBER (M) OF BACTERIA FROM DIFFERENT GROUPS AND ENRICHMENT
FACTOR (E) AT STATIONS IN THE NORTHWEST PACIFIC OCEAN
CO
o
Station
No. Level
1
2
3
4
5
6
7
8
9
10
11
12
220
1
220
1
220
1
220
1
220
1
220
1
220
1
220
1
220
1
220
1
220
- 1
220
1
pm
m
pm
m
pm
m
pm
m
pm
m
pm
m
pm
m
pm
m
pm
m
pm
m
pm
m
pm
m
Saprophytes
M
(c/100 ml) log M
1350
205
103
220
6200
500
1290
550
55000
925
1900
1300
24000
1050
300
230
425
1600
3900
3100
115
100
70
355
3.13
2.31
2.01
2.34
3.79
2.70
3.11
2.74
4.74
2.97
3.28
3.11
4.38
3.02
2.48
2.36
2.63
3.20
3.59
3.49
2.06
2.00
1.85
2.55
E
0.6
0.5
12.4
2.6
59.5
1.5
22.9
1.3
0.3
1.3
1.2
0.2
Oligocarbophils
M
(c/100 ml) log M
200
100
85
4
2000
4
116
1
340
180
500
1000
14000
70
440
47
620
360
3800
800
40
2
1
1000
2.30
2.00
1.93
0.60
3.30
0.60
2.06
0.00
2.53
2.26
2.70
3.00
3.15
1.85
2.64
1.67
2.79
2.56
3.58
2.90
1.60
0.30
0.00
3.00
E
2.0
21.3
500.0
116.0
1.9
0.5
200.0
9.4
1.7
4.8
20.0
0.0
HC-oxidizing+HC- resistant
M
(c/100 ml) log M E
60
50
11
20
1000
21
255
1
850
265
570
1400
6450
90
245
160
250
1605
2200
2100
60
4
4
30
1.78
1.70
1.04
1.30
3.00
1.32
2.41
0.00
2.93
2.42
2.76
3.15
3.81
1.96
2.39
2.21
2.40
3.21
3.34
3.32
1.78
0.70
0.60
1.49
1.2
0.6
47.6
255.0
3.2
0.4
71.7
1.5
0.2
1.1
15.0
0.1
-------
The number of oligocarbophilic bacteria in the surface film exceeded the
number of hydrocarbon-oxidizing and hydrocarbon-resistant bacteria at seven
stations and varied within the limits of 1 to 14,000 c/100 ml; however, the
number of bacteria from this group exceeded 1000 c/100 ml at only three
stations. The enrichment factor did not exceed 10 at five stations, and
varied from 20 to 500 at five stations. A higher number was observed at the
meter level at two stations. The number of this group of bacteria in the
meter level did not exceed 100 c/100 ml at seven stations; there were 1-4
c/100 ml at four stations, and the number varied from 180 to 1000 c/100 ml at
the other five stations. No correlation was found between the number of
oligocapnophilic bacteria at the two levels studied.
The number of hydrocarbon-oxidizing and hydrocarbon-resistant bacteria in
the surface film varied from 4 to 1000 c/100 ml at the ten stations, and
reached 2200 and 6450 c/100 ml at only two stations. At the meter level, the
number was 100 c/100 ml at seven stations, and varied from 160 to 2100 at the
other five. The enrichment factor for the microsurface level did not exceed
3.2 at four stations, and varied from 15 to 71.7 elsewhere; the factor was 255
at only one station. A high number of bacteria was found at the meter level
at only four stations. No correlation was found between the number of
bacteria in the surface film and at the meter level; this indicates that there
is no direct correlation between the number of bacteria in these two levels.
Overall, the mean values for the number of different bacteria in the
surface film were (per 100 ml): saprophytes 7888 c, oligocapnophilic 1845 c,
hydrocarbon-oxidizing + hydrocarbon-resistant 996 c; the values for the meter
level were, respectively: 845, 297, and 479 cells.
A correlation was found between the number of all of the groups of
bacteria studied both in the surface film and on the meter level (Table 2).
The data obtained concerning the high number of saprophytic bacteria in the
surface film in comparison to the meter level generally agree with the data in
the literature.
However, an inverse correlation was sometimes found; this could be due to
the presence of substances in the surface film which inhibit the development
of bacteria (Sieburth 1972), and to a combination of other unfavorable
factors—intensive solar radiation, high surface tension, high oxidation-
reduction potential, hydrologic conditions, etc. The result of this could be
a decrease in the number of bacterioneustons; as a consequence, accumulation
of bacteria in the surface film would not adequately replace the cells from
dead microorganisms (Dietz et aj. 1976).
There was a close correlation between the number of hydrocarbon-oxidizing
and oligocapnophilic bacteria and also between the hydrocarbon-oxidizing
bacteria and the saprophytes; the relatively high correlation coefficient is
an indication of this (Table 2). This apparently indicates that representa-
tives of both the saprophytes and the oligocapnophilic bacteria are included
in the composition of the hydrocarbon-oxidizing and hydrocarbon-resistant
bacteria, and a closer correlation between this group and the number of
oligocapnophilic bacteria was observed on the meter level.
31
-------
TABLE 2. CORRELATION COEFFICIENT FOR THE NUMBER OF BACTERIA FROM THE DIFFERENT
PHYSIOLOGIC GROUPS
Level
Physiologic Group 220 urn 1m
1. Oligocapnophilic -
saprophytes 0.67 0.59
2. Hydrocarbon-oxidizing and
hydrocarbon-resislant
saprophytes 0.85 0.70
3. Hydrocarbon-oxidizing and
hydrocarbon-resistant
Oligocapnophilic 0.82 0.84
The hydrocarbon-oxidizing and hydrocarbon-resistant bacteria at seven of
the 12 stations represented a high percentage of the total number of the
heterotrophic population at the meter level (Table 3). This indicated that
the biochemical activity of the bacteria in the surface film is no higher than
the activity of the bacteria in the water layer with respect to petroleum
products.
TABLE 3. NUMBER OF HYDROCARBON-OXIDIZING AND HYDROCARBON-RESISTANT BACTERIA
IN PERCENTAGES OF THE TOTAL NUMBER OF HETEROTROPHIC BACTERIA
Level
Station No.
12 3 4 5 6 7 8 9 10 11 12
Surface
layer
220 pm 3.9 5.9 12.2 18.3 1.5 23.8 17.0 33.1 23.9 28.6 38.7 5.6
Meter
layer 16.4 8.9 4.0 0.2 24.0 60.9 8.0 57.8 81.9 53.8 3.9 1.9
The highest quantity of hydrocarbons was contained in the surface layer
(Table 4): from 0.38 to 3.22 ing/liter.
The concentrations of hydrocarbons on the meter level were significantly
lower: from 0.06 to 0.19 mg/liter. The indicative significance of petroleum-
oxidizing bacteria has been discussed in the literature.
32
-------
TABLE 4. CONCENTRATION OF HYDROCARBONS IN SAMPLES OF WATER FROM THE NORTHWEST
PACIFIC OCEAN (in mg/liter)
Station No.
Level
220 M"i
1 m
1
3.22
0.19
2
2.94
0.14
3
3.22
0.06
4
1.62
0.64
5
1.94
0.10
6
1.74
0.18
7
0.80
0.14
8
0.88
0.15
9
1.06
0.12
10
0.48
0.08
11 12
0.72 0.38
0.18 0.15
This concerns whether or not it is possible to determine the degree of
pollution based on the number of hydrocarbon-oxidizing bacteria. As our data
indicate, no correlations between the number of hydrocarbon-oxidizing and
hydrocarbon-resistant bacteria and between the fractions of these bacteria in
the entire heterotrophic population and the concentration of hydrocarbons were
found for the surface film or for the meter level of open waters (Figures 2,
3).
The presence of bacteria which could grow in hydrocarbons and mixtures of
hydrocarbons in the water samples were not correlated with the concentration
of hydrocarbons in these samples. Walker and Colwell (1976) arrived at a
similar conclusion in conducting studies in the Chesapeake Straits. This
conclusion is logical if we consider studies of hydrocarbon absorption by the
cells of microorganisms. It has been found that oxidation of hydrocarbons by
microorganisms takes place when the cells are in direct contact with the
substrate; the hydrocarbon enters the cell by passive diffusion and is
solubilized in the lipophilic parts of the cell wall (Koronelli 1979). The
concentration of the hydrocarbon in the medium should not be less than 0.1
g/liter for this process to take place. The quantities of petroleum products
usually observed in open waters are significantly lower than the "substrate"
concentrations; this explains the absence of any correlation between their
concentration "and the number of petroleum-oxidizing microorganisms. The
necessary "substrate" concentrations usually occur in chronically and
critically polluted bodies of water (ports, petroleum products and discharge
sites, accidental discharges of petroleum). However, it is necessary to
remember that the number of hydrocarbon-oxidizing bacteria does not
necessarily increase proportionally at such sites, since their development
will be limited by a number of other factors, primarily a shortage of biogenic
elements—nitrogen and phosphorus.
We conducted studies in the northern part of the Pacific Ocean in the
area of Mednyy Island to investigate the distribution of hydrocarbon-oxidizing
microflora in unpolluted waters. This island, which is part of the
Komandorskiy Archipelago, is one of the few places where the effect of human
activity on the surrounding environment is almost absent. There are not
permanent settlements on the island, and the marine freight routes are far
away. For this reason, the coastal waters can be considered an unpolluted
ecosystem.
33
-------
LOG A
%A
40
30
20
10
0
0
2
I
0
0
3L
3 C
H
Figure 2. Correlation of the number of hydrocarbon-oxidizing and hydrocarbon-resistant bacteria (I) and
their percentage of the total heterotrophic population (II) with the amount of hydrocarbons in
the surface film;
j>
log A - Logarithm of the number of bacteria,
%A - percentage of the number of hydrocarbon-oxidizing and hydrocarbon-resistant
bacteria in the total heterotrophic population,
C. - concentration of hydrocarbons in mg/liter.
-------
LOG A
0
%A
80
60
40
20
0 0.1
0.2 C
0
H
0
0.1
0.2 CH
Figure 3. Correlation of the number of hydrocarbon-oxidizing and hydro-
carbon-resistant bacteria (I) and their percentage of the total
heterotrophic population (II) with the amount of hydrocarbons at the
meter level; symbols same as for Figure 2.
Although all of the island systems in the North Pacific share many common
features, each one has its own characteristics which determine not only the
topography and climate, but also the communities which populate them. In
particular, there are large lairs of marine animals on Mednyy Island—seals
and dotterel. Rich populations of seaweed have developed in the coastal
waters, forming dense overgrowth in some places.
The microbiologic studies in the coastal waters off Mednyy Island were
conducted during the Northern Expedition of the Biological Faculty of MSU in
June-July 1976. The total number of microorganisms and the number of
individual groups—heterotrophic, oligocapnophilic and hydrocarbon-oxidizing
microorganisms were determined. The microbiologic studies also included both
the western and eastern coasts of the southeastern part of Mednyy Island;
there were 23 stations (Figure 4). The total number of bacteria varied within
the limits of 10-15 million c/ml. Oligocapnophilic bacteria were most
prolific; the number at the different stations varied from 120 to 24,000 c/ml,
and they were predominant at 12 out of 23 stations (Figure 5).
35
-------
16
Figure 4. Southeastern part of Mednyy Island.
taken.
LOGm
^Stations where samples were
4
3
2
1
STATION
SAMPLING SITE
-
-
— •,
\
t
•
w
•r
H •
»
ff
0
•
_
-
^^^
^
fj
1 1 7 I 8 1 9 1 13 1 10 1
•1
1 12
GLIUKA BAY JPODEMNAYA BA
a
-
17 1 18 1 19
Y OZHIDDIYA BAY
6 1
PERESHSEI BAY
LOGm
J
4
3
2
\
1
STATION
SAMPLING SITE
~ rn n •-. ?
-
•
*
*
25 1 24
S.WPROMON
f
1.
H
5 1 4 1 3 1 2
SEKACHINSKAYA BA^
,
•
I
r
.'•1
1
1
•m_
20 1 21 1 15 1 16 1 1?
f PALATPROM1LEVYAZHIYABAY
Figure 5. Number of bacteria in the coastal waters of the southeastern part of
Mednyy Island: a) east coast; b) west coast. Number of bacteria
(c/ml) o hydrocarbon-oxidizing and hydrocarbon-resistant; a sapro-
phytes; • oligocapnophilic; *number of bacteria less than 1 c/ml.
36
-------
The number of heterotrophs was generally small and did not exceed 200
c/ml at 13 stations; this indicated a very insignificant concentration of the
easily accessible organic substance in coastal waters. On the western coast
of the island, the number of heterotrophs was considerably higher than on the
east coast: It was over 5000 c/ml at 5 out of 11 stations. It was
particularly prolific at the stations near the seal and dotterel lairs
(stations 24 and 25 at the Southeast Cape and stations 20 and 21 at the Palat
Cape). This was probably due to the enrichment of the coastal waters by
organic substances as a result of the vital activities of these animals.
Studying the number of hydrocarbon-oxidizing bacteria was of more interest.
No petroleum hydrocarbons were found on analysis of the samples, although the
sensitivity of the recording device was relatively high—0.05 mg/liter.
Nevertheless, in using the method of counting in agar-treated medium
containing diesel fuel, we obtained high values for the number of
hydrocarbon-oxidizing and hydrocarbon-resistant bacteria: up to 20,000 c/ml
at some stations and below 200 c/ml at only eight stations. However, only 50%
of the strains isolated grew in the silica gel medium. This was also true of
the real hydrocarbon-oxidizing bacteria. The remaining bacteria isolated in
the agar-treated medium containing a petroleum product did not assimilate the
petroleum product and grew because of the organic substances contained in the
agar. In analyzing the data presented in Figure 5, we see that the number of
hydrocarbon-oxidizing bacteria isolated in the agar medium correlate with the
number of oligocapnophilic bacteria in the overwhelming majority of the cases.
A correlation between the hydrocarbon-oxidizing and heterotrophic bacteria was
observed less often and only when there was a large number of heterotrophs.
However, the number of hydrocarbon-oxidizing bacteria at a number of
stations remained significant and was in no way related to pollution. The
presence of these microorganisms in unpolluted coastal waters is due to the
variety of their food requirements, since we know that there are no highly
specialized forms in this group of organisms (ShlegeV 1972). The high
concentration of hydrocarbon-oxidizing bacteria at some stations could be due
to the following causes. It has been shown that a large part of microbial
enzyme systems responsible for oxidizing hydrocarbons from the paraffin series
are inducers (Rozanova 1975; Klug and Marcovetz 1971). Oxidized hydrocarbons,
the higher alcohols, aldehydes, and acids can be inductors. The formation of
lipid substances of this type could be related to the vital activity of the
seals and is very highly probable in this region. In addition, the copious
seaweed in the coastal waters of the island could also be a source of hydro-
carbons and their derivatives (Caparello and LaRock 1975). Hydrocarbon-
oxidizing bacteria are thus a normal component of the coastal waters of Mednyy
Island, and the large amount found at a number of stations is not an
indication of pollution from petroleum products.
Similar results were recently obtained in foreign studies of microflora
in the estuary of the Neuse River (North Carolina). It was found that a large
number of microorganisms capable of using hydrocarbons could be found even in
natural waters with comparatively low levels of pollution from petroleum
hydrocarbons (5-79-10-9 g/ml) (Buckley et aJL 1976).
37
-------
With respect to the results of the expedition's studies, it is necessary
to mention that the presence of petroleum-oxidizing microorganisms in regions
moderately polluted by petroleum and in unpolluted regions is not related to
the presence of petroleum products and consequently these microorganisms
cannot be used as a pollution indicator in these conditions.
PETROLEUM-OXIDIZING MICROFLORA IN MARINE ECOSYSTEMS
Microorganisms which assimilate petroleum products were isolated in pure
cultures and studied, and the most active forms were determined by species.
Pure cultures were obtained by inoculating water samples in agar-treated
medium containing diesel fuel or liquid paraffin or by the cumulative culture
method in medium containing 1% paraffin (a natural source of hydrocarbons and
energy). Petroleum-oxidizing hydrocarbons from the following regions were
investigated: the coastal waters of the Finnish straits, Arctic Seas from
Arkhangel to the Tiks Bay, the coastal waters of Wrangel I. and the
Komandorskiy Islands. All of these places were in communication and were
located in regions with cold and very cold climates (55-75°N latitute). The
waters in the Finnish and Yenisey Straits are characterized by pronounced
distillation. The concentration of salts in these regions varied from 2.3 to
37.5 g/liter.
As noted above, in inoculating the water samples in the agar-treated
medium containing a petroleum product, a significant number of microorganisms
which weakly assimilated or did not assimilate hydrocarbons grew in the
medium. Most of them lost their capacity to grow in the medium with the
hydrocarbons after a series of repeated inoculations. These bacteria are
apparently either totally incapable of assimilating hydrocarbons and grow
because of the organic substances in the agar, or are capable of oxidizing the
hydrocarbons in special conditions. In contrast to this, the cumulative
culture method made it possible to distinguish the active forms immediately.
The study of the cultural, morphologic and physiologic-biochemical
properties showed that microorganisms with a stable capacity to oxidize
aliphatic hydrocarbons and petroleum products are basically arthrobacteria and
saprophytic mycobacteria. Hydrocarbon-oxidizing actinomycetes and fungi were
found in an insignificant number. Mobile forms were only found in freshly
collected materials; they usually grew weakly in the medium containing diesel
fuel (a natural source of hydrocarbons and energy) and died after several
inoculations. No mobile forms were found when the cumulative culture method
wa^ used.
Petroleum-oxidizing arthrobacteria were widely distributed in all of the
regions examined and represented approximately half of the active microflora.
They all belonged to the A. ceroformans species (old name: Mycobacteriurn
ceroformans) described by Krasil'nikov et al_. (1971). The characteristic
features include: active growth in a medium containing paraffin and formation
of colorless colonies which fluoresce in transient light, formation of
filiform cells which rapidly decompose into short bacilli and cocci in MPA
medium + 7% glycerin, gram-variability, weak growth in media containing sugar,
38
-------
formation of large amounts of wax when grown in media containing hydrocarbons
>C14, absence of mycolic acids; the cells are typically oval or coccal in
shape and 1.0-0.8 x 0.8 pm in size.
The saprophytic mycobacteria were determined according to Krasil'nikov
(1949). They all went through the bacillus -» coccus -»• bacillus cycle in
developing. The Mycobacteriurn mucosum species was widely distributed, and
representatives of this species were found in all of the regions investigated.
The related species, Mycobacteriurn convolutum and M. planum, were respectively
isolated from water samples taken in the coastal waters of the Finnish straits
and the Komandorskiy Islands. The brightly stained forms are similar to the
species M. phlei and M. brevicale. Both species were encountered in all
examinations of the Arctic Seas, the latter species was also found in the
region of the Komandorskiy Islands. According to other classifications, the
saprophytic mycobacteria described belong to the group of Coryneform bacteria,
or Rhodococcus.
Arthrobacteria and mycobacteria thus play a leading role in the oxidation
of petroleum hydrocarbons in the northern seas. All of the petroleum-
oxidizing bacteria isolated could develop at temperatures below 10°, but to
different degrees. All grew well in both fresh and salt media. In laboratory
conditions, these microorganisms actively oxidized a petroleum product, and
the presence of 3% sodium chloride did not interfere in this process (Table
5).
TABLE 5. GROWTH OF ACTIVE PETROLEUM-OXIDIZING BACTERIA WITH VARIED SALINITY
Biomass, g/liter
Medium with diesel fuel Medium with petroleum
Strain
Mycobacteri urn mucosum
AR-25 (5 days)
Mycobacteri urn phlei
AR-18 (10 days)
Mycobacteri urn phlei
AR-19 (10 days)
Without
NaCl
2.6
2.6
2.6
+3%
NaCl
2.8
2.8
2.9
Without
NaCl
0.7
0.8
0.9
+3%
NaCl
0.5
0.4
0.8
Growth of all strains was significantly weaker in the medium containing
petroleum, and salinity had a negative effect in this case. In the flasks
where strain AR-25 developed, the petroleum film disappeared in the fresh
water after 2 days, and after 4 days in the salt water; in the flasks con-
taining strains AR-18 and AR-19, it disappeared after 3 and 5 days,
respectively.
39
-------
Consumption of petroleum products by Arctic mycobacteria was investigated
at low temperatures (Table 6).
TABLE 6. GROWTH OF MYCOBACTERIA IN A MEDIUM CONTAINING DIESEL
FUEL AT LOW TEMPERATURES (growth—20 days)
Culture Temperature
Organism
1.6C
5.2°
Mycobacteri urn mucosum AR-25
Mycobacteri urn brevicale MST-32
Mycobacteriurn phlei AR-18
Mycobacteri urn phlei AR-19
Strong
Medi urn
Medi urn
Strong
Strong
Strong
Medi urn
Strong
Assimilation of the petroleum product was slowed at low temperatures, but
most of the strains grew well at a temperature of 5° over 20 days, and two—M.
mucosum AR-25 and M. phlei AR-19--also grew at a temperature close to zero.
Mycobacteria and the forms related to them which have a high capacity to
decompose petroleum products in a model medium have been found by other
authors in waters which differed in climatic conditions: in the Odessa
Straits of the Black Sea (Krasil'nikov et a^. 1973); in the Atlantic Ocean on
the coast of New Jersey (Atlas and Bartha 1972); in different regions of the
Pacific Ocean (Soli and Bens 1972; Cundell and Traxler 1973). There is also
an extensive literature on oxidation of aliphatic hydrocarbons and petroleum
products by mycobacteria isolated from river water and soil. These include
representatives of the species which we found in the marine ecosystems.
Salinity is not a significant obstacle for the development of arthrobacteria
and mycobacteria in a marine environment. The data obtained indicate that the
active hydrocarbon-oxidizing microflora in marine ecosystems are not specific.
CONCLUSIONS
The studies that we conducted in the northern seas of the USSR and the
northwest part of the Pacific Ocean showed that microorganisms which oxidize
hydrocarbons are widely distributed in these regions. They are found in
particularly large quantities in places which are significantly polluted by
petroleum products (port, local spills)—up to 3500 c/ml, and also in places
enriched with organic matter from a non-anthropogenic origin (coastal waters
in regions containing colonies of marine animals and birds, overgrowth of
macrophytes)—up to 10,000 c/ml. The concentration of petroleum-oxidizing
bacteria is low in open waters where no petroleum pollution is visually
detectable (an average of 500 c/100 ml in the surface film) and is not cor-
related with the concentration of petroleum products. A large number of
hydrocarbon-oxidizing bacteria always accompanies a high level of petroleum
pollution, but the opposite is not always true, and a high number' of these
microorganisms is not an indication of pollution.
40
-------
One of the most widely distributed hydrocarbon-oxidizing microorganisms
is oligocapnophilic bacteria. At the same time, many marine oligocapnophilic
forms cannot oxidize petroleum products, but carry high (up to 1%) concentra-
tions in the environment. This is due to the fact that the method for
counting the petroleum-oxidizing bacteria, based on using agar-treated media
containing a petroleum product, produces results which are approximately twice
as high. We are currently developing a method in the laboratory for preparing
silica gel medium with a petroleum product which would allow counting natural
petroleum-oxidizing bacteria and would be suitable for field conditions. The
petroleum-degrading microorganisms isolated from sea water are arthrobacteria
and saprophytic mycrobacteria (Coryneform group). The latter group includes
brightly colored forms. The arthrobacteria are represented by the species
Arthrobacter ceroformans; a characteristic feature of this species is the
formation of large amounts of wax when liquid hydrocarbons are used. No forms
specific for defined living sites were found; the same species were found in
different regions. Marine petroleum-oxidizing arthrobacteria and mycobacteria
are inherently euryhaline.
REFERENCES
Atlas, R. M. , and R. Bartha. 1972. Degradation and mineralization of
petroleum by two bacteria from coastal waters. Biotechnol. Bioeng.
14:297.
Buckley, E. N., R. B. Jonas, and F. K. Pfaender. 1976. Characterization of
microbial isolates from an estuarine ecosystem: Relationship of hydro-
carbon utilization to ambient hydrocarbon concentrations. Appl. and
Environ. Microbiol. 32:232.
Caparello, A. M. , and P- A. LaRock. 1975. A radioisotope assay for the
quantification of hydrocarbon biodegradation potential in environmental
samples. Microbial Ecol. 2:28.
Cundell, A. M. , and R. W. Traxler. 1973. Microbial degradation of petroleum
at low temperature. Marine Poll. Bull. 4:125.
Dietz, A. S., L. J. Albright, and T. Tuominen. 1976. Heterotrophic
activities of bacterioneuston and bacterioplankton. Can. J. Microbiol.
22:1699.
Klug, M. J. , and A. J. Markovetz. 1971. Utilization of aliphatic hydro-
carbons by microorganisms. Adv. in Microbial Physio!. 5(1).
Koronelli, T. V. 1979. Absorption of hydrocarbons by microorganisms.
Uspeichi mikrobiologii 14.
Krasil'nikov, N. A. 1949. Detector of bacteria and actinomycetes. Moscow,
Izd-vo AN SSSR.
Krasil'nikov, N. A., L. N. Stepanova, T. V. Koronelli, and V. I. Duda. 1971.
A new species of paraffin-oxidizing mycobacteria. Mikrobiologya 40:1040.
41
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Krasil'nikov, N. A., A. V. Tsyban1 , and T. V. Koronelli. 1973. Uptake of
normal alkanes and crude oil by marine bacteria. Okeanologiya 13:877.
Rozanova, Ye.P. 1975. The enzymatic apparatus of hydrocarbon-oxidizing
microorganisms and models of mechanisms of hydrocarbon oxidation.
Uspekhi mikrobiologii 10(3).
Seki, H. 1973. Silica gel medium for enumeration of petroleumlytic micro-
organisms in the marine environment. Appl. Microbiol. 26:318.
Shlegel1, G. 1972. General Microbiology, izd. Mir, Moscow.
Sieburth, J.McN. 1972. An instance of bacterial inhibition in oceanic
surface waters. Mar. Biol. 11:98.
Soli, G., and E. M. Bens. 1972. Bacteria which attack petroleum in the
saline medium. Biotechnol. Bioeng. 14:219.
Walker, J. D. , and R. R. Colwell. 1976. Enumeration of petroleum-degrading
microorganisms. Appl. and Environ. Microbiol. 31:198.
42
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DEVELOPMENT OF PLANKTON ALGAE IN CONDITIONS OF
PETROLEUM POLLUTION
by
0. G. Mironov
Institute of Biology for Southern Seas,
AN USSR, Sevastopol1
Petroleum and petroleum products constitute one of the basic toxins which
currently enter sea and fresh water. Without exaggeration, we can say that
petroleum pollution is universal and the effect of other pollutants is exer-
cised on the background of petroleum pollution of the aqueous medium. Other
toxic substances often predominate in inland bodies of watei—rivers, lakes,
canals, etc., caused by the discharge of wastewater from different industrial
enterprises. At the same time, petroleum pollution is characteristically the
predominant type of pollution in oceans and seas, and marine flora and fauna
primarily encounter a "pure form" of this type of pollution. The petroleum
which enters the ocean is carried by currents hundreds and thousands of miles
away from the discharge site, enters the subsurface layers of the water,
accumulates in bottom residue, and thus affects all aspects of marine life.
In this respect, the effect of petroleum on the primary production chain in
the ocean is of special interest.
Galtsoff et al. (1936) found that the concentrations of petroleum
products in ports, lagoons and bottom residue unfavorably affect diatomaceous
and other algae. However, the petroleum film, in their opinion, does not kill
diatomaceous algae, but only disturbs their normal fission. For example,
Nitzschia closterium developed in an aqueous medium under a layer of different
types of petroleum as well as in the controls. A 12% concentration of crude
oil extracts had a stimulating effect on the growth of most diatomaceous
cultures; a 25% concentration inhibited growth, and normal development of the
algae stopped only with a 50% concentration. The authors concluded that both
the observations under natural conditions and the laboratory data indicated
that these organisms are damaged by prolonged exposure to large quantities of
petroleum.
Alfimov (1956) reported that Licmophora ehrenbergii died 1 day after a
0.1 mg/liter concentration of solar oil was added to sea water, while the same
concentration of solar oil increased the number of cells by 5 and 7.5 times in
5 days in Melosira monifliforim's and Grammatophora marina. ZoBell (1964),
citing a personal communication to Ruff Patrick, indicated that the petroleum
film does not destroy diatomaceous algae in the underlying water, but can
affect their normal reproduction.
43
-------
Lacaze (1969) tested a 1% extract of crude oil of the type spilled during
the TORREY CANYON catastrophe on the single-celled marine algae Phaeodactylum
tricornutum and obtained a 10% decrease in growth. Aubert et aj. (1969)
showed experimentally that different petroleum derivatives have a harmful
effect on the plankton algae Asterionella japonica. A 0.075 ml concentration
caused the cells to die rapidly.
Fractions of several types of petroleum which are soluble in sea water
and pollute the water off the coast of New York state exhibit different
degrees of toxicity for Phaeodactylum tricornutum, Sceletonema costatum,
Chlorella sp. and for the natural phytoplankton colonies in samples of sea
water, according to Nuzzi's observations (1973).
Laboratory experiments with mass species of plankton and benthoplankton
algae isolated from different seas were conducted in the Institute of Biology
for the Southern Seas, AN USSR (Mironov and Lanskaya 1967ab, 1968, 1969;
Mironov 1970; Roukhiyaynen and Mironov 1973).
As the concentration of mazut and kerosene in the sea water increased,
multiplication of the algae decreased, and with a concentration of 1.0
mg/liter, the cells died in the first days. Inhibition of growth and in many
cases the death of the cells were also observed with smaller concentrations of
the petroleum products. Concentrations of 0.001 and 0.0001 ml/liter kerosene
and mazut in sea water did not cause the algae to die during the experiment
(i.e., for 5 days). No significant differences were found in the development
of the experimental and control cells with these concentrations. In many
cases, small quantities of petroleum products even had a stimulating effect on
division of some types of phytoplankton.
In discussing the effect of petroleum pollution, we emphasize the dif-
ferences in the sensitivity of the individual species. Thus, Actinacyclus
ehrenbergii and Hyalodiscus sp. remained viable for 5 days with a 1.0 ml/liter
concentration of mazut, while Gymnodinium kovalevskii and Gymnodinium sp. died
when exposed to a 0.01 ml/liter concentration, i.e., 100 times less. Even
greater differences were observed in Ditylum brightwellii and Melosira
moniliformis. D. brightwellii died with a 0.01 ml/liter concentration of
kerosene in the first days, and in 3 days with a 0.001 ml/liter concentration.
At the same time, M. moniliformis remained viable after a 5-day exposure to
sea water containing a 10.0 ml/liter concentration of kerosene. In this case,
the difference in the sensitivity of the species to pollution of sea water was
3-4 orders of magnitude. Benthos and benthoplankton species were usually more
resistant to pollution of the water.
Exposure of microscopic algae to different types of petroleum showed that
despite slight differences in the chemical composition of the petroleum, its
effect on division of plankton cells was basically a function of the con-
centration of the substance in the water and not a function of the type of
petroleum used. A similar mechanism was observed both with species of algae
which weakly divided in the experimental conditions (twice the number of cells
in the control) and with intensively dividing species (5-6 and 30-fold
increase in the number of cells). This apparently indicates some common
feature in the effect of petroleum hydrocarbons on phytoplankton. ThTs natur-
ally forms no basis for suggesting that these types of petroleum have the same
*
44
-------
degree of toxicity, since this could only be determined by special studies.
Kauss et ah (1972) showed in laboratory experiments that aqueous extracts of
seven crude oils had an inhibiting effect on phytoplankton and differed in the
degree of toxicity. They also observed significant changes in the pH of the
—,j,-..m ,.,,-4.1, petroleum pollution which could inhibit growth of the algae by
medium with
itself.
The differences in the species sensitivity of algae to petroleum pol-
lution are shown in Table 1, which indicates that the cells die within a wide
range of concentrations, from 1.0 to 10-4 ml/liter. The absence of division
or inhibition of division in comparison to the controls was also observed as a
function of the species of algae with a 0.1-0.00001 ml/liter concentration of
petroleum. In these concentration ranges, the rate of division in a number of
species of algae did not differ from the controls. Of the ten species used
for the experiment, death or inhibition of cell division occurred in six
species with up to 10-2 ml/liter concentrations of petroleum and in two
species with 1.0-0.1 ml/liter concentrations. This indicates the species
sensitivity of most species of plankton algae isolated from different seas to
pollution of the sea water by crude oil.
Kauss et al. (1972) also observed a difference in the species sensitivity
of phytoplankton to petroleum. This phenomenon is more evident in natural
conditions, where a stimulating, indifferent, or damaging effect has been
observed in plankton algae.
TABLE 1. REACTION OF ALGAE TO DIFFERENT CONCENTRATIONS OF PETROLEUM
(concentrations expressed as mg/liter sea water)
Algae
Glenodinium foliaceum
Chaetoceras curvisetus
Gymnodi n i urn wu 1 f f i i
Di ty 1 urn bri ghtwel 1 i i
Gymnodi nium kovalevskii
Prorocentrum mi cans
Peri dim* urn trochoideum
Licmophora ehrenbergi i
Platimonas viridis
Coscinodiscus granii
Cells
Died
1.0-0.1
1.0-0.01
1.0-0.1
1.0-0.0001
1.0-0.0001
1.0
1.0
1.0
1.0
1.0
Absence of Division
or Inhibition of
Cell Division
0.1-0. 01
0.01
0.01-0.0001
0.001-0.0001
0.1-0.00001
1.0
0.1-0.001
0.01-0.001
1.0-0.1
Did not Differ
From Controls
0.001-0.0001
0.001-0.0001
0.00001
0.1-0.00001
0.0001-0.00001
0.0001
0.1-0.0001
In some conditions, low concentrations of Venezuelan crude oil can stimu-
late photosynthesis according to the data of Gordon et aJL (197?). On the
whole, the effect of the three types of petroleum studied (Venezuelan crude,
fuel oil No. 2 and fuel oil No. 6) on photosynthesis of colonies of natural
phytoplankton from the basin in Bedford, New Scotland (Canada) and the
northern part of the Atlantic Ocean (between Halifax and the Bermuda Islands)
45
-------
showed that all three toxins could inhibit photosynthesis. The current levels
of petroleum pollution in the Bedford basin are capable of decreasing photo-
synthesis by several percent.
Intermittent contact of algae with petroleum products also results in
inhibition of cell division and death of the cells. Exposing Prorocentrum
mi cans and Coscinodiscus sp. to sea water containing 1.0 ml/liter of mazut and
kerosene for one-half hour inhibited multiplication in these algae. In this
case, £. mi cans was found to be highly sensitive to mazut pollution. Thus,
after a 4-hour exposure to mazut, it died in 3 days in clean sea water, and
died in 1 day after a 6-hour exposure. At the same time, 6-hour exposure of
the same algae to the same concentration of kerosene only caused inhibition of
cell division.
M. moniliformis did not lose the capacity to divide after 1-day exposure
in sea water containing 10.0 ml/liter of mazut. Multiplication was not
observed in the first days, however, indicating that mazut has a definite
toxic effect on this alga. On the other hand, 5-minute contact of D.
brightwellii with sea water containing 1.0 ml/liter of mazut caused statistic-
ally significant inhibition of its growth after it was placed in clean sea
water. If this alga is left in sea water containing mazut for 1 hour, the
cells begin to die on the third day in clean sea water.
Adding mazut and crude oil to sea water in the concentration of 0.01
ml/liter resulted in a decrease in the number of generations of D.
brightwellii (Figure 1). Although 30% of the control cells produced ten
generations, the cells exposed to the petroleum and mazut produced seven and
six generations, respectively. Some cells stopped dividing in the second
generation and the number of dividing cells decreased by up to 30-40% in the
fourth to fifth generations.
Cell division proceeded somewhat differently in the presence of solar
oil. The number of dividing cells in the experimental and control groups was
initially about the same, while there were twice as many in the experimental
group at the end of the experiment, i.e., in the tenth generation. The dif-
ference in the chemical composition of the petroleum products used apparently
plays a role here.
Division of Coscinodiscus granii cells was approximately the same as
division of the first species. However, due to the great resistance of C.
granii, some of the cells generated up to ten generations in sea water
containing petroleum and mazut.
Studies of the effect of petroleum, solar oil and mazut in concentrations
of 0.001 ml/liter, 0.01 ml/liter and 0.1 ml/liter on the development of small
Flagellata algae (pirophytic-- Cryptomonas vulgaris, golden-- Ochromonas sp.,
green—Platymonas mediterranea) showed (Figures 2-4) that all three species
were characterized by a definite type of development both in the presence of
petroleum products and in the controls. On the whole, the experiments demon-
strated the clearly negative effect of petroleum products on small Flagellata
algae. This was very clearly traced in £. vulgaris and Ochromonas
46
-------
1000
500
0
1000
CO 500
_l
_J
8 °
Q 1000
NUMBER
Cl
o 8
1000
500
0
A
1
Inll.inArm QAI 1 ^>^n>\ry^
1 5 10 1 5 10
II
nnn||/\AA/^ /iClTI A/\
15 10 1 5 10
III
ll.n>vlnlll |llAAlnLl
15 10 1 5 10
IV
AJ-II-I fin • np-j 1 AA^
15 10 1 5 10
GENERATION
100
50
1 1 - • A>\A «
B
1
| ,, |Mi ,, i§|
1 5 10 1 5 10 1 5 10 1 5 10
100
A I
02 50
CO
_J
_J
UJ 0
o
u_ 100
o
a:
UJ 50
QQ
_Ari_j_i_1 ^
1 5 10
100
50
0
II
Al..i.Atll A.AAAA^lll |I.|AAA||
15 10 1 5 10 1 5 10
III
• AAA| AA^| A.AA||AAAI iAAAAABiAl
15 10 1 5 10 1 5 10
IV
IA.AA.I/VAI. |/\AAAtAy\l ..IAA/IA.I
15 10 1 5 10 1 5 10
GENERATION
jure 1. Maximum number of Ditylum brightwellii (A) and Coscinodiscus granii (B) cells in differer
generations. I-controls; II-petroleum; Ill-solar oil; IV-mazut; 1, single-celled; 2, under 20
cells.
-------
Cl
320
280
240
200
160
120
80
40
0
I
II III IV
PETROLEUM
Figure 2. Development of Crypto-
monas vulgaris in the presence of
petroleum.
Cl
20,000 r
10,000
7000
5000
3000
1200
800
400
0
n 2
• 3
The growth of the other species of
nanoplankton algae, Platymonas
tetrahele, in the presence of crude oil,
was investigated by Mommaerts-Billiet
(1973). A decrease in the growth rate
was observed with a concentration of
more than 50 mg/liter. The toxicity of
"fresh" and weathered petroleum was the
same. The author emphasized the hazard
of an even insignificant delay in growth
for the species, as this could ulti-
mately result in its elimination from
the ecosystem.
The studies conducted by Kustenko
on several species of Black Sea plankton
algae showed their different sensitivity
not only to petroleum in general, but
also to its individual fractions--film
and soluble. The petroleum film was
most toxic for most of the species. The
degree of toxicity was a function of the
lighting. With continuous light, the
effect of the hydrocarbons fell within
the range of the optimal levels of
lighting for each of the species of
algae. We know that cell division in
light decreases the intensity of photo-
B
IV I
PETROLEUM
II
IV
Figure 3. Development of Ochromonas sp. in the presence of petroleum.
48
-------
Cl
320
280
240
200
160
120
80
40
n
—
-
—
—
—
—
—
—
^
?
a
^J
\
r
J*
*
*
* •
'.;
A
E3 1
D 2
• 3
0 4
1 Sk,
L
•
•
•:
B
J3
|
r
•^
# «
*."»
n
a*
1 •
i £
-
-
-
—
—
-
-
—
^
h"
I II III I I
PETROLEUM
III
Figure 4. Development of Platymonas
in the presence of petroleum.
synthesis. However, certain
reserves of assimilates are even
created in this case. Enrichment of
the environment with oxygen also
takes place. As a result, the cells
can divide further and the condi-
tions are created for multiplication
of petroleum-oxidizing bacteria in
aerobic conditions. For this
reason, adding petroleum results in
enrichment of the medium with carbon
which can be utilized by the cells
for treating energetic materials.
The decrease in the toxic
effect of hydrocarbons observed with
an 8-hour period of light could be
explained in the following way:
First, with this lighting regime,
the rate of cell division signif-
icantly decreases. However, in this
case, the light energy is apparently
used more intensively by the cells
than in other regimes. Neverthe-
less, a significant decrease in the
light period does not provide suffi-
cient reserves of assimilates for
the vital activity of the cells in
the remaining part of the day in
conditions of total darkness. As a
result, attenuation of cellular pro-
cesses can begin, and the intensity
with which the pollutants partici-
pate in cell metabolism decreases.
Great differences were thus observed in the sensitivity of the individual
species of microscopic algae to pollution of sea water by petroleum and petro-
leum products. The lethal concentrations of these substances in sea water
differ by several thousand times for the individual species. Demonstrating
the toxic effect of petroleum products found in sea water on plankton and
benthoplankton algae isolated from different bodies of water forms the basis
for hypothesizing the presence of some common mechanisms in the effect of
petroleum pollution on algae.
We can hypothesize that when the water contains concentrations of petro-
leum products on the order of 10-4 to 10-5 ml/liter, the petroleum could
sometimes significantly increase the rate of cell division in some species of
microscopic algae (which can then be followed by no less significant inhibi-
tion). But the finding that individual species of phytoplankton multiply
under the effect of the pollution still cannot be used as a water quality
criterion.
49
-------
It is very possible that small quantities of some "wastes," including
hydrocarbons, in sea water result in a temporary increase in the cell division
rate which can subsequently be replaced by equally pronounced inhibition, and
possibly the death of the cells. The unfavorable reorganizations of the
biologic structure of a given body of water whose catastrophic consequences
are often observed in bodies of fresh water (in the blossoming period) are
very possible.
SUMMARY
At present oil and oil products are the major toxic substances entering
the marine environment. Of special interest is the impact of oil on the
primary production link in the sea. There is a great discrepancy in the
sensitivity of microscopic alga species to oil and oil products-polluted sea
water. The lethal concentrations of these substances vary for individual
species by several thousand times. Oil causes injury to the development of
plankton algae generations. By considerably diluting oil products until
concentrations of 10-4 to 10-5 ml/liter in sea water are reached, the oil may
increase the rate of division in some microscopic algae species. However, the
effect of pollution on reproduction of individual species of phytoplankton may
not serve as a useful pollution criterion.
REFERENCES
Alfimov, N. N. 1956. Use of cultures of diatomaceous algae for evaluating
the degree of pollution of sea water. Botanicheskiy zhurnal, 41(11).
Aubert, M. , J. Aubert, S. Daniel, and J.-P. Gambarotta. 1969. Study of the
effects of chemical pollution on plankton. Degradation of fuel by
telluric and marine microorganisms. Rev. Int. d'Oceanogr. Med.
Galtsoff, P. S. , H. F. Prytherch, R. 0. Smith, and V. Koehring. 1936.
Effects of crude oil pollution on oysters in Louisiana waters. Bulletin
of the Bureau of Fisheries.
Gordon, D. C., and N. J. Prouse. 1973. The effects of three oils on marine
phytoplankton photosynthesis. Mar. Biol. 22(4).
Kauss, P. R., T. C. Hutchinson, and M. Griffiths. 1972. Field and laboratory
studies of the effects of crude oil spills on phytoplankton. Inst.
Environ. Sci. and Educ. Proc. 18th Ann. Techn. Meet., New York.
Lacaze, J. C. 1969. Effects of pollution of the "Torrey Canyon" type on the
unicellular marine alga Phaeodactylurn tricornutum. Rev. Int. d'Oceanogr.
Med.
Mironov, 0. G. 1970. Division of some diatomaceous algae in sea water
containing petroleum products. Biol. Nauki, 7.
50
-------
Mironov, 0. G.
algae in
Distribution
and L. A. Lanskaya. 1967a. Development of some diatomaceous
sea water polluted with petroleum products. In_ Biology and
of Plankton in the Southern Seas. Moscow.
Mironov, 0. G., and L. A. Lanskaya. 1967b. Effect of petroleum products on
development of marine phytoplankton. Ir\ Topics in Oceanography. Naukova
dumka Press, Kiev.
Mironov, 0. G. , and L. A. Lanskaya.
and benthoplankton algae in sea
Bot. zhurn., 53(5).
1968. Survival of some marine plankton
water polluted with petroleum products.
Mironov, 0. G.
algae in
Processes
, and L. A. Lanskaya. 1969. Development of marine microscopic
sea water polluted with hydrocarbons. In Productive-Biologic
in Plankton in the Southern Seas. Naukova dumka Press, Kiev.
Mommaerts-Billiet, F. 1973. Growth and toxicity tests on the marine nano-
planktonic alga Platymonas tetrathele g. s. west in the presence of crude
oil and emulsifiers. Environ. Pollut. 4(4).
Roukhiyaynen, M. I., and 0. G. Mironov. 1973. Development of some small
marine flagellata algae in the presence of hydrocarbons. Hydrobiologic
studies of the northeast part of the Black Sea. Rostov-Don, Izdat.
Rostov. Univ.
ZoBell, C. E.
sea. Adv.
1964. The occurrence, effects and fate of oil polluting the
Wat. Pollut. Res., No. 3.
51
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LONG-TERM BIOLOGICAL VARIABILITY AND STRESS IN COASTAL SYSTEMS
by
Robert J. Livingston
Department of Biological Science
• Florida State University
Tallahassee, Florida
ABSTRACT
Coastal biological systems undergo long-term cyclic changes which are
variously influenced by short-term and prolonged climatic phenomena. The
biological reaction to physical stress is complex and does not necessarily
follow strictly linear response patterns. This non-linearity tends to
complicate the analysis of the impact of a given pollutant or pollutants. Our
research group has just completed a continuous, eight-year, multi-disciplinary
study of two very different bay systems in the Gulf of Mexico. This study has
concentrated on long-term changes of benthic assemblages including both plant
and animaT species. In addition to the usual physico-chemical and biological
measurements, we have concentrated on detailed trophic interactions of several
fish species in an effort to understand the nature of community changes over
short and long periods. Although much of the data base is still under
analysis, several trends are emerging. Often such features as macrophyte
standing crop and species richness of diverse groups of organisms are
temporally uncoupled. Individual species strategies are diverse enough to
complicate generalization based on cause and effect phenomena. The results
have led us to redirect our emphasis from the taxonomic species to the
"trophic unit," a group of organisms (usually a particular size class of a
species) which have common feeding habits. Overall, the trophic structure of
grassbed systems in the Gulf of Mexico does not follow the traditional
concepts of distinct trophic levels. An understanding of the impact of
habitat alteration and its effect on key units of a given trophic spectrum may
be critical to the understanding of system resilience and stability. However,
such an understanding will require analysis of actual overlapping successions
of plant and animal assemblages which result in observed, long-term changes in
coastal communities.
INTRODUCTION
The actual demonstration of cause and effect relationships due to
pollution is extremely difficult because of the complexity of coastal systems
and the natural variation in assemblages at various levels of biological
organization. These fluctuations reflect changes in key climatic features,
microhabitat distribution, water quality functions, and biological tnter-
52
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actions such as predator-prey relationships and competition. Temporal cycles
may have periods ranging from minutes to years. Despite the fact that long-
term cycles with periods of many years are not only possible, but probable,
most "long-term" studies of coastal systems do not exceed one year (Coull and
Fleeger 1977). Rarely is enough information gathered to establish causal
effects on the biological system. If specific hypotheses are to be tested
regarding a system's reaction to stress, the frequency and duration of the
sampling effort should be designed to fit its natural patterns of biological
variability. Otherwise, the approach becomes tautological. Inadequate
sampling can lead to contradiction and over-generalization. In dynamic
coastal systems, there is no simple way to eliminate background variation so
as to establish causal mechanisms.
Many studies have been made of long-term changes in diverse groups of
coastal organisms (Longhurst et aj. 1972; Williams 1972; McErlean et aj. 1973;
Coull and Fleeger 1977; Reid 1977; Bayley et aj. 1978). However, most such
studies are descriptive and without substantial insight into causal
mechanisms. Coastal biological systems do not usually have a linear response
to causal physico-chemical factors (Livingston et a_L 1978; Meeter and
Livingston 1978). The relationships of spatial variability (from system to
system) and response to catastrophic natural occurrences remain poorly
defined. Heavy rainfall (Boesch et aj. 1976; Saloman and Naughton 1977) and
temperature (Snelson and Bradley 1978) are often mentioned as important
driving factors, and have been hypothesized to direct long-term trends in
commercial fisheries data (Van Winkle et aJL 1979; Sutcliffe et a_l_ 1977;
Meeter et al. 1979). However, fisheries data are often affected by fishing
effort and/or socio-economic conditions (Clark and Brown 1977; Meeter et al.
1979), and ultimately do not represent a satisfactory substitute for
quantitative, multi-disciplinary, scientific data on long-term (supra-annual)
changes in coastal systems.
Despite a distinct lack of actual data, there has been considerable
theoretical discussion concerning the basis of stability (the ability of a
system to return to normal after periods of stress or perturbation) in natural
populations and communities. Orians (1974) recognized various aspects of
stability, and attempts have been made to express such concepts as mathe-
matical formations (Harrison 1979). There are numerous models which predict
functional mechanisms with the usual semantic diversity (Lewontin 1969;
Holling 1973; May 1973; Webster et aj. 1975). That so much modeling should
have been done with so little empirical testing is indeed, as McNaughton
(1977) has said, "fascinating." The central problem, of course, arises from
the extreme spatial and temporal variability of natural systems, especially
coastal ones. The assumption that the influence of environmental changes on
population variability is stochastic has been effectively questioned (Wiens
1977). Once again, from an entirely different viewpoint, the need for long-
term studies is emphasized. Long-term population data (relative to generation
time) is necessary because of the possibility of temporal variability in the
overlapping of resource utilization traits of individual populations. At
present, without these long-term data, the nature of impact due to stress
(natural or anthropogenic) and system recovery remains largely undefined
within the context of long-term changes in coastal systems.
53
-------
AREA OF STUDY
Apalachee Bay (Figure 1) is a shal'low bay system in the Gulf of Mexico
which receives runoff from a series of small rivers in north Florida. It is
characterized by relatively clear water (i.e. low turbidity and color), and
its biological systems are dominated by benthic macrophytes which serve as a
source of primary production and determine microhabitat distribution
(Livingston 1974). From 1971 through 1979, a comparison was made between two
of the rivers feeding Apalachee Bay, together with their offshore drainage
areas: the unpolluted Econfina (Livingston 1975; Zimmerman and Livingston
1976ab, 1979) and the polluted Fenholloway. The latter received pulp mill
effluents from 1954 to 1974. By changing the primary patterns of productivity
and microhabitat distribution through alteration of the benthic macrophyte
assemblages, the pulp mill effluents altered the trophic system drastically.
During late 1974, however, a pollution-control program was initiated which
eliminated considerable portions of the kraft pulp mill effluent. As a
result, during the final six years of the study, we were able to study the
system's capacity for return to its pre-pollution state.
METHODS
Because of similarities in water quality and flow rates to the
Fenholloway River (above polluted areas), the Econfina system was chosen as a
control. By comparison of the two systems, we hoped to determine the impact
of pulp mill effluents on the water quality and the offshore biota of the
Fenholloway, as well as to evaluate its subsequent recovery. Cognate stations
were set up in the two systems (Figure 1) to facilitate statistical analysis
of relative impact. Station pairs were chosen which did not differ in the
value of variables not affected by the pulp mill effluents, such as salinity
and temperature (Livingston 1975), so as to provide valid comparison for the
evaluation of impact due to pollution (as represented by high levels of
color). The data collection and analysis methods used for physio-chemical
(salinity, temperature, color turbidity, pH, dissolved oxygen, Secchi depth)
and biological (benthic macrophytes, epibenthic organisms) variables are given
by Livingston (1975), Livingston et a!. (1976), and Zimmerman and Livingston
(1976ab, 1979). The principal comparisons made were between single-station
pairs and between seven-station totals of monthly data for the period of
study.
RESULTS AND DISCUSSION
Long-Term Trends
Nearshore areas of Apalachee Bay are influenced by local climatic
features such as temperature and rainfall. Usually, rainfall in north Florida
is seasonal and major peaks occur during summer months. There are indications
that upland vegetation buffers the effects of this major input of runoff
through evapotranspiration, to the extent that inshore coastal systems undergo
only minimal changes in water quality functions such as salinity and color.
During 1973, however, there was a major rainfall during the winter and-rearly
spring period (Figure 2) which seems to have had a major influence on coastal
54
-------
Figure 1. The Apalachee Bay area showing permanent sampling stations in the
Econfina and Fenholloway drainage systems.
55
-------
en
cn
0
•WW*«A ^^
1| WINTER-SPRING
78
Figure 2. Total monthly rainfall (cm) in the Econfina-Fenholloway drainage area from January, 1971
through September, 1978.
-------
water quality (Figure 3). Presumably, the (relative) dormancy of the wetlands
vegetation reduced the buffering effect, resulting in the lowest salinities
and the highest color levels encountered in the Econfina system during the
seven-year period of sampling.
Simultaneously with the early spring rainfall of 1973, divers noted
die-offs of benthic macrophytes throughout the Econfina offshore area. Macro-
phyte biomass remained low during the following year (Figure 4). Turtlegrass
(Thalassia testudinum), the dominant macrophyte species, was particularly
adversely affected by the high runoff. Recovery by this species took four
years.
Long-term changes in number of macrophyte species, however, did not
follow the same trends as did biomass (Figure 5). The peak in numbers of
macrophyte species in the Econfina system occurred in 1974-75, the same point
at which standing crops reached seven-year lows. Fish species richness in the
Econfina area followed the same pattern; the peak in fish species numbers
occurred during the summer of 1974. The peak in this parameter in the
Fenholloway system occurred somewhat later, in 1975. The species richness
curves for the two systems converged in succeeding years. Numbers of indi-
viduals, both of all fisnes and of the numerically dominant species, pinfish
( Lagodon rhomboides), followed the same trends as did fish species richness
(Figure 6).
It is clear from these results that a single climatic shock, in the form
of an unusual pattern of rainfall, can have long-lasting effects on various
levels of biological organization. Because such rainfall patterns, and
possibly long-term trends in winter low temperatures, show evidence of follow-
ing distinct temporal patterns, an understanding of the timing of such
phenomena is essential to our comprehension of the forces controlling biolog-
ical variability in coastal systems.
In the present case, our long-term data base enabled us to conclude that
macrophyte biomass is not coupled with species richness, either of macrophytes
or of fishes. On the other hand, the close correlation between certain other
long-term biological trends in both plant and animal populations seems to
indicate some (direct or indirect) relationships between the controlling or
modifying factors and biological response. This possibility is currently
under analysis.
Trophic Structure of Fish Populations
The analysis of observed population and community changes, though impor-
tant as a first estimate of impact due to stress, does not answer functional
questions about dynamic interactions of various populations. The trophic
structure of the system may be extremely important in the determination of
population interactions and the impact of natural and anthropogenic stress on
marine systems. In an attempt to understand further the nature of the changes
caused by the addition of pulp mill effluents to one of the two rivers under
study, a four-year study of the trophic structure of the fish assemblages in
the Econfina and Fenholloway offshore systems was undertaken. The basis for
this study was established by Livingston (1978). The stomach contents of the
57
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70
c/)
h-
40
O
O
•
I-
CL
10
25
O
15
RAINFALL
PEAK
COLOR
E7
---- F9
SALINITY
i
1971 72
Figure 3. Salinity (parts per
taken monthly at stations E7
1978.
73 74 75 76 77 78
thousand) and water color (APHA Pt-Co units)
and F9 from June, 1971 through March,
58
-------
I
CO
CO
o
GO ,
H
cr
0.5
TOTAL - ALL SPP
ECON.
FEN.
THALASSIA
1972 73 74 75 76 77
78
Figure 4. Total dry weight biomass of benthic macrophytes taken at 7 stations
in the Econfina and Fenholloway drainage areas. Figures represent
multiple samples extrapolated to biomass m-2 at each station. Also
shown are total biomass figures for the dominant seagrass,
Thalassia testudinum.
59
-------
12
8
c/)
UJ
o
UJ
Q_
C/)
£
GO
30
20
10
ECONr
FEN.
PLANTS
1971 72 73 74 75 76 77 78
Figure 5. Total numbers of species of benthic macrophytes (gm m-2 at E8 and
F10) and fishes (21 2-minute trawl-tows at E7, E8, E10, and F8,
F10, Fll) taken monthly from June, 1971 through May, 1978.
60
-------
200
c/)
_J
<
Q
100
tr
LL)
GO
80
40
0
TOTAL FISHES
- PINFISH
FEN.
ECON.
1971 72 73 74 75 76 77
Figure 6. Total numbers of individuals of fishes taken by repetitive (49)
otter-trawl tows at 7 permanent stations in the Econfina and
Fenholloway offshore systems from March, 1971 through September,
1977. Also shown are total numbers of individuals of the dominant
pinfish, Lagodon rhomboides.
61
-------
20 numerically dominant fish species in the Econfina and Fenholloway offshore
areas were identified (to species wherever possible) and quantified according
to methods established by Carr and Adams (1972). The analyses included fishes
taken during monthly sampling trips from 1971 through 1977 at stations chosen
to allow spatial (station-specific) as well as temporal (short- and long-term)
comparisons.
Because many marine organisms pass through a series of developmental
stages whose ecological requirements vary considerably from one stage to the
next, the species may not be the appropriate taxonomic unit to use in the
establishment of the functional or causative ecological relationships. There-
fore, information on the size and the developmental stage of the fish was
included as a factor in the analysis. By studying the feeding habits of
fishes according to species-specific ontogenetic changes and categorizing the
resulting patterns by clustering techniques (Sheridan and Livingston 1979), we
were able to break down the taxonomic species into a series of functional or
trophic units. These units could then be analyzed within the context of
considerable background information regarding basic physico-chemical and
biological components of the system in an attempt to obtain some understanding
of the functional basis of this shallow coastal system. The relative impor-
tance of dominance, species richness, and components of the productivity
phenomena, all part of the community structure, could thus be analyzed within
the context both of a spatial/temporal comparison with control stations and of
long-term changes in a recovering system. Questions could be asked regarding
equilibrium, resilience, and stability over a time span long enough to be
relevant to the systems under analysis. Ultimately, these data will be
compared to a similar series of projects in the Apalachicola estuary
(Livingston 1976; Livingston et aJL 1974, 1977, 1978; Sheridan and Livingston
1979) in an attempt to test hypotheses related to spatial variability of two
disparate estuarine systems (in this case the turbid, river-dominated
Apalachicola estuary and the clear, benthic-macrophyte-dominated Apalachee Bay
system).
The central questions regarding system stability and the relation of
biological diversity to physico-chemical controlling functions are thus being
approached, in the present study, on a trophic basis, with an emphasis on
spatial and temporal variability. Trophic units (groups of organisms which
have common feeding habits), rather than taxonomic species, serve as the basis
for analysis.
The preliminary results, based largely on Stoner's (1979) study of
pinfish (Lagodon rhomboides), the dominant fish species in the Apalachee Bay
area, are as follows:
1. The numbers of individuals and species of macrobenthic organisms in
a given area in Apalachee Bay were directly related to the species
composition of benthic macrophytes in that area. The organisms
studied included suspension-feeding and carnivorous polychaetes and
epifaunal amphipods and polychaetes. Deposit-feeder and omnivore
numbers decreased with increased macrophyte standing crop, and
seasonal changes in amphipod numbers were primarily related to
reproductive mechanisms and relative abundance of fish predators.
62
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2. There were five ontogenetic stages in the trophic composition of the
pinfish, ranging from planktonic through carnivorous, omnivorous,
and herbivorous phases. This succession of stages was related to
morphological changes, with variation based on relative abundances
of prey and macrophyte density. Selection of specific prey species
was important and increased with macrophyte standing crop.
3. The key to an understanding of the trophic relationships of the fish
assemblages in Apalachee Bay was the organization of benthic macro-
phytes in the area. There was evidence that the "trophic species"
was functional as an ecological unit, casting doubt on established
trophic-level approaches and the assumptions inherent in the
specialist-generalist dichotomy.
SOME CONCLUSIONS
Analyses relating trophic phenomena to hypotheses concerning variability
and stability of coastal systems are still being tested. However, preliminary
results indicate that established theory relating impact due to stress to
community structure in coastal systems may need revision. Spatially diverse
estuarine systems tend to follow long-term temporal sequences of changes in
species composition which reflect seasonal and supra-annual progressions.
Such changes are controlled by catastrophic climatic events and long-term
patterns of key physical controlling features. These cyclic phenomena have
varying periodicities at different levels of biological organization and are
superimposed over each other in time so that long-term aspects of community
organization are complex and dynamic. Specific response to stress is based on
these relationships and an understanding of trophic organization is critical
to the establishment of causal phenomena. Such responses must be viewed
within the context of differential phase relationships, uncoupled biomass and
species richness indices, and ontogenetic changes in species ecological roles.
All such functions are influenced by specific changes in productivity, micro-
habitat distribution, and species strategies. The durations of the relevant
climatic and biological phenomena exceed the one- to three-year periods so
common in established ecological surveys. Traditional (theoretical) concepts
of stability and resilience often reflect naive and faulty assumptions based
on inadequate empirical data. Such speculation has led to the exalted status
of simplified models and indices such as species diversity which take no
account of the extent of natural variability in time and space. These models
and indices often involve a tacit assumption of equilibrium, a state which has
not been demonstrated to exist, and is not in fact probable. Multi-dimen-
sional variability at different levels of biological organization is simply
ignored in impact analysis and, often, the natural variation of a given system
exceeds that caused by a given pollutant, masking or exaggerating its effects.
Biological relationships are not necessarily linear and there is simply
too little understanding of functional mechanisms in marine systems to make
quantitative comparisons and generalizations possible. This fact should not
be a deterrent to further modeling and speculation, but investigators should
recognize that without adequate empirical data, mere hypotheses may be
advanced to the startus of broad generalizations. Models which are neither
applicable nor predictive are being developed without necessary background
63
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information. Consequently, there is little basis for hypothesis testing. In
other words, if marine ecology and pollution biology are to reach an accept-
able level of credibility, considerably more "dirty" work will have to be done
on natural history and basic ecological function. Hypotheses will have to be
fitted to system-specific cycles of variability and factors which are critical
to the functions of a given natural system will have to be identified before
adequate management can become a reality.
ACKNOWLEDGEMENTS
Data analysis was supported by EPA Program Element No. 1 BA025 under
Grant No. R-803339.
REFERENCES
Bayley, S. , V. D. Stotts, P. F. Springer, and J. Steenis. 1978. Changes in
submerged aquatic macrophyte populations at the head of Chesapeake Bay,
1958-1975. Estuaries 1:171-182.
Boesch, D. F. , R. J. Diaz, and R. K. Virnstein. 1976. Effects of tropical
storm Agnes on soft-bottom macrobenthic communities of the James and York
estuaries and the lower Chesapeake Bay. Chesapeake Sci. 17:246-259.
Carr, W. E. S. and C. A. Adams. 1972. Food habits of juvenile marine fishes:
evidence of the cleaning habit in the leatherjacket, Oligoplites saurus,
and the spottail pinfish, Diplodus holbrooki. Fishery Bulletin 70:1111-
1120.
Clark, S. H. , and B. E. Brown. 1977. Changes in biomass of finfishes and
squids from the Gulf of Maine to Cape Hatteras, 1963-74, as determined
from research vessel survey data. Fishery Bulletin 75:1-22.
Coull, B. C. , and J. W. Fleeger. 1977. Long-term temporal variation and
community dynamics of meiobenthic copepods. Ecology 58:1136-1143.
Harrison, G. W. 1979. Stability under environmental stress: resistance,
resilience, persistence, and variability. Amer. Nat. 113:659-669.
Holling, C. S. 1973. Resilience and stability in ecological systems. Ann.
Rev. Ecol. Syst. 5:1-24.
Lewontin, R. C. 1969. The meaning of stability. .In Diversity and Stability
in Ecological Systems, G. M. Woodwell and H. H. Smith (eds.), 13-24.
Livingston, R. J. 1974. The ecological impact of pulp mill effluents on
aquatic flora and fauna in north Florida: comparison of a polluted
drainage system (Fenholloway) with an unpolluted one (Econfina). Florida
Coastal Coordinating Council. 87 pp.
Livingston, R. J. 1975. Impact of kraft pulp-mill effluents on estuarine and
coastal fishes in Apalachee Bay, Florida, USA. Marine Biology 32-19-48.
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Livingston, R. J. 1976. Time as a factor in environmental sampling programs:
diurnal and seasonal fluctuations of estuarine and coastal populations
and communities. Ij] Symposium on the Biological Monitoring of Water
Ecosystems, J. Cairns, Jr. (ed.), ASTM STP 607:212-234.
Livingston, R. J. 1978. Multiple factor interactions and stress in coastal
systems: a review of experimental approaches and field implications. Iji
Marine Pollution: Functional Responses, W. B. Vernberg and F. J.
Vernberg (eds.). Academic Press, N.Y.
Livingston, R. J. , R. L. Iverson, R. H. Estabrook, V. E. Keys, and J. Taylor,
Jr. 1974. Major features of the Apalachicola Bay system: physiography,
biota, and resource management. Florida Scientist 37:245-271.
Livingston, R. J. , R. S. Lloyd, and M. S. Zimmerman. 1976. Determination of
adequate sample size for collections of benthic macrophytes in polluted
and unpolluted coastal areas. Bull. Mar. Sci. 26:569-575.
Livingston, R. J. , N. Thompson, and D. Meeter. 1978. Long-term variation of
organochlorine residues and assemblages of epibenthic organisms in a
shallow north Florida (USA) estuary. Marine Biology 46:355-372.
Livingston, R. J. , D. C. White, and R. L. Iverson. 1977. Energy relation-
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66
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SOME FEATURES OF THE BIOLOGIC EFFECTS
OF POLLUTANTS ON MARINE ORGANISMS
by
S. A. Patin
All-Union Fisheries and Oceanography
Scientific Research Institute
Moscow
Extensive data have been collected on the effect of pollutants on many
species of marine organisms. The variety of this material, the difference in
methods, procedures and experimental conditions all complicate general inter-
pretation of the vast toxicological data base which increases every year.
Several attempts have been made; the results of one will be presented in this
paper.
We previously emphasized the characteristic tendency toward an increase
in the intensity (coefficient) of accumulation of radioactive and chemical
microcomponents in marine organisms as the size of the hydrobionts decreases
and, consequently, the size of their surface and contact with the water
increases (Patin 1971). If this is true, then the stable differences in the
degree of accumulation of toxic microimpurities in the organisms of different
sized groups should result in corresponding differences in the formation of
the dose loads of the toxic substances in the biomass of the organisms in
these groups and, consequently, to different effects of their reaction with
the same initial concentrations of harmful substances in the water.
We used the data from the literature (over 300 published works) to test
this hypothesis; a summary of the results was presented in our paper (Patin
1979). The values of the logarithms of the mean lethal concentration (LC50)
in 48-96-hour experiments for each of the most common toxic substances,
including mercury, copper, lead, cadmium, zinc dissolved in petroleum products
and organochlorine compounds, were plotted graphically as a function of the
logarithm of the mean size of the hydrobionts. One of these graphs is shown
in Figure 1.
An evident decrease was found in the values of the LCs0 as the mean size
of the hydrobionts representing the basic groups of the biotic population in
the world's oceans decreased, including the simplest plankton and benthos,
mollusks, worms, and fish in different stages of ontogenesis. The correlation
coefficients for the LC50 and mean dimensions of the organisms were equal to
67
-------
o:
z>
o
cc
I
O)
T-- 2
co \
II O>
o
0
y=0.92x-1.56
n = 32, r= + 0.87
A
A
4
'A
I
A A
O 1
02.
03
A4
A5
Figure V.
2345
LOGARITHM OF MEAN SIZE (jum)
Correlation of the toxicity of mercury (Hg2 ) for marine hydro-
bionts with their size: 1,3-data from the literature; 2,4-data
from the present study; 1,2-adult organisms; 3,4-early stages of
ontogenesis; 5-value of the EC50 for unicellular algae; n-number of
data; r-correlation coefficient; data which were not considered in
calculating the regression equation are delimited by the broken
line.
0.87 for mercury, 0.78 for copper, 0.88 for zinc, 0.84 for cadmium, 0.87 for
lead, and 0.73 for dissolved petroleum products, 0.91 for polychlorinated
biphenyls and 0.47 for all groups of organochlorine substances.
In determining these correlations, we did not consider the data relating
to unicellular algae, since these data do not reflect the survival rate, as in
animal experiments, but the relative change in the bioproduction indices and
the cell division rate. In addition, note the correspondence of these results
with respect to the general tendency toward a decrease in the resistance of
hydrobionts to toxins as their dimensions decrease (Figure 1). The data on
bacteria were also excluded due to the small number and the difficulty in
comparing them with other materials.
The results obtained form the basis for attributing the size of hydro-
bionts which reflects the degree of development of the area of contact with
the environment and the capacity to accumulate contaminants to one of the
basic factors which predetermine the intensity and nature of toxic effects in
marine ecosystems. In each concrete case, the different representatives of
the marine population with the enormous variety of physiological, biochemical,
morphological, trophic and other features of the existence and vital activity
in different biotopes, naturally have their own specific features with respect
68
-------
to their reaction to any effect, including toxic effects. If we see an
obvious correlation between the toxic effect and the size of the hydrobionts
on this background, we must assume that the tendency is sufficiently universal
and reflects objectively existing relationships and phenomena in the complex
picture of the reaction of marine organisms to toxic factors in the environ-
ment.
In addition to the reasons for the low resistance of small species and
forms noted above, we also note the fact that organisms in different stages of
development are also included in this category: As all of the current toxico-
logical data, including the data obtained in our laboratory indicate (Patin et
aj. 1978), these forms are significantly more sensitive than the adult
specimens. The question of the nature and mechanism of this interesting
phenomenon is still debated, although it would be useful for solving a number
of problems in aquatic toxicology.
A particularly strong correlation between the LC50 and the size of the
organisms is characteristic of metals, while the correlation is not as strong
for dissolved petroleum products and organochlorine compounds. The reason for
this probably lies in the known multicomponent nature of organic toxins and
the difficulty of obtaining comparable data in conducting experiments with
these substances. The variations in the experiments with organochlorine
compounds, which include DDT and its metabolites, polychlorinated biphenyls,
aldrin and other substances which differ in composition and properties, are
particularly significant. This is the reason the correlation between the LC50
and size of the hydrobionts was weaker for all organochlorine compounds than
for the polychlorinated biphenyls alone: The correlation coefficients were
0.47 and 0.91, respectively.
Extrapolating the experimental results and conclusions to natural con-
ditions is difficult. Strictly as a first approximation, we can hypothesize
that phytoplankton, microzooplankton filter feeders, eggs and larvae of
benthos and nekton animals, especially in hyponeuston, are among the groups
which are most sensitive and react most rapidly to toxic pollution. We can
also hypothesize that the most rapid structural and functional disorders
should arise in associations of plankton and hyponeuston where the small
species and forms of hydrobtonts with low resistance to toxins predominate.
Similar disorders probably begin more slowly in associations of nekton and
benthos where large forms predominate, but they are more stable in nature.
A more definite conclusion can be drawn with respect to the prospects for
practical utilization of small species and forms of hydrobionts as test
objects for experimental toxicologic control of the hazard of waste waters,
their components and other substances which enter the seas. The relative
simplicity and ease of working with such species (especially with unicellular
algae, the simplest, smallest plankton crustaceans), the possibility of year-
round cultivation, the high sensitivity and rapidity of the reaction to toxic
contaminants in the environment--these and other advantages raise the hope
that small species of marine organisms will be widely used in determining the
danger of waste water from different enterprises and monitoring the quality of
natural sea water.
69
-------
SUMMARY
On the basis of data cited in scientific literature, the existence of a
negative correlation was shown between the rate of sensitivity (based on 48-96
hr LC50) of marine organisms of various systematic groups to the major pollut-
ants in the marine environment (dissolved petroleum products, organo-chlorine
intoxicants, mercury, lead, cadmium, copper, zinc) and the size of hydro-
bionts. This conclusion is of particular interest for evaluating biological
damage due to pollution in the marine environment and also links theory with
the practice of use of minor species and forms of hydrobionts in sewage water
biotesting.
REFERENCES
Patin, S. A. 1979. Effect of pollution on biologic resources and produc-
tivity in the world's oceans. Moscow, Pishchepromizdat.
Patin, S. A., V. K. Dokholyan, N. S. Chernyshev, A. M. Akhemdov. 1978.
Toxicology of some species of Caspian and Atlantic fauna. Tr. VNIRO,
vol. 134.
70
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FISH AS AN OBJECT FOR MONITORING PETROLEUM
POLLUTION OF THE MARINE ENVIRONMENT
by
N. D. Mazmanidi, G. I. Kovaleva, and A. M. Kotov
Georgian Branch of the VNIRO
In monitoring pollution of the marine environment, an important place
must be assigned to fish as an integral link in the trophic chain of the
marine bioecosystem.
The position could be based on a number of hypotheses. First, fish are
the most highly organized aquatic animals with a very differentiated nervous
system (not including marine mammals, of course). Second, fish are more
sensitive to toxic substances, particularly petroleum, than other aquatic
organisms (Stroganov 1976; Rice et aJL 1976). Third, it is not only possible
to conduct analytic studies on the molecular level in fish, but also to
observe all symptoms of poisoning (coordination, behavior, color, respiration,
etc.).
In selecting representative organisms, we departed from the principle
that experimental toxicological studies should be conducted not only for the
purpose of determining admissible levels of pollution in and evaluating the
quality of an aquatic medium, but also as a component part of any ecotoxico-
logical prognosis. In this respect, the object (or objects) to be investi-
gated should be a constituent element of the ecosystem in the region under
consideration. Stating the problem in this way predetermined our approach to
selecting representative species of fish for the southeastern and eastern
parts of the Black Sea. For this reason, before beginning the toxicological
experiments on fish, we decided to consider a series of questions:
a) Which species of fish were most common in the zone under investi-
gation and what were the dynamics of catching them?
b) What was the basic weight-size composition of the captured
specimens?
c) How would the species adapt to long periods of aquarium living?
d) How easy is it to distinguish between males and females (for correct
grouping of specimens in experiments and for obtaining mature sexual
products and working with them in different stages of ontogenesis)?
e) What is the sensitivity of these species to toxic chemicals?
71
-------
The fish were caught with fixed nets and divided into three groups:
Frequently found in large numbers;
Infrequently found in small numbers;
Individual specimens.
The representatives of the second and third groups could not be used for
our toxicological studies because of the small number found. For this reason,
we concentrated on the first group. It included:
Horse mackerel - Trachurus mediterraneus ponticus
Smarida - Spicara smaris
Mullet - Mullus barbatus ponticus
Crucian carp - Diplodus annularis
Sole - Solea lascaris nasuta
Plaice - Platichthus flesus luscus
Each of these species had to meet two important requirements: incidence
and size-weight composition.
The final conclusion was drawn after orientational experiments, acute and
subacute, were conducted on the survival rate and the sensitivity of these
species to dissolved petroleum products had been determined.
The experiments showed that pelagic active species of smaris and horse
mackerel were the most sensitive, while sole and plaice were the most resis-
tant benthoic species. The other two species occupied an intermediate
position. However, the horse mackerel did not survive in aquarium conditions.
For this reason, smaris and sole were selected for the toxicological studies.
The toxicological results for smaris should be used to establish the MAC
for petroleum products and the data on sole should be used for ecotoxicolog-
ical prognoses.
The method used in the experiments has been described in detail in our
previous studies (Mazmanidi et al. 1972; Mazmanidi and Kovaleva 1972;
Mazmanidi et al. 1975). We note only that the experiments were conducted with
dissolved petroleum products, and that we made certain that the experimental
materials and experimental conditions were compatible. In addition to the
survival rate, a number of morphological and biochemical indices for the
blood, carbohydrate metabolism in the organs and tissues, and the histopatho-
logic picture were selected as toxicity criteria. We departed from Federov's
conception (1976): First, these indices to a great degree ensure homeostasis
in fish, and second, changes in these indices are nonspecific response
reactions to very different factors.
We emphasize that the toxicological studies were preceded by many years
of work on the seasonal changes in all of the indices considered for the
purpose of determining the lower and upper limits for basal data for use in
interpreting the results obtained in the experiment. We also note that we
concentrated on conducting chronic experiments with low concentrations of
dissolved petroleum products.
72
-------
When considering poisoning in fish from dissolved petroleum products, it
is first necessary to discuss the symptoms of toxicosis.
Disorders in the central nervous system were observed in acute (25-29
mg/liter concentration of dissolved petroleum products) and subacute (15-19
mg/liter) experiments on active smaris. These disorders were manifested by
different degrees of excitation with subsequent inhibition of motor activity.
Narcotic and paralytic symptoms predominated in the less mobile benthic
fish in acute and subacute poisoning.
The investigation of the three response reactions (primary, anodic,
electroshock) in an electrical current field in smaris and plaice poisoned
with petroleum products confirmed the clinical symptoms: The threshold of
these reactions was lower in smaris than in plaice (Mazmanidi and Balayev
1974; Balayev et al_. 1976). Local pathological processes were simultaneously
observed in the fish in acute poisoning, in addition to the CNS disorders. In
addition to significant affection of the gill epithelium (hemorrhaging, edema)
observed on pathological-anatomical autopsy, histological studies revealed
some disorders in the structure of the organs and tissues (Mazmanidi 1974:
Mazmanidi and Zambakhidze 1974).
There were also disorders in production of mucus. Hypothesizing that the
skin, as well as the gills, are one of the barriers whose rupture allows
petroleum poisoning to begin, we investigated the effect of dissolved petro-
leum products on a number of mucus indices for the skin and olfactory bulbs of
smaris. In the course of petroleum toxicosis, significant changes were
observed in the concentration of total protein and the ratio of protein
fractions in mucus from the olfactory lining and the isoenzymatic protein
fraction picture and peroxidase activity in mucus from the skin of the fish
(Korolev et al_. 1978).
Asphyxia was one of the major symptoms of acute poisoning. As our
studies showed (Mazmanidi 1977), the clinical signs characteristic of this
symptom were closely correlated with a sharp decrease in blood oxygen satura-
tion (Table 1).
TABLE 1. OXYGEN SATURATION OF BLOOD IN ACUTE AND SUBACUTE POISONING BY
DISSOLVED PETROLEUM PRODUCTS (in %).
Fish N 1-10 h 24 h 48 h 72 h 96 h
Smaris
Sole
10
12
8
10
65.5
45.0
75.0
65.0
56.5
40.0
—
—
72.0
72.0
52.0
38.0
— — —
—
68.0
58.0
Note: The numerator indicates the controls; the denominator indicates the
experimental fish; n = number of specimens examined.
73
-------
As the above data indicate, this process is slightly more pronounced in
the smaris. This also indicates the high sensitivity of pelagic fish to the
toxic components in petroleum in comparison to benthic fish.
The decrease in blood oxygen saturation is related to a disorder in blood
hemoglobin metabolism, as the blood is the principal supplier of energy for
metabolic processes in fish (Table 2).
To briefly summarize the data presented in these tables, we can conclude
that anoxia, clinically manifested by asphyxia and functionally by a decrease
in blood oxygen saturation and disorders in the metabolsm of the different
hemoglobin fractions with pronounced meth- and sulfhemoglobinemia, occupies an
important position in the pathogenesis of acute and subacute poisoning from
petroleum toxicoses in fish.
In reference to the morphological and biochemical blood indices, leuko-
cytosis was observed in acute and subacute petroleum poisoning; it was within
the limits of 2a in the sole and reached 3a in the smaris (Table 3). The
same concentrations of petroleum also caused a change in the qualitative
composition of the white blood, manifested by monocytopenia and lymphocyto-
penia; it exceeded 2a in the smaris, and was manifested by neutrophilic
leukocytosis in the sole.
The changes in the red blood cells were characterized by oligochromemia
and erythropenia.
The experiments conducted with small concentrations of dissolved petro-
leum products showed that a 0.05 mg/liter concentration does not substantially
affect the blood of benthic, less mobile marine fish (sole), while only a 0.02
mg/liter concentration is not toxic for demersal forms (smaris).
In addition to the quantitative red and white blood indices, the effect
of petroleum on the qualitative composition of the blood was also studied.
Acute and subacute poisoning resulted in significant disorders in the form and
composition of the cellular elements in the blood.
Normoblasts and polychromatophilic erythrocytes were often found in the
blood smears. There was a significant number of immature erythrocytes with
unevenly stained cytoplasm, indicating irregular distribution of hemoglobin
inside the cells. The agglutination property was very pronounced. An
increase in the relative number of immature erythrocytes in the blood of
smaris in acute poisoning was maximum, and consisted of 455 cells per 100
fields of vision. The figure was 89 cells with a 1.0 mg/liter concentration;
47 with a 0.5 mg/liter concentration; and 26 cells with a 0.02 mg/liter
concentration. The average number of cells in the control fish was 24. In
addition to the above changes, signs of primitive cell-hemocytoblasts (up to
2.6%) were also observed in the blood smears. Vacuolization of the cytoplasm
was observed in the white blood cells. The changes observed in the quali-
tative composition of the red and white blood in marine fish in petroleum
poisoning were nonspecific.
74
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TABLE 2. FRACTIONAL COMPOSITION OF BLOOD HEMOGLOBIN IN FISH IN POISONING BY DISSOLVED PETROLEUM PRODUCTS
en
Type of
Experiment
Acute
Smari s
Acute
Sole
Subacute
Smari s
Duration of
Experiment N
10 h 10
10
24 h 10
10
24 h 10
10
72 h 10
10
C Met
mEq/liter
0.0010
0.0040
0.0003
0.0010
0.0010
0.0030
0.0005
0.0020
Hb
C Total
Hb
-SHb
% mEq/liter
2.0
6.0
0.9
4.2
1.5
5.3
1.3
3.4
0.
0.
0.
0.
0.
0.
0.
0.
051
063
032
023
066
055
039
059
C
Hb02
mEq/liter
0.
0.
0.
0.
0.
0.
0.
0.
050
059
031
022
065
052
038
052
C SHb
mEq/liter %
0.
0.
0.
0.
—
0030 4.5
_-_
0010 4.2
—
0008 1.4
—
0005 0.8
Total
Pigment
mEq/liter
0.051
0.066
0.032
0.024
0.066
0.056
0.039
0.059
Total
Inactive
Pigment
%
2.0
10.5
0.9
8.4
1.5
6.7
1.3
4.2
Note: The numerator indicates the controls; the denominator indicates the experimental fish; n = number
of specimens examined.
-------
TABLE 3. HEMATOLOG1CAL INDICES IN ACUTE AND SUBACUTE PETROLEUM POISONING IN
FISH
Hb
g %
+3 16.3
+2 14.3
+1 12.3
M 10.3
-1 8.3
-2 6.3
-3 4.3
a 2.0
+3 8.3
H2 7.7
H 7.1
M 6.5
-1 5.9
-2 5.3
-3 4.7
a 0.6
Millions
/mm3
2.697
2.478
2.259
2.040
1.821
Thou-
sands
/mm3
171350 )
141800/y
li
11225D7
It
82/00
53150
i
1 . 602 1 ^23600
I//
.1.383 "I/ —
1 *
0.219
2.388
2.137
1.886
1.635
29550
171390
142560 /
/
11373(//
8496
-------
The disturbances in carbohydrate metabolism in the first phase of poison-
ing were manifested by pronounced hyperglycemia, an increase in the amount of
glycogen, and a decrease in the lactic acid in the blood (Table 4).
TABLE 4. CHANGES IN CARBOHYDRATE METABOLISM INDICES IN THE BLOOD OF SMARIS
AND SOLE IN ACUTE POISONING BY DISSOLVED PETROLEUM PRODUCTS
Time,
Hours
No.
of Fish
Examined
Glucose, mg%
M + m
Glycogen, mg%
M ± m
Lactate, mg%
M ± m
SMARIS
Starting
Point
2
4
6
8
10
Controls
Time,
Hours
Starting
Point
1
2
3
Controls
8
8
8
8
8
8
8
12
14
12
13
14
55.0 ± 6.0
71.6 ± 7.4
109.7 ± 5.6
79.2 ± 4.4
49.3 ± 5.2
26.2 ± 4.8
61.2 ± 4.2
SOLE
«J \J L_ L.
30.2 ± 3.3
41.0 ± 1.1
54.8 ± 7.7
24.4 ±2.4
35.1 ± 1.7
35.8 ± 2.8
47.4 ± 4.2
55.9 ± 1.1
54.2 ±2.2
51.1 ± 3.4
48.2 ± 1.9
29.4 ± 3.5
33.5 ± 4.4
61.9 ± 3.7
53.5 ± 3.5
50.5 ± 5.0
43.7 ± 3.6
76.7 ±8.9
43.6 ± 7.1
30.2 ± 5.9
36.2 ± 3.1
48.7 ±3.6
60.6 ±10.0
73.5 ± 7.3
25.0 ± 3.9
16.1 ± 3.6
17.4 ± 2.6
28.0 ± 1.5
39.5 ± 2.7
The significant hyperglycemia in the first half of the experiment indi-
cated intensive glucose transport to organs and tissues, primarily to the
central nervous system, as the most important energetic substrate. The simul-
taneous hypolactacidemia was probably the result of decreased glycolysis.
Accumulation of glycogen in whole blood cells apparently promotes its
synthesis from plasma glucose by white blood cells. The second part of the
experiment was characterized by pronounced hypoglycemia and a decrease in the
concentration of glycogen in the blood of the poisoned fish. In this case,
the organism responds to the effect of the toxic factor by progressive deter-
ioration in its condition on all levels.
The chronic experiments were conducted with a wide range of concentra-
tions of dissolved petroleum products; the 0.01 mg/liter dose caused no
significant changes in the carbohydrate metabolism indices in the blood of
77
-------
smaris. Carbohydrate metabolism in the tissues of the fish also changed sig-
nificantly in poisoning from dissolved petroleum hydrocarbons. There was a
progressive decrease in carbohydrate reserves in all tissues studied in acute
poisoning in smaris (Table 5).
TABLE 5. CONCENTRATION OF GLYCOGEN (M ± m mg%) IN SMARIS TISSUES IN ACUTE
POISONING BY DISSOLVED PETROLEUM PRODUCTS
Hours
Controls
2
4
10
N
14
14
14
15
Glycemia
mg %
59.9 ± 6.3
*71.3 ± 8.6
*82.4 ±1.2
—
Liver
625 ± 73
*390 ± 63
*380 ± 60
*365 ± 64
Heart
440 ± 39
347 ± 35
*168 ± 25
*261 ± 47
Red
Muscles
251 ± 41
170 ± 23
*153 ± 19
*130 ± 26
Brain
35 ± 1
36 ±11
20 ± 3
—
Note: * indicates significant results p < 0.05.
The degree of these changes in the different tissues was almost ident-
ical: approximately 50% of the control values, i.e., the disorders in
glycogen metabolism in the liver, heart and red muscles and brain were similar
in nature. The changes in the concentration of polysaccharides in the dif-
ferent tissues were similar - the fundamental disorders in glycogen metabolism
began in the first hours of contact with the toxic substances and gradually
increased in time. When sections of intoxicated liver were
saline solution, a slight decrease was observed in production
comparison with intact sections. Acute poisoning also caused
glycogen in the liver and production of glucose by sections of
sole, and the latter process was very pronounced.
incubated In
of glucose in
a decrease in
this organ in
The decrease observed in the concentration of glycogen in the livers of
the poisoned fish, accompanied by a decrease in glucose production by the
sections, was apparently due to processes for detoxifying petroleum hydro-
carbons which principally occur in the liver. The glucose formed from the
glycogen is used as a source of energy for detoxification reactions and also
as a substrate for metabolic conversions of the petroleum products.
Subacute poisoning of smaris.and sole resulted in different disorders in
glycogen metabolism in the various tissues. A hepatotoxic effect was found in
smaris - a more significant decrease in glycogen in the liver in comparison to
the other tissues. Glycogen metabolism in heart muscle was primarily and more
significantly disturbed in sole, and inhibition
was observed only after seven days of exposure.
experiment, sections of liver from both species
icant decrease in glucose production, but cardiac
exhibited increased consumption of sugar from the incubation medium in compar-
ison to the controls.
of hepatic glycogen function
Under the conditions of this
of fish exhibited a signif-
muscle from the sole instead
78
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The experiments with the highest concentration which caused chronic
toxicosis in smaris (9 nig/liter) were conducted in different seasons of the
year: spring-summer, which corresponded to the pre-spawning and beginning of
the spawning periods, and in the autumn, when smaris are in the post-spawning
period. In these conditions, the hepatotoxic effect of the petroleum products
was primarily manifested (Table 6).
TABLE 6. CONCENTRATION OF GLYCOGEN (M±m, mg%) IN SMARIS TISSUES IN CHRONIC
POISONING BY DISSOLVED PETROLEUM PRODUCTS (9 mg/liter)
Exposure
Time
(days)
Liver Heart Red Muscle White Muscle
SPRING-SUMMER
7
15
20
25
508 ± 15 (8)
*292 ± 35 (19)
242 ±71 (11)
176 ± 29 (17)
242 ±71 (11)
183 ± 36 (10)
163 ± 38 (6)
*41 ± 11 (8)
297 ± 39
267 ± 24
232 ± 17
247 ± 20
232 ± 17
248 ± 37
74 ± 23
33 ± 7
149 ± 13
140 ± 13
93 ± 15
116 ± 12
93 ± 15
69 + 6
38 ± 8
43 ±10
55 ± 8
71 ± 4
41 ± 3
*60 ± 7
41 ± 3
61 ± 7
35 ± 5
*60 ± 3
AUTUMN
5
15
20
25
392 ± 87 (6)
*115 ± 26 (9)
427 ±84 (6)
*161 ± 15 (6)
427 ± 84 (6)
*174 ± 14 (6)
488 ± 84 (6)
*156 ± 36 (6)
286 ± 33
252 ± 28
319 ± 31
386 ± 54
319 ± 31
366 ± 61
359 ± 13
*470 ± 45
79 ± 5
91 ± 18
80 ± 11
*123 ± 11
80 ± 11
130 ± 14
70 ± 11
92 ± 17
60 ± 7
*35 ± 5
90 ± 6
80 ± 5
90 ± 6
67 ±12
51 ± 8
52 ± 3
Note: The number of specimens examined is indicated in parentheses;
* Indicates the significance of the results p<0.05.
In both variants (spring and autumn), the greatest decrease in the con-
centration of glycogen was observed after 5-7 days of exposure. However, the
decrease in the polysaccharide reserves in the liver of the poisoned fish
remained at approximately the same level in the autumn experiments--30-38% of
79
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the control values—during the entire experiment, while the polysaccharide
reserves in the poisoned fish did not significantly differ from the control
values on the 15th and 20th days in the spring variant. A pronounced decrease
in liver glycogen reserves in the poisoned fish was only observed at the end
of the experiment, on the 25th day.
The significant decrease in the concentration of glycogen in the liver in
the initial period of both variants of the experiment was apparently due to
processes for detoxication of the dissolved petroleum products, since it has
been shown that microsomal enzymes in the liver of fish actively metabolize
various petroleum products, included in the detoxication process as a basic
component of glucose. However, the metabolic reaction of the fish in the
spring and autumn experiments differed; this was apparently due to the dif-
ferent physiological condition of the animals. In the pre-spawning period,
all metabolic processes were highly active, indicating a high degree of induc-
tion activity of microsomal enzymes which participate in detoxication of
heterologous compounds and their excretion from the organism. This allows for
more rapid metabolization of the petroleum products and for the active elim-
ination of the metabolites. The results of this toxic effect of the dis-
solved petroleum products decreased, and a state of equilibrium was observed
in glycogen metabolism at some time (data on the 15th and 20th days). We can
thus conclude that the liver of smaris is more resistant to the effect of
petroleum products in the pre-spawning period than in the post-spawning
period. As the concentration of dissolved petroleum products used in the
experiment decreased, the effect of compensatory mechanisms appeared: The
decrease in the concentration of glycogen in the liver was followed by
restoration of glycogen reserves and then another decrease. As a result, the
evolution of the toxic process was triphasic in nature. The changes in the
concentration of glycogen in the tissues of smaris could be considered
fluctuations in the normal level only with the 0.01 mg/liter concentration.
Liver sections from fish exposed to this concentration for 30 days precipi-
tated glucose in amounts which did not differ from those of the controls.
Determination of several enzymes involved in carbohydrate metabolism in
the tissues of the fish showed that the changes observed in the concentration
of glycogen were first related to a decrease in the glycogen-synthesizing
enzyme in the liver and second, to an increase in the activity of glycogen-
splitting enzymes such as glycogenphosphorylase and -amylase. The change in
the activity of glycogenolytic enzymes did not always occur concomitantly.
The decrease in the glycogen reserves in the livers of the poisoned fish was
also caused by an increase in glucose-6-phosphatase activity. The activity of
this enzyme in the brain changed in the same way as in the liver. However,
the high glucose-6-phosphatase activity in the brain corresponded with a high
concentration of glycogen in the brains of the sole in comparison with the
smaris. Glucose split from glucose-6-phosphate apparently plays a substantial
role in cerebral metabolism in sole in a toxic environment; cerebral
metabolism thus becomes more independent of the level of glycemia, which is an
additional factor determining the high resistance of this species to toxic
substances.
The studies which we conducted thus indicate that fish can be widely used
as objects for monitoring pollution of the marine environment based on a
number of their features.
80
-------
REFERENCES
Balayev, L. A., N. D. Mazmanidi, and R. R. Bazhashvili. 1976. Reactions of
some Black Sea hydrobionts in an electrical current field in poisoning by
dissolved petroleum products. Vopr. ikhtiologii 16(4).
Fedorov, V. D. 1976. Problem of admissible effects of the anthropogenic
factor from an ecologist's point of view. Proceedings of the III VGBO
Conference, 1.
Korolev, A. M. , Yu. Yu. Frolov, and N. D. Mazmanidi. 1978. Study of the
physical-chemical properties of mucus from Black Sea smaris in conditions
of petroleum toxicosis. Biol. nauki, 9.
Mazmanidi, N. D. 1974. Symptoms of petroleum poisoning in hydrobionts.
Rybnoye khoz-vo, 9.
Mazmanidi, N. D. 1977. Effect of petroleum poisoning on oxygen saturation of
the blood of the Black Sea fish. Rybnoye khoz-vo, 6.
Mazmanidi, N. D. et aj. 1972. Study of the effect of petroleum pollution of
seas on hydrobionts. Scientific report on subject No. 21, VNIRO, Moscow.
Mazmanidi, N. D. , and L. A. Balayev. 1974. Effect of petroleum poisoning on
the behavior of some Black Sea fish in an electrical current field.
Rybnoye knoz-vo, 11.
Mazmanidi, N. D., and G. I. Kovaleva. 1972. Experimental data on the effect
of petroleum on some chemical properties of the marine environment.
Okeanologiya, 5.
Mazmanidi, N. D., and G. I. Kovaleva. 1975. Effect of dissolved petroleum
products on some elements in carbohydrate metabolism in fish and inverte-
brates. Vopr. ikhtiologii, 15(5).
Mazmanidi, N. D., G. I. Kovaleva, and N. A. Zobova. 1975. Determination of
petroleum products and naphthenic acids dissolved in sea water.
Okeanologiya, 15(3).
Mazmanidi, N. D. , and N. P. Zambakhidze. 1974. On pathological-anatomical
changes in fish in acute and subacute poisoning by petroleum. Trudy
Gruzinskogo otd. VNIRO, 16.
Rice, S. D. , J. V. Short, and J. F. Karinen. 1976. Effects of hydrocarbons
on biological systems: behavioral, physiological and morphological;
toxicity of Cook Inlet crude oil to several Alaskan marine fishes and
invertebrates. Symposium on sources, effects and sinks on hydrocarbons
in the aquatic environment, Washington D.C., U.S.A.
Stroganov, N. S. 1976. Comparative sensitivity of hydrobionts to toxic
substances. General ecology. Biocoenology. Hydrobiology, 3 (Scientific
and technical results. VNITI AN SSSR), Moscow.
81
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IMPROVED ASSESSMENT OF ECOLOGICAL EFFECTS BY INCORPORATING
PHYSICAL SIMULATIONS IN BIOASSAY PROCEDURES
by
D. J. Baumgartner
Marine Division, Environmental Research Laboratory
U.S. Environmental Protection Agency
Corvallis, Oregon
INTRODUCTION
Bioassay procedures are customarily used in the United States by marine
pollution control regulatory authorities to determine the acceptability of
municipal and industrial wastes for disposal in the ocean. In some cases the
results of bioassays are of advisory value while in other cases, such as
disposal of material dredged from waterways, failure to pass one or more
bioassay tests can result in denial of a permit. Since a permit is required
for ocean dumping from barges and for continuous operation of an ocean out-
fall, denial of a permit can result in significant economic and technical
difficulties for municipalities, industries, and dredging authorities. Conse-
quently, dischargers as they may be conveniently labeled, are urging regula-
tory authorities to adopt bioassays that are realistic indicators of environ-
mental stress, but which are not arbitrarily restrictive.
Pollution control authorities are, of course, also interested in reliable
indicator tests, and since they are responsible for protecting the environment
they are motivated to prescribe tests that will be effective in screening out
chemicals and wastes that are likely to be troublesome. Recognizing that
laboratory bioassays are neither infallible, nor exact representations of the
environment, authorities are quite conservative in adopting new techniques
that are projected as being more realistic. It seems to this observer that
pollution control authorities, at least in the United States, favor standards
and uniform tests justified no doubt on legal grounds as well as administra-
tive ease, whereas many scientists generally tend to favor tests that are
tuned to the conditions of the specific case, rather than being standardized.
In recent years major advances in bioassay testing have been incorporated
in the administration of waste discharge permits, because of changes in the
legal basis for pollution control and the tremendous economic impact of permit
decisions. There are however improvements which are still necessary and the
purpose of this paper will be to describe a few that are motivated by con-
sideration of the physical and chemical factors influencing the distribution
of wastes discharged from barges and pipelines.
82
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PRESENT BIOASSAYS
Both the Environmental Protection Agency and the Corps of Engineers have
published bioassay manuals for evaluation of ocean dumping (USEPA 1978; U.S.
Army Engineers 1977). Generally mortality is the endpoint of the bioassay
even though it is recognized that ecological significance is difficult to
interpret directly from this information. The only nonlethal tests being
utilized are the phytoplankton growth tests and one attempting to evaluate
bioaccumulation of toxic chemicals, again recognizing that interpretation is
difficult. Federal ocean dumping permit regulations require use of bioassays
for the liquid phase, suspended phase, and sediment phase.
Most tests are conducted in aquaria without flowing seawater (sometimes
called "static" bioassays) although the recommended procedure for the solid
phase bioassays employs flow-through aquaria. In this test, the benthic
animal survival is determined after 10 days, under a 15-mm blanket of material
to be tested, but in other tests the LD50 is calculated for exposures of 4 to
96 hours, using at least three concentrations of test solution. Results are
compared to the expected concentrations to be found in the dumping zone using
field data or model results. For ocean outfall permits a standard test of
bioassay requirements does not yet exist.
INCORPORATION OF TIME VARYING FACTORS IN BIOASSAYS
Ocean Outfalls
Many models in use today have been developed from the same simple plume
mechanics that have been found to adequately describe the behavior of smoke
plumes. In Figure 1 the main features of an ocean outfall plume are sketched.
The features which are important in relationship to the regulatory use of
bioassays are: (a) the rising plume; (b) the transition zone; and (c) the
drift flow. These three zones are shown again in Figure 2 in terms of the
relative concentrations and length of times to which organisms are exposed.
The initial dilution rates are based on plume calculations for typical ocean
outfall configurations (Baumgartner 1971; Teeter and Baumgartner 1979). The
example used in Figure 1 snowing an initial dilution of 100 to 1 in one minute
is in the range of dilutions frequently found, i.e., between 50/1 and 300/1
within several minutes at most.
In order to operate a bioassay chamber in which the appropriate dilution
rate can be achieved in such a way as to expose the assay organisms it would
be necessary to compute the expected dilution according to accepted models and
adjust the plumbing controls accordingly. The reliability of this initial
dilution estimate is in the range of ± 20% (Brooks 1973) and therefore is
considerably better than many other environmental predictions. The range of
values usually prescribed for bioassays should conveniently bracket this
estimate. The bioassay organisms used in this test should be the planktonic
and fish species selected for the region. (Benthic organisms would not be
expected to be exposed in nature to this concentration/exposure regime). The
96 hour (5,760 minutes) steady state bioassay, if applicable at all, is seen
83
-------
CURRENT
PYCNOCLINE
REGION
TRANSITION
ZONE
DRIFT '
FIELD •
PARTICULATES
RISING
PLUME
DIFFUSER
PORTS
ENTRAPMENT OF
DILUTION WATER
SEABED
Figure 1. Waste field generated by simple ocean outfall
84
-------
INITIAL DILUTION (PLUME)
IN DEEP WATER
HIGH TURBULENCE
TRANSITION ZONE
LOW TURBULENCE
\ RANGE OF
f DRIFT FLOWS
:J
(TYPICAL VALUES USED)
lilt
I 10 100 1000
TIME, MINUTES
i
1000 10,000 100,000
DISTANCE, FEET
Figure 2. Relative efficiency of dilution in flow regimes associated with
municipal outfalls.
85
-------
to be applicable only to the drift flow region, and only for concentrations
that result after initial dilution has been achieved in the rising plume.
Even in this case the test would be more appropriate if concentrations were
decreased in time, at a rate found in field observations or as predicted by
calculations of expected dilutions and pollutant decay. It may be possible to
incorporate these changes in concentration in the drift flow region into
bioassays so that they would more reasonably represent the environmental
conditions. it may however be much more difficult to incorporate the rapid
initial dilution phase into the bioassay test, but this may be the most signi-
ficant exposure experience for organisms caught in the plume. What is most
obvious from this figure is that a 96-hour LD50 based on concentrations higher
than that achieved at the end of the initial dilution is not realistic from
the point of view of environmenal exposures.
Barge Dumping
Figures 3 and 4 show schematically two variations in barge dumping models
(Clark et aj. 1971). The convective descent phase is modeled the same way as
the buoyant plume of ocean outfalls. The long term dispersion is modeled the
same as the drift flow region of ocean outfalls. The barge dumping models are
more complicated because of the necessity to account for a wide variety of
discharge methods and materials which are dumped, particulates being espe-
cially variable. In the case of a single barge dumping event (Figure 3) a
bolus of dense waste is rapidly carried downward until reaching a level of
buoyant equilibrium. As an example of the use of graphical solutions prepared
by Koh and Fan (1968), the top of Figure 5 shows that for a densimetric Froude
number (F1) of 10 for the initial barge dumping conditions, a penetration
depth (Y.) of 10 times the barge diameter (b ) would result for a normalized
stratification value (E) of approximately 10-2. In this case, then, if the
barge were about. 8 meters wide, the penetration depth would be about 80 meters
in a few minutes. If the depth of water at the dumping site were shallower
than 80 meters the waste cloud would immediately begin to deposit particulates
in a relatively compact dome on the seabed. The approximate size of the mound
is computed from the bottom half of Figure 5 as four times the width.
In the above example, if the depth is considerably greater than 80 meters
at the dump site the waste cloud would begin to spread horizontally and be
carried by ambient currents. Since the dilution is approximately proportional
to the size increase cubed, the concentration of pollutants at this level
would be about 1/64 the initial concentration in the barge. The rate of
additional dilution in the long term drift flow, as with ocean outfalls, is
much slower, and suspended phase bioassay conditions may be quite easily
designed to match these expected physical conditions.
Benthic bioassays may be more easily designed when physical factors
suggest bottom encounter because the particulate and water phase concentra-
tions are much more precise than in deep water. In water deeper than the
penetration depth the sedimentation of particulates is much more difficult to
predict because of uncertainties in long term transport calculations.
Another barge dumping configuration frequently employed is discharge into
the wake, as in Figure 4. The theoretical analysis of dispersion is not
*^
86
-------
00
LONG TERM (MONTHS) COLLAPSE
(MINUTES)
BOTTOM TRANSPORT
CONVECTIVE DESCENT
(MINUTES)
Figure 3. Basic transport phases (after Clark et al_. 1971).
-------
ZONE
CO
00
MIXING
DEPTH
(h)
MIXED SURFACE LAYER
Figure 4. Schematic presentation of fate of material when discharged in
barge wake (after Clark et aj. 1971).
-------
Convective descent terminal depth (after Koh and Fan 1968),
bf
Figure 5. Convective descent terminal size (after Koh and Fan 1968).
89
-------
unlike jet (or ocean outfall) analysis (Birkhoff et al_. 1957) and full scale
demonstrations have confirmed this (Abraham et aj. 1972). Because of barge
and propeller variations as well as the varying arrangements of discharge
pipes, especially orifice size and direction, the rates of dilution can easily
vary by a factor of 30 (Redfield and Wai ford 1957; Ketchum and Ford 1952).
Under some variations of wake discharge that appear physically possible,
long term dispersion may be described by methods identical to slug releases
from barges and continuous discharges from ocean outfalls that form a lens at
an intermediate depth.
Long Term Dispersion
Figure 6 graphically displays results of an example (Clark et al. 1971)
of long term dispersion according to the method of Koh and Fan (1968). Four
situations were considered. In one (U.) the waste cloud formed a buoyant
spherical cloud at 50 meters, and was transported for 8.5 hours before the
pollutant concentration at the surface reached a maximum. The surface con-
centration at that point was between 10-4 and 10-5 of the initial value. In a
second case (C, ) the materials in the waste cloud were described in such a way
that the buoyant spherical cloud collapsed into a thin saucer at the initial
stage of transport. In this case it was 35 hours before the surface maximum
concentration was reached. The concentration would be expected to be about
10-6 of the initial concentration. A third case (Up) describes a spherical
mass influenced by a strong pycnocline above the penetration depth. The
pycnocline retards vertical migration as seen by the reduced surface concen-
tration of 10-7 times the initial value. The fourth case is not described
graphically because the surface concentration remained essentially zero. In
this example the sea again was considered as possessing a strong pycnocline
and in addition the waste material in the cloud was of such nature that the
buoyant spherical cloud collapsed into a saucer shape.
These curves, along with other curves that could be computed for the
centroid of the waste flow and the seabed surface can be used to describe
exposure time and time varying concentrations to be employed in bioassays.
Depending on the type of material discharged and the physical arrangements of
the barge or outfall, these calculations also provide insight to the relative
importance of surface, water column, and seabed in selecting bioassay organ-
isms as well as in evaluating the relative ecological significance of bioassay
results.
The above example also shows the possibility of vastly increased vari-
ability in results as more complex, but more realistic, environmental factors
are considered. Recall that the plume calculations were ± 20%.
Physical Variability
Greater variability occurs with dumping of dredge spoils and sludges than
with ocean outfall discharges because the material may range from fine suspen-
sions to large cohesive clumps of clay. The mathematical models used to
describe the physical behavior must be quite elaborate to account for the wide
range of conditions. This can be accomplished by incorporation of adjustable
90
-------
C= COLLAPSED CLOUD
U= UNCOLLAPSED CLOUD
L = LINEAR DENSITY GRADIENT
P = PYCNOCLINE
actual
FIRST APPEARANCE OF
SURFACE CONCENTRATION
I
HOURS
Figure 6.
10
100
10
r4
10
-52
10"
10
I
LU
o
o
LU
8
-7 cr
CO
Q
LU
ia
-8
10'
0
LU
cr
CL
LU
_
LU
cr
Predicted relative surface concentration for lonq-term dispersion
(after Clark et al_. 1971).
phase
-------
coefficients. For example, for the illustrations in Figures 5 and 6 a
modestly large entrainment coefficient was used as would be appropriate for a
liquid or a slurry of nonagglomerated particles. To employ the same entrain-
ment value for a barge of cohesive clays of low water content would be mis-
leading. The unwary or unwise may be led to believe the model is defective
rather than a case of misapplication.
Laboratory hydraulic model studies are reasonably efficient for deter-
mining values of coefficients to describe behavior of dumped material in the
convective descent phase. As an example, hydraulic model studies based on
Froude and Reynolds numbers of dredge spoil from the Duwamish River estuary,
as conducted at our laboratory suggested an entrainment coefficient of essen-
tially zero for barge dumping in 67 meters of water in Elliott Bay. Using a
mathematical model developed by Koh and Chang (1973) we predicted the material
would impact the seabed without measurable dilution. Field observations
during dumping confirmed the hydraulic and mathematical model predictions
(Baumgartner et al. 1978).
Hydraulic models may not be effective for obtaining coefficients relating
to barge performance and oceanic turbulence, both of which influence disper-
sion, chemical interactions, and biological exposure, as do the properties of
waste material itself. Consequently it is important to acquire field data,
which even in limited circumstances may be sufficient to characterize the
coefficients needed to model exposure conditions.
Bioassays may need to be conducted without adequate field information.
The range of exposure conditions to be employed in the bioassays, or at least
considered before final bioassay design, can be provided by repeated mathe-
matical simulations. Done in an orderly way this exercise may require only
one day and can highlight the natural situations which are most influential in
controlling the exposure condition. A sensitivity analysis of this sort was
conducted on the Koh-Chang model (Teeter and Johnson 1979). Figures 7 and 8
show the range of exposure conditions appropriate for benthic bioassays depend-
ing on the values that are determined to be appropriate for two coefficients,
a and CHRAG' *n eacn fiQure tne top section shows the percent of particulate
material reaching the seabed as a function of time, for coefficient values
adjacent to each curve. The lower portion of each figure is a measure (VoO of
the distribution of material around the mass centroid assuming a Gaussian
form. These two sets of graphs may be used to determine the mass of material
to be placed on a simulated seabed for benthic bioassays. This exercise may
well elicit biological insights to be fed back to the physical scientists
prior to field studies so that effort might be maximized on those coefficients
which are most important biologically.
Chemical Variability
Chemical models can be used to describe the behavior of some chemical
species under varying conditions of the environment including salinity, pH,
oxygen content (Eh), and the concentration of cations and anions (Ingle et al_.
1978). Figure 9 is an example of the effect of salinity on the distribution
of silver species as described by the EPA model using data from Jenne et a_[.
(in press). This shows that it may be very important to incorporate^ time
92
-------
100 r
0.300
400
800
1200
1600
o Hiooo
O ^. 500
LJ Q.
Og
100
50
< O
Fm
10
0
0.059
400
1
800 1200
B
1000
500
9*
J> 100
10
0
.400
0.470
0.300
400 800 1200
TIME, SEC.
Figure 7. Sensitivity of model output to a in bottom dump
disposal (after Teeter and JohnsSn 1979).
93
-------
Q
8g
100
75
50
ui
a: o 25
ui
a.
0
0 H 300
a u!
Sfcf
t
en
UI fL
Og
g ^200
ujg
o r~
IOO
• 0.50
6.50
A
•2.0
400 800 1200 1600
B
2.0
500
1000
1500
200
H
u.
h-
100
500 1000
TIME, SEC.
1500
Figure 8. Sensitivity of model output to CDRAG
(after Teeter and Johnson 1979).
in bottom dump disposal
94
-------
AgCi;3
25 50 75
PERCENT SEAWATER
100
Figure 9. Concentration of silver solute species in freshwater-
saltwater mixtures. Total silver concentration:
10-6 molar (Ingle and Baumgartner, unpublished).
95
-------
varying concentrations into bioassays intended to assess the effects of wastes
containing silver to the marine environment. As most discharge practices can
achieve a mixture 90% seawater within several minutes, the kinetics of such
reactions may be important. Assuming that reactions, such as the silver
speciation reactions, are as rapid as the physical dilution reactions, or even
an order-of-magnitude slower, it is obvious that current bioassays based on
96-hour "static" exposures do not adequately represent exposure conditions in
the initial dilution. Figure 9 also demonstrates that exposure conditions in
the upper reaches of estuaries (salinity < 25%) are quite different than in
the open ocean.
CONCLUSIONS
Recent advances in modeling and in acquisition of data to describe chemi-
cal and physical behavior of pollutants have provided the basis for improved
bioassay procedures. The benthic bioassays used for the solid phase evalua-
tion of dredge spoils have incorporated some of the advances. Laboratory
systems have become more sophisticated and elaborate in response to the need
for more effective bioassay procedures and it may now be possible to incor-
porate time varying concentrations in bioassays through controlled continuous
dilution techniques.
REFERENCES
Abraham, G. , W. D. Eysink, G. C. van Dam, J. S. Sydow, and K. Miller. 1972.
Full scale experiments on disposal of waste fluids into propeller stream
of ships. pp. 471-474 _In Marine Pollution and Sea Life. Mario Ruivo
(ed.) Proc. FAO Tech. Conf. on Marine Pollution and its Effects on
Living Resources and Fishing, Rome, Italy, December 19, 1980. Fishing
News (Books) Ltd: Surrey and London.
Baumgartner, D. J. 1971. Disposal of liquid and particulate wastes to the
ocean. Chemical Engineering Symposium Series. Water 67(107):46-53.
Baumgartner, D. J. , D. W. Schults, and J. B. Carkin. 1978. Chemical and
physical analyses of water and sediment in relation to disposal of
dredged material in Elliott Bay. Army Corps of Engineers Dredged
Material Research Program Technical Report D-77-24. 215 pp.
Birkhoff, G. , and E. H. Zarantonello. 1957. Jets, wakes and cavities. New
York Academic Press. 353 p. (Applied Mathematics and Mechanics, vol.
2).
Brooks, N. H. 1973. Dispersion in hydrologic and coastal environments. U.S.
Environmental Protection Agency. EPA-600/3-73-010. 136 pp.
Clark, B. W., W. F. Rittall, D. J. Baumgartner, and K. V. Byram. 1971. The
barged ocean disposal of wastes. A review of current practice and
methods of evaluation. U.S. Environmental Protection Agency, Corvallis,
Oregon. 120 pp.
96
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Ingle, S. E. , M. D. Schuldt, and D. W. Schults. 1978. A user's guide for
REDEQL.EPA: A computer program for chemical equilibria in aqueous
systems. U.S. Environmental Protection Agency, EPA-600/3-78-024,
Corvallis, Oregon.
Jenne, E. A., D. C. Birvin, J. W. Ball, and J. M. Burchard. In press. Inor-
ganic speciation of silver in natural waters—fresh to marine, _In
Environmental Impacts of Nucleating Agents Used in Weather Modification
Programs, D. A. Klein, (ed.), Dowder, Hutchinson and Oss, Straudsberg,
Pennsylvania.
Ketchum, B. H., and W. L. Ford. 1952. Rate of dispersion in the wake of a
barge at sea. Trans. AGU. 33(5):680-684.
Koh, R. C. Y. , and Y. C. Chang. 1973. Mathematical model for barged ocean
disposal of wastes. U.S. Environmental Protection Agency, EPA-600/
2-73-029. Corvallis, Oregon.
Koh, R. C. Y. , and L. Fan. 1968. Prediction of the radioactive debris cloud
distribution subsequent to a deep underwater nuclear explosion (U). A
final report prepared for the Naval Radiological Defense Laboratory,
Contract No. N00228-68-C-0684. 266 p.
Redfield, A. C., and L. A. Walford. 1951. A study of the disposal of
chemical waste at sea. NAS-NRC Pub. #201.
Teeter, A. M., and D. J. Baumgartner. 1979. Prediction of initial mixing for
municipal ocean discharges. U.S. Environmental Protection Agency,
Environmental Research Laboratory Report, CERL-043, Corvallis, Oregon.
88 p.
Teeter, A. M. , and B. H. Johnson. 1979. A computer study of the Koh-Chang
model for dredged material disposal. U.S. Environmenal Protection
Agency, EPA-600/3-79-027, Corvallis, Oregon.
U.S. Army Engineer Waterway Experiment Station. 1977. Ecological evaluation
of proposed discharge of dredged material into ocean waters. Implementa-
tion manual for Section 103 of Public Law 92-532. Vicksburg,
Mississippi.
U.S. Environmental Protection Agency. 1978. Bioassay procedures for the
ocean disposal permit program. EPA-600/9-78-010, Gulf Breeze, Florida.
97
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A PROGRAM FOR INVESTIGATING POLLUTION OF MARINE
MACROBENTHOS BY HEAVY METALS AND ITS POSITION
IN THE SYSTEM FOR MONITORING THE ENVIRONMENT
by
K. S. Burdin and I. B. Savel'yev
Department of Biology
Moscow State University
Until recently, the collection of information on the state of the envi-
ronment was typically random. Awareness of the unsatisfactory condition of
the environment has occurred after the fact. The global nature of the anthro-
pogenic influence on Earth's biosphere in recent years has led to gradual
deterioration of the quality of the human environment. The recently created
procedure for assessing the state of the environment on local, regional,
national, and global levels has been termed "environmental monitoring."
Theoretical and applied aspects of environmental monitoring have been
described and discussed in the literature (Gerasimov 1977; Gasilina and
Rovinskiy 1977; Simonov 1977). Definitive plans for data flow based on
previously developed programs to provide maximum satisfaction of three
criteria—observation, evaluation, and prediction of the state of the environ-
ment for the purpose of making decisions concerning regulation of the quality
of the environment (IzraeV 1977)--are of special interest.
In addition to discussing the generalized schemes for environmental
monitoring, biological monitoring and monitoring pollution of the marine
biota, concrete programs for diagnostic and prognostic monitoring of heavy
metal pollution in marine macrobenthos are discussed, and an attempt to use
the proposed programs for monitoring pollution of macrobenthos is presented.
ENVIRONMENTAL MONITORING
Structure of Environmental Monitoring
In this paper, "environment" will refer to the natural and artificial
physical and chemical factors and the social changes which could have a direct
or indirect effect on the condition of the abiotic and biotic component of the
biosphere and on man. With respect to this determination, environmental
monitoring should be constructed on the basis of three independent elements:
monitoring the natural environment, monitoring the artificial or transformed
environment, and monitoring the anthroposphere (Figure 1). "Natural environ-
ment" means the natural factors which have not yet been used by man as sources
of raw materials or energy, work tools, and consumer goods, but are very
98
-------
Geophysical
monitoring
Monitoring
the natural
environment
Geochemical
monitoring
Monitoring the
transformed and
artificial environment
Montoring the
anthroposphere
Biologic
monitoring
• — ••• • • '
Physical-
geographic
monitoring
Monitoring
anthropogenic
changes
Demographic
monitoring
Sanitary-
hygienic
monitoring
Monitoring
pollution
in biota
Monitoring
productivity
of the
biosphere
Monitoring
disappearing
species and
species at the
brink of
extinction
Monitoring the
most important
species, popula-
tions, societies,
and ecosystems
(genofund)
Figure 1. Diagram of the environmental monitoring process.
-------
significant for the level of human life and the economic potential of the
society. "Artificial environment," most often called the "technosphere" in-
the literature, includes objects and edifices created by man while working to
provide production materials, consumer goods, and the nonindustrial needs of
the society. "Transformed environment" chiefly includes the agrosphere used
by man to produce food and raw materials from cultivated plants and animals.
Figure 1 shows the structure of environmental monitoring which includes
monitoring of the anthroposphere, i.e., the population of the Earth, which
basically coincides with the technosphere and the agrosphere.
Monitoring the natural environment includes monitoring of the abiotic and
biotic components. The basic indices of the condition of the abiotic compo-
nent of the natural environment, including the lithosphere, the hydrosphere,
the pedosphere, and the atmosphere, can be measured within the framework of
geophysical, geochemical, and physical-geographic monitoring (IzraeV 1977).
Monitoring the anthropogenic changes is the basis for observations, evalua-
tions, and prognosis of the condition of the transformed and artificial
environment. Demographic and sanitary-hygienic monitoring can now be used to
monitor the anthroposphere.
The condition of the biotic component of the natural environment is
observed by biological monitoring, including not less than four types of
monitoring at present: monitoring pollution of the biota, monitoring produc-
tivity of the biosphere, monitoring disappearing species of animals and plants
and species at the brink of extinction, and monitoring the most important
species, populations, societies, and ecosystems.
Structure of Monitoring Pollution of the Biota
We know that substances of an anthropogenic origin entering the environ-
ment are included in the cycle of matter and consequently appear in the biotic
component in some interval of time. Migration and the time in the abiotic
component, as indicated above, are observed within the framework of geochem-
ical and geophysical monitoring, while pollution of biota is monitored by
observing the level of pollution of the biotic component, movement along the
food chain, and the time of accumulation and elimination of pollutants.
Monitoring pollution of biota as a function of the living environment is
subdivided into monitoring pollution of biota in fresh water, monitoring
pollution of biota in salt water, and monitoring pollution of biota on land
(Figure 2). Basal monitoring of pollution is now done within the framework of
monitoring pollution of biospheric national forests.
We shall examine the structure of monitoring pollution of marine biota
which includes monitoring pollution of benthic organisms and monitoring pollu-
tion of pelagic organisms. Pollution of benthic organisms is monitored by
observing the level of pollution in macrobenthic organisms and meso- and
microbenthic organisms (Figure 2). Monitoring pollution in pelagic organisms
is based on the trophic relationships of the organisms and can include four
types of monitoring—in phytoplankton, zooplankton, fish, and mammals.
100
-------
Monitoring pollution
in biospheric national
forests
Monitoring pollution
in freshwater biota
Monitoring pollution
in marine biota
Monitoring pollution
land biota
in
Monitoring pollution
in benthic organisms
Monitoring pollution
in pelagic organisms
Monitoring
pollution
in macro-
benthos
Monitoring
pollution
in micro-
benthos
Monitoring
pollution
in phy to-
pi ankton
Monitoring
pollution
in zoo-
plankton
Monitoring
pollution
in fish
Monitoring
jollution
in mammals
Figure 2. Diagram of the pollution morrftoring process in the biota.
-------
Structure of Monitoring Pollution in Macrobenthos
We shall examine one of the possible variants of the structure of moni-
toring pollution in the macrobenthos as part of the overall structure of moni-
toring environmental pollution (Figure 3). Using the spatial and trophic
indices to form the structure for monitoring pollution in marine biota, we can
distinguish the littoral, neritic, bathyal, and abyssal zones of the ocean.
The biogeochemical role of benthic organisms which live in the neritic,
bathyal, and abyssal zones has not been studied in depth. Most data on
pollution in the macrobenthos relate to organisms which live in the coastal
zone. Benthic organisms which live in this zone play a major role in purify-
ing sea water from pollutants which enter from the shore and by way of
currents from the open sea. With respect to the existing notions of moni-
toring and the set of activities related to observation, evaluation, and
prognosis, monitoring of pollution in macrobenthos in the coastal zone will
require two types of activity—diagnostic and prognostic monitoring of macro-
benthos pollution.
The program for monitoring heavy metal pollution of the macrobenthos
should include three stages. The first two involve reconnaissance to deter-
mine the levels of pollution and prognostic experiments. The third stage
involves creating a network of diagnostic monitoring stations whose scale is a
function of the problem posed and can vary from the local to the global level.
Diagnostic Monitoring Program
The framework of our approach to diagnostic monitoring is a set of
systematic observations of the level of pollution in macrophytobenthos and
macrozoobenthos. Measurement of any biological response to any factor of an
anthropogenic origin can be included in the diagnostic monitoring program.
The capacity of many organisms to store pollutants is one of the most easily
detected responses of biota to the presence of pollutants. For this reason,
periodic measurement of the quantity of pollution in the Jiving components of
the environment is a basic task in current diagnostic monitoring.
With respect to the program for diagnostic monitoring of heavy metal
pollution in the coastal zone, the criteria for selecting organisms for
monitoring are: extensive living areas, mass, and possible use of the
organism as food or as an industrial resource. Biological samples at differ-
ent geographic points in the world's oceans should be collected at the same
time of the year. Potential errors in measuring the concentrations of metals
can be minimized by unified collection, storage, and analysis of the samples
for the concentration of heavy metals (Martin 1979). A comparison of the
available data on the level of pollution of macrobenthos organisms collected
at different geographic points will make it possible to detect origins with
normal and high levels of pollution in the marine environment.
Prognostic Monitoring Program
Prognostic monitoring can be conducted if data are available on the
concentration of heavy metals in sea water and hydrobionts, i.e., a basis for
active experiments can be constructed. Any type of experiment on the effect
102
-------
o
u>
Monitoring pollution
in macrobenthos in
the coastal zone
Monitoring pollution
in macrobenthos in
the neritic zone
Diagnostic monitoring
Observation of
the level of
pollution in
macrophytobenthos
Observation of
the level of
pollution in
macrozoobenthos
Monitoring pollution
in macrobenthos in
the bathyal zone
Monitoring pollution
in macrobenthos in
the abyssal zone
Prognostic monitoring
Toxicologic
evaluation of
pollution on
test organisms
Toxicologic
evaluation of
pollution on
small microcosms
Toxicologic
evaluation of
pollution on
field microcosms
Figure 3. Diagram of the pollution monitoring process in the macrobenthos.
-------
of pollutants on individual species or on complex ecological systems conducted
in laboratory or field conditions will be called prognostic monitoring. As
shown in Figure 3, prognostic monitoring includes conducting laboratory and
field experiments on test organisms in large and small microcosms for toxico-
logic evaluation of pollutants, for determining the limits and rates of accum-
ulation and elimination, and for studying degradation of pollutants in the
presence of test organisms used and migration of pollutants along the food
chain of the organisms.
Any prognostic monitoring program, regardless of the scale of observa-
tion, includes at least the following types of activity: maintenance or
cultivation of the test organisms in laboratory conditions, creation of
laboratory models of natural ecosystems of different degrees of complexity
(static, flowing systems, and systems with a controlled degree of pollution),
treating mathematical models of pollution which have predictive value.
A POTENTIAL PROGRAM FOR MONITORING HEAVY
METAL POLLUTION IN MACROBENTHOS
We conducted an experiment on the Crimean and Caucasian shores of the
Black Sea to implement the program developed to monitor pollution of macro-
benthos by heavy metals. The preliminary research primarily consisted of
reviewing the literature on the concentrations of copper, zinc, and lead in
sea water and hydrobionts.
Diagnostic Monitoring
Selecting the Biomonitor. Representatives of phytobenthos and zoobenthos
which have been adequately studied and meet established requirements can be
used as the monitoring organisms. The selected organisms should be widely
distributed, cosmopolitan species which can accumulate the pollutants under
consideration, yet preserve the basic indices for vital activity and genetic
stability when relatively high concentrations of the pollutants are present in
the environment. In addition, it is desirable for the monitoring organisms to
be readily available for collection and to have a long life span. It has been
shown in previously published studies that bivalve mollusks and macrophytes
are promising organisms for monitoring heavy metal pollution i-n the marine
environment (Berner et al. 1972; Goldberg 1975).
Two species of: organisms which possess the features listed—the mussel
Myti1 us galloprovincialis and the brown alga Cystoseira barbata--were selected
to monitor heavy metal pollution of the Black Sea macrobenthos. Colonies of
Myti1 us and Cystoseira were almost universal in the sublittoral zone of the
Crimean and Caucasian coasts in natural and artificial solid substrates.
Collecting the Biological Samples. The biological samples were collected
in the winter of 1977 and 1978 in the sublittoral zone of the Crimean and
Caucasian coasts. Sampl
0.5-3.0 m; the organisms
mine their age. Each d
as of mussels and algae were collected at a depth of
were subsequently analyzed morphometrically to deter-
mensional group contained not less than 10 specimens
104
-------
of mussel and 5 specimens of Cystoseira. The bodies of mussels and algae of
the same size were dried, preserved in desiccators, and reduced to a constant
weight immediately before analysis to determine the concentrations of metals.
Method of Calcining the Biosamples and Analyzing the Heavy Metal Concen-
tration. The samples were prepared for analyses using the method of dry
calcination at a maximum temperature of 450°C and subsequently dissolved in
hydrochloric acid. The concentration of Cu, Zn, and Pb in the solutions was
measured by the atomic absorption method in Perkin-Elmer 403 and AAS-1 (Karl
Zeiss, Jena) devices. The preliminary calibration of the devices showed a
satisfactory convergence of the data. The concentrations of metals in the
biosamples were calculated with respect to the weight of the dry residue.
Concentration Ranges £f Cu, Zn, and Pb i_n the Samples. The data on the
concentrations of Cu, Zn, and Pb in the bodies of mussels from the average
size group having a shell length of 40-45 mm (mean age of approximately 2
years) are shown in Table 1. Stations 1, 2, and 3 were located on the Crimean
coast and Stations 4 and 5 were located on the Caucasian coast. The lowest
concentrations for all metals were found in the samples from Station 3,
located in the region of the Karadagskaya biostation of the Institute of
Southern Seas Biology of the AN USSR. This formed the basis for selecting
this station as the "basal" station and conducting the corresponding experi-
ments in the prognostic monitoring program.
TABLE 1. CONCENTRATIONS OF METALS IN BLACK SEA MUSSELS
Concentrations, pg/g dry weight
Station Zn Cu Pb
1
2
3
4
5
250
179
150
166
231
5.93
5.36
7.60
9.27
31.90
2.76
2.94
0.92
3.17
7.89
The range of concentrations of Cu in the samples of marine algae
collected on the Crimean and Caucasian coasts varied within the limits of 2.0
to 40.0; from 10.0 to 60.0 for Zn, and from 3.0 to 50.0 ug/g dry residue for
Pb.
Changes iji Metal Concentrations _ni Samples as a Function £f Size (Age).
The concentrations of Cu and Zn in M. galloprovinciatis living in one colony
were not a function of the size of the shells in the broad range of size
groups considered (Table 2). The high concentrations of Pb in the first two
size groups and Zn in the last two were apparently due to the corresponding
changes in the concentrations of these metals in the marine environment.
105
-------
TABLE 2. CONCENTRATION OF METALS IN MUSSELS AS A FUNCTION
OF SHELL LENGTH
Length of shell,
mm
Concentration, ug/g dry weight
Cu
Zn
Pb
20-26
30-32
35-38
42-46
50-53
53-56
58-60
67-68
74-78
8.64
6.62
6.39
5.93
5.87
6.11
6.91
5.01
6.34
269
210
226
250
218
214
256
457
387
20.8
19.4
3.5
2.8
3.9
4.1
3.3
4.3
4.6
We selected two size groups of mussels with shell lengths of 40-45 and
60-70 mm for subsequent study in the diagnostic monitoring program.
The changes in the concentrations of Cu, Zn, and Pb in the samples of
macrophytes as a function of the length of the axial stipes of the algae are
shown in Table 3. Algae with axial stipe length of 20-25 cm were selected for
the work done in aquarium conditions. The age of the plants was determined
according to the method of Sabinin and Schapova (1954). Based on this method,
Thallophyta with stipe length of 20-25 cm were approximately one year of age.
TABLE 3. CONCENTRATIONS OF METALS IN SAMPLES OF
Cystoseira AS A FUNCTION OF LENGTH OF THE AXIAL
STIPES
Length of axial
stipes, cm
Concentration, ug/g dry weight
Cu
Zn
Pb
20
100
10.3
5.4
39.8
29.4
8.0
4.7
Prognostic Monitoring Program
Experimental studies must be conducted under rigorously controlled condi-
tions to determine the limits of accumulation and elimination of Cu, Zn, and
Pb by monitoring organisms. The experiments should first concern the
following question: With what range of concentrations of metals in the envi-
106
-------
ronment can the organisms be reliable monitors? The answer could serve as the
basis for developing standard prognostic programs for an organized network of
diagnostic monitoring stations.
Mussels and Cystoseira of the same size were selected and acclimatized in
flowing marine aquaria for approximately 10 days. The experiments involving
fixed additions of solutions of Cu, Zn, and Pb salts were conducted in
30-liter plastic tanks; the water was changed twice a day. A fixed number of
specimens of the same size was exposed to different concentrations of the
metals; when the water was changed samples were taken from 5 specimens of
mussel and 3 Cystoseira plants. Seventy-two hours later, the remaining
mussels and plants were placed in special nets, carried out to sea 100 m away
from shore and lowered to a depth of 1 meter below the surface. Samples were
taken twice a day for the next three days to determine the limits of elimina-
tion of the metals. Figure 4 shows a diagram of the changes in the concentra-
tion of metals in the samples of mussels and Cystoseira as a function of the
concentration in the sea water. The columns in the diagram indicate the
maximum values for the concentrations of metals in the mussels and plants. As
the diagram shows, mussels and Cystoseira accumulate the metals in direct
proportion to their concentration in the water. The data from this brief
experiment did not permit establishment of any correlations for this process.
£ 200
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1.0
tun/litm
MUSSELS
CYSTOSEIRA
MUSSELS
CYSTOSEIRA
MUSSELS
CYSTOSEIRA
Figure 4. Diagram of changes in the concentrations of copper, zinc, and lead
in mussels and Cystoseira as a function of the concentration of
metals in the water (first column: controls; second column: 72
hours after accumulation; third column: 72 hours after removal).
107
-------
functon,
fu^tlols
valuesfor6
1977? J
cnnr-rat7 -*°K- thf * exPeriments allowed us to establish the range of
concentrations with! n which mussels and Cystoseira could be reliable monitor-
r8 if°VU' Zn' ^ Pb °n the C0ast of the B1ack Sea- I? is ev dent
il accumulatl°n .f which the organisms preserve their basic
1s tne uPPer ^mit of this range. In the course of the exper-
col?cei?V-ations of c°PP*r, zinc, and lead exceeded the mean
Concentrations of these metals found (based on Patin's data,
°f the Black Sea (Cu 4 MQ/Hter; Zn 30 ug/liter- Pb 3
and the -«1-» concentrations did no?
rnn t - mUSSe1s and al9ae in Se3 water to which a 10
- Con5en.tra,tlon of !ead had been added for three days caused some
0inm°n/iV1ttal frct1ons' but did not result in death, while a concentra-
Knin Ic9i * °f copper resulted in a 50% death rate in the same period.
Keeping mussels and algae in water containing a concentration of 10 mg/liter
of thP /in1" K66. ^ d,ld not have any ne9at1ve effect ^ the vital activity
nln * i 9 ' ^ klll6d isolated specimens of the mollusks and resulted in
funct'ln SofPPtehSS10Hn. °f their Vital functio"s (disorders in the closing
function of the adductors and detachment of the byssus to the substrate)
5 ?i:esfnts the data on the accumulation of copper and the survival
u iU!uS.anu algae as a functl"°n of the concentration of Cu in the
We found that the mollusks began to die with a concentration of copper
w=t
water.
1000
cc
UJ
I
ui
o
o
o
o
10
150
100
CO
UJ
-8
i
50
o
3
1000
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\
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\
4 6
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\
\
i\
468
k-IO'2
I
I
1000
o>
I001
10
8
3
50 100
PERCENT SURVIVAL RATE
Figure 5.
50 100
PERCENT SURVIVAL RATE
Effect of the concentration
galloprovincialis (A) and the brown alga Cystoseira barbata"(B) on
their survival and the concentration factor. *
of copper in the water and in Myti1 us
108
-------
of over 100 jjg/g (Figure 5A), near the LD50 (150 ug/g) for mussels when they
were kept in water containing a 1 mg/liter concentration of copper for three
days. Keeping Cystoseira in sea water containing copper (1 mg/liter) with a
concentration of approximately 300 ug/g in the thai 1 us turned the plant dark,
gave it a sharp odor, and killed it.
The data allowed us to determine four basic ranges of concentrations for
the metals in water and the corresponding ranges of concentrations in the
tissues of the mollusks and algae at which they exhibit certain changes
related to their vital activities. The ranges of concentrations, principally
characterized by the different mechanism of action of the metals on the
organism sampled, are indicated by Roman numerals in Figure 5.
I - The range of a deficit of the metal is characterized by the fact that
its concentration in the water is so small that a shortage of the element can
result in inhibition of growth and even the death of the animal or plant.
This range was hypothetically distinguished, since this situation seldom
occurs in a marine environment in natural conditions. The growth curves for
unicellular marine algae grown in environments poor in microelements were
determined in experimental conditions, but there are no data on conducting
analogous experiments on macrophyte and mollusks in the literature.
II - The range of biogeochemical response corresponds to the basal level
of the concentration of the metal in the organisms and is exclusively a
function of the ecologic conditions. This range is related to fluctuations in
the concentrations of metals in the marine environment determined by the
hydrochemical and geochemical conditions in the living place. The investiga-
tion of the elemental composition of the organisms in this range of concentra-
tions belongs to the field of biogeochemical studies.
Ill - The monitoring range reflects the response of the organisms to
changes in the concentration of metals in the water which are primarily of
anthropogenic origin. In this range of concentrations, the monitoring organ-
isms are exposed to an effect responsible for certain biological effects which
are a function of the concentration of the metal in the marine environment,
the duration of the effect, and the form of its existence. The greatest
difficulties are related to establishing the upper threshold of the range of
concentrations which is actually the lower threshold of the appearance of a
toxic effect. In our work, we conducted brief acute experiments to establish
this range. We shall subsequently specify the upper limit of the monitoring
range by conducting longer experiments.
IV - The range of the toxic effect corresponding to manifest disorders in
the basic physiologic functions of the organisms which lead to their death.
In the proposed classification of concentration ranges in monitoring
organisms, the monitoring range for copper in mollusks is from 20 to 100 ug/g
of dry residue and is from 3 to 400 ug/g for Cystoseira. The zinc monitoring
range for mollusks is within the limits of 150-500 ug/g of dry residue when
the concentration of Zn in the water varies from 30 to 3,000 ug/liter. Since
almost no lead is detected in the tissues of mollusks with a concentration of
109
-------
3 (jg/liter in the marine environment, we can assume that the lead monitoring
range is within the limits of 0-300 ug/g of dry residue when the concentra-
tions in the marine environment varies from 3 to 10,000 ug/liter.
We intend to determine the rate of accumulation and elimination of metals
by selected monitoring organisms, to determine the biologic nature of these
processes based on elementary and biochemical studies of different organs, and
to extend the list of heavy metals which will be investigated in subsequent,
more prolonged studies based on the prognostic monitoring program.
SUMMARY
The paper deals with a general scheme for monitoring the environment,
biological monitoring, and the monitoring of the pollution of marine biota.
The state of coordination of various types of monitoring in the hierarchic
structure of monitoring the environment has been shown, the interaction of
different research trends and the place of our investigation in this structure
which is devoted to developing a program for monitoring the pollution of
macrobenthos by heavy metals.
The results of the experiments in this monitoring program are given in
the second part of the report. A study has been made of mussels (Mytilus
gal 1oprovincialis) and brown algae (Cystoseira barbata) collected in the
littoral zone of the Crimea and Caucasus. The samples have been studied
morphometrically, weighing the parameters and concentration of Cu, Zn, and Pb
by atomic absorption spectrophotometer. At the first stage, reconnaissance
observations were conducted within the framework of diagnostic monitoring
which made it possible to estimate the change in the concentrations of metals
depending on the size (age) of the organism, to determine the ranges of
concentrations of Cu, Zn, and Pb in samples collected at different stations,
as well as to choose a "basic" station for conducting experiments regarding a
program of forecasting. These experiments were made in aquaria of the
Institute of Biology of the Southern Seas. Each experiment took six days
(three days of monitoring organisms treated in sea water with one of the
metals, and after that they were withdrawn and placed in the natural water of
a marine bay for three days). These experiments showed the change in the
concentration of heavy metals in the tissues of mussels and algae at the stage
of accumulation and extermination in relation to the concentration of metals
in sea water. The limits of concentrations were determined at which these
organisms may be used as reliable monitors of Cu, Zn, and Pb in sea water of
the littoral zone of the Black Sea.
110
-------
REFERENCES
Berner, L. , J. H. Martin, J. McGowan, and J. Teal. 1972. Sampling marine
organisms. Jji Marine Pollution Monitoring: Strategies for a National
Program. Deliberations of a workshop held at Santa Catalina Marine
Biology Laboratory of the University of Southern California, Allan
Hancock Foundation, October 25-28, 1972.
Gasilina, N. K., and F. Ya. Rovinskiy. 1977. A national system for moni-
toring the environment in the USSR. Jji Proceedings of the First Soviet-
English Symposium. Environmental Monitoring, Gidrometeoizdat Press,
Leningrad.
Gerasimov, I. P. 1977. Scientific principles of environmental monitoring.
Ibid.
Goldberg, E. D. 1975. The mussel watch—a first step in global marine moni-
toring. Mar. Pollut. Bull. 6(6).
Izrael', Yu. A. 1977. The concept of monitoring the state of the biosphere.
In Proceedings of the First Soviet-English Symposium. Environmental
Monitoring, Gidrometeoizdat Press, Leningrad.
Martin, J. G. 1979. The effect of metals on the marine environment. Li Man
and the Biosphere, 3rd ed., MGU, Moscow.
Patin, S. A. 1977. Ecologic toxicology and biogeochemistry of pollutants in
the world's oceans. Doctoral Dissertation, Moscow.
Sabinin, D. A., and T. F. Schapova. 1954. Growth rate, age, and production
of Cystoseira barbata in the Black Sea. Tr. In-ta okeanologii, 8.
Simonov, A. I. 1977. Oceanographic aspects of the pollution of seas and
oceans. In Proceedings of the First Soviet-English Symposium. Environ-
mental Monitoring, Gidrometeoizdat Press, Leningrad.
Ill
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EFFECTS OF LIQUID INDUSTRIAL WASTES ON ESTUARINE
ALGAE, PLANTS, CRUSTACEANS, AND FISHES
by
Gerald E. Walsh and Richard L. Garnas
Environmental Research Laboratory
U.S. Environmental Protection Agency
Gulf Breeze, Florida 32561
ABSTRACT
A chemical fractionation scheme using ion-exchange resins for separation
of organic and inorganic constituents of complex industrial wastes has been
developed for use with bioassays in order to identify the toxic components of
such wastes.
Grass shrimp (Palaemonetes pugio) and sheepshead minnows (Cyprinodon
variegatus) were not as good indicators of possible effects of complex wastes
on estuarine organisms as the diatom, Skeletonema costatum, in bioassays per-
formed at our laboratory. The diatom was affected by wastes in either of the
following three ways: stimulation, inhibition, or stimulation at low concen-
trations but inhibition at higher concentrations. Toxicity to algae corre-
lated with toxicity to the mysid, Mysidopsis bahia.
Chemical fractionation with appropriate biological testing provides a
means of estimating a more complete potential effect of an effluent on receiv-
ing waters. Effects of growth stimulators, for example, can be masked by
toxicants in whole waste.
We conclude that chemical fractionation is required for comprehensive
analysis of possible effects of complex industrial wastes on estuarine organ-
isms, and that fractionation should be coupled to bioassays that use at least
an alga and a crustacean as test organisms.
INTRODUCTION
The possible impact of increasing industrialization along coastlines on
marine ecosystems is great because industrial plants often emit large volumes
of liquid effluents that contain bioactive organic and inorganic substances.
Also, because the population of coastal areas is increasing, the volume of
sewage and municipal wastes discharged will increase, thus placing a double
burden upon marine ecosystems.
112
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Very little is known about effects of complex liquid effluents on aquatic
organisms. Data that describe effects are lacking for most substances found
in such wastes, and even when effects are known, it is impossible to predict
how a bioactive substance will behave in the presence of other substances. In
order to estimate the potential effect of a complex waste, chemists and
biologists must work together so that bioactive substances can be identified
and their effects measured.
The purpose of the work reported here was to evaluate the potential
hazard of liquid wastes from industrial plants and one sewage treatment plant
to aquatic organisms. The report discusses use of chemical analyses and
bioassays to evaluate potential impacts of complex wastes on estuarine algae,
seagrass, mysids, shrimp, and fish.
METHODS
Liquid waste samples were collected from the outfalls of industrial
plants and a sewage treatment plant located near estuarine areas and shipped
by air, either in glass jars under ice in insulated containers or uniced in
non-toxic polyethylene containers. Samples were collected and shipped in the
morning and received in the afternoon, when tests were begun.
Upon receipt, each waste was examined for color, odor, suspended matter,
pH, and salinity. It was then divided between the chemistry and bioassay
laboratories for detailed analyses.
Chemical Methods
A fractionation scheme for chemical analysis was devised to be used in
conjunction with bioassays (Figure 1). If a complex waste was toxic, it was
passed through an XAD resin column that adsorbed dissolved organic matter.
Organic compounds were described from the resin with acetone and used in
bioassays. If the organic fraction was toxic, it was subfractionated into
compounds that were extractable with acetone under acidic, basic, and neutral
conditions.
The liquid that passed through the XAD resin column was considered to be
the inorganic fraction. If this fraction was toxic, it was subfractionated
into the heavy metal portion by use of a Dowex1 strongly basic anion exchange
resin, and into the non-heavy metal portion by use of a Dowex strongly basic
cation exchange resin.
Organic and inorganic fractions and subfractions were reconstituted in
artificial seawater before use in biological tests.
1 Dow Chemical Company, Midland, Michigan. Mention of trade names does not
constitute endorsement by the Environmental Protection Agency.
113
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RECEIVE WASTE
BIOASSAY
.TOXia
ORGANIC FRACTION
BIOASSAY.
TOXIC
-NON TOXIC
DISCARD
-NON TOXIC
I
DISCARD
•INORGANIC FRACTION
.BIOASSAY
TOXIC
ACID BASS NEUTRAL
SUBFRACTION SUBFRACTION SUBFRACTION
BIOASSAY
BIOASSAY
BIOASSAY
HEAVY METAL
SUBFRACTION
BIOASSAY
OTHER
SUBFRACTION
BIOASSAY
Figure 1. Integrated chemical
industrial waste.
and biological tests for analysis of complex
Biological Methods
fractionation
pugio, and C.
Bioassays were used to direct the course of chemical
according to the flow-chart in Figure 1. For S. costatum, P.
variegatus methods for bioassay are given by EPA (1977), and for mysids by the
EPA Ocean Disposal Bioassay Working Group (1978). Toxicity to animals is
expressed as the LC50, which is the calculated concentration that would be
lethal to 50% of the exposed animals.
1. Skeletonema costatum, a chain- forming diatom, is known to respond to
toxicants and growth-stimulating substances. Growth was measured as increase
in absorbance at 525 nm and by cell counts. In this report, EC50 is the
calculated concentration that would inhibit growth by 50% as compared to
control growth; the SC2o is the calculated concentration that would stimulate
growth by 20% as compared to control growth.
114
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2. Mysidopsis bahia, a mysid, was used to test for survival of animals
in some bioassays.
3. Palaemonetes pugio, the grass shrimp, was used to test for survival
in textile industry waste.
4. Cyprinodon variegatus, the sheepshead minnow, was also used in
survival tests with textile industry wastes. Only juveniles were used in our
tests.
5. Zostera marina, a seagrass from Chesapeake Bay, was shipped to the
U.S. EPA Environmental Research Laboratory, Gulf Breeze. This species
survived in a healthy state for up to three months in our culture room when
planted in sand in aquaria.
6. Thai assia testudinum, a seagrass that grows in abundance near the
Gulf Breeze laboratory, was also used in phytotoxicity tests.
The method of exposure for seagrass used a 4-liter volume reaction kettle
fitted with a false bottom. Seagrasses were planted in sand in the false
bottom (10 plants per kettle), and the waste was stirred continuously by a
magnetic stirrer. Salinity of waste was increased to 30 parts per thousand by
adding artificial sea salts (Rila Products, Teaneck, NJ). The waste was
diluted with artificial seawater prepared with deionized water and concentra-
tions tested were 25, 50, 75, and 100% water. Controls were prepared from
artificial sea salts and deionized water. Plants were considered dead when
the leaves turned brown and began to disintegrate.
A new computer method to estimate effects of wastes on algal population
growth was used (Walsh et al. , in press). Response of S. costatum was esti-
mated as (1) the calculated concentration that would inhibit growth by 50%
(EC50) and (2) the calculated concentration that would stimulate growth by 20%
(SC2o)- Stimulation or inhibition of algal growth was calculated by plotting
the non-linear regression of absorbance on waste concentration by the equation
Y = A A
1 + Bt (X - C) 1 + Dt (X - E)
where A = mean maximum biomass, B = increasing slope, C = waste concentration
(increasing slope) where Y = 0.5 x A, D = decreasing slope, E = waste concen-
tration (decreasing slope) where Y = 0.5 x A, X = range of waste concentra-
tions tested, and Y = estimated population density over range of X. Calcula-
tions were made on a Digital Equipment Corporation POP 11/45 computer. The
EC50 and SC2o were calculated from the regression curve.
On the average, algal growth stimulation of approximately 20% above the
control value (SC2o)- can be considered to be a statistically significant
increase in growth in these tests. The mean upper 95% confidence interval for
algal growth in control cultures was 17% above the control mean growth value
115
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in 20 textile and other industrial effluent assays. Therefore, growth stimu-
lation greater than 17% above control was a significant response. To predict
a number that represents a statistically significant increase in growth
requires that the standard deviation of the predicted value also be con-
sidered. For 20 effluent assay analyses of textile and other wastes, 27%
predicted growth stimulation that was significant (a = 0.05} as compared to
growth. This value supports the concept of SC20> which can be used as an
index of significant algaT growth stimulation in effluent bioassays (Banner
and Oglesby, in press).
RESULTS AND DISCUSSION
Whole Wastes
Animals and Algae. Shrimp and fish are generally more sensitive to
individual toxicants such as chlorinated hydrocarbons than algae or plants. ' A
major finding of our work is that algae are better indicators of complex waste
bioactivity in static tests than grass shrimp and sheepshead minnows. In a
comparative study using 14 wastes from the textile industry, all 14 wastes
affected algal growth, whereas only five were toxic to shrimp and fisn (Table
1).
TABLE 1. COMPARISON OF RESPONSE OF Skeletonema costatum, Palaemonetes pugio,
AND Cyprinodon. variegatus TO TEXTILE WASTES (SC20 =• percentage
waste at which growth was stimulated by 20%; EC50 = percentage wast*
at which growth was inhibited by 50%; LC50 = percentage waste lethal
to 50% of the animals; NE = no effect.)
Waste
Code
C
N
T
W
A
B
F
G
K
L
S
U
V
X
E
P
S.
SC20
1.50
NE
2.00
1.50
—
0.50
NE
2.75
1.00
NE
2.25
1.50
21.75
0.50
--
NE
costatum
EC50
76
2.0
66
50
--
NE
84
59
79
1.5
NE
NE
93
NE
--
9.0
P. pugi o
LC50
12.8
26.3
34.5
19.6
21.2
NE
NE
2/
NE
NE
2/
NE
NE
NE
3/
-._
C. variegatus
LC50
69.5
47.5
68.Q
37.5
62.0
NE
NE
3/
NE
4/
2/
NE
NE
NE
I/
~^
10% mortality in 100% waste; f/ 20% mortality in 100% waste;
40% mortality in 100% waste; - 50% mortality in 100% waste.
•»*
116
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Bioactivity toward algae occurred in three patterns: growth stimulation,
growth inhibition, and growth stimulation at low concentrations and growth
inhibition at higher concentrations (Figure 2). Results of all three types
are common and indicate that industrial effluents are sources not only of
toxic materials but also of growth stimulants that may accelerate the eutro-
phication process.
Data generated from bioassays of whole waste can be used to estimate
potential impact relative to other wastes. We compared our data for 14 tex-
tile plant effluents to that reported for the same plants with freshwater
organisms (Selenastrum capricornutum, Daphnia pulex, Pimephales promelas)
(Walsh et aJL , in press) and found that S. costatum was the most sensitive
species in 12 cases. The 14 samples were then ranked according to their
potential impact by relating the volume of discharge to response concentration
according to the expression:
MEU/d =
100
EC50 or SC20
(discharge rate)
where MEU/d = million effective units per day. Sample X, which had a low SC2o
and a high discharge rate, is judged to have the greatest potential for
adverse effect upon its receiving water. Waste W, with a low SC20 and a low
discharge rate, was judged as having a relatively low potential for adverse
effect (Table 2).
TABLE 2. RELATIVE POTENTIAL IMPACT, in MEU/d, OF 14 SECONDARY TEXTILE MILL
WASTES TESTED WITH FRESHWATER AND ESTUARINE ALGAE, CRUSTACEANS, AND
FISHES
Waste
Discharge
m3/d x 103
Most sensitive
organism
Response
Percentage waste
Lowest Value
MEU/d x 10
_3
B
C
F
G
K
L
N
P
S
T
U
V
W
X
4.5
3.8
7.6
7.6
9.5
2.8
3.8
4.5
4.7
2.3
1.1
3.0
1.0
9.8
S.
S.
D.
S.
S.
S.
S.
S.
S.
S.
S.
D.
S.
S.
costatum
costatum
pulex
costatum
costatum
costatum
costatum
costatum
costatum
costatum
costatum
pulex
costatum
costatum
SC20
EC50
SC20
SC20
EC50
EC50
EC50
SC20
SC20
SC20
EC50
SC20
SC20
0.5
1.5
81.7
2.8
1.0
1.5
2.0
9.0
2.3
2.0
1.5
9.4
1.5
0.5
900
253
9
271
950
187
190
50
204
115
73
32
67
1,960
117
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0.30
0.20 -
0.10
£ 0.20
c
10
CM
uj 0.10
o
00
cc
o
en
GO
Q60r
0.45 -
030 -
QI5
SC20= 1.5%
EC 50= 50.0%
60 80 100
EC 50= 2.0 %
20 40 60 80
PERCENTAGE WASTE
100
Figure 2. Effects of three textile mill wastes on growth of Skeletonema
costatum. EC50 = calculated concentration that would inhibit growth
by 50%; SC2o = calculated concentration that would stimulate growth
by 20%. **.
118
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Algal data are not widely used for setting water quality standards
because algae usually are not as sensitive to individual organic pollutants as
Our data show that S. costatum was the most sensi-
the complex wastes tested. It was especially valu-
often contain nutrients and toxicants that do not
We suggest that algal tests should be an integral
invertebrates or fishes.
tive organism to most of
able because such wastes
affect the test animals.
part of testing programs for complex wastes.
Seagrass. Seagrasses are important components of estuarine ecosystems,
and we have just begun a long-term study of effects of pollutants on them. In
preliminary experiments, seagrasses were exposed to six waste effluent
samples, and three of the samples were toxic to them (Table 3). Seagrasses
are declining in many parts of the world, mainly in industrialized countries,
but this is the only study that we are aware of that demonstrates direct
toxicity of industrial wastes to them.
TABLE 3. EFFECTS OF INDUSTRIAL WASTE ON SEAGRASS AFTER THREE WEEKS OF
EXPOSURES
Industry
Species
Effect
Creosoting
Chemical
Steel
T. testudinum
Not toxic
Steel
Chemical
Sewage Plant
T. testudinum All plants dead in all concentrations
T. testudinum
T. testudinum
T. testudinum
Z. marina
(1) Two plants dead in 50% waste
(2) Three plants dead in 75% waste
(3) Four plants dead in 100% waste
(4) All surviving plants in 50, 75, and
100% waste were chlorotic
Not toxic
Not toxic
(1) All plants exposed to 25 and 50%
waste were degraded
(2) All plants exposed to 75 and 100%
waste appeared dead
Field Studies. We received 11 waste samples that were tested in flow-
through bioassays in the field. Field bioassays done in a mobile laboratory
demonstrated that four were toxic and seven were non-toxic to shrimp and fish.
In laboratory bioassays, seven were toxic and four were non-toxic to algae,
shrimp, and fish. Algae responded to all, either by growth stimulation or
growth inhibition. We" recommend, therefore, that expensive and time-consuming
on-site bioassays need not always be done. Laboratory tests identify bio-
active wastes and can be used effectively for screening waste outfalls.
119
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Fractionated Wastes
When toxic whole wastes were fractionated according to the scheme given
in Figure 1, toxicity was found to be in the organic or inorganic fractions,
seldom in both. Table 4 gives results from a few industries whose wastes were
tested with S. costatum and M. bahia. Alga and mysid were generally useful in
identifying toxic fractions and non-toxic wastes.
TABLE 4. RESPONSES OF MYSIDS (M. bahia) AND ALGAE (S. costatum) TO FRACTIONS
OF INDUSTRIAL WASTES (+ = toxic effect; - = no toxic effect)
Industry
Mysid
Algae
Toxic Fraction
Gunpowder
Titanium oxide
Aliphatic amines
Oil refinery
Tall oil products
Phosphoric products
Nylon
Carpeting
Citric acid
Synthetics
Carpeting
Amine products
Textiles
Paper products
Unbleached paper
Wire and Rod
Municipal STP
Heavy metals
Neutral organic
Neutral organic
Heavy metals
Heavy metals
Heavy metals
Heavy metals
Neutral and acid organics
Heavy metals
Fractionation followed by bioassay simplifies identification of toxic
components in complex wastes because only the toxic fraction needs to be
analyzed chemically. For example, waste from a titanium oxide plant was found
to be highly toxic, and all toxicity was present in the heavy metal subfrac-
tion. No other subfraction was toxic. Heavy metals analysis of the heavy
metal subfraction revealed very high concentrations of some metals (Table 5)
and it was recommended that the industrial plant apply control technology to
the metals portion of its effluent. In that way, only a single fraction of
the effluent needed to be treated and highly expensive treatment of whole
waste was avoided.
It is important to test fractions of an effluent because effects of
bioactive substances may not be detected in whole waste. Effects of three
wastes and their fraction on S. costatum are given in Table 6. Whole waste
was toxic in each case, and none stimulated growth. The organic fractions had
no effect on algal growth. Toxicity of each waste was in cationic subfraction
of the inorganic fraction. *<
120
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TABLE 5. HEAVY METALS ANALYSIS OF A TITANIUM OXIDE PLANT WHOSE EFFLUENT
TOXICITY WAS IN THE HEAVY METALS SUBFRACTION (Analysis by induc-
tively coupled argon plasma method)
Element
Concentration, ug/L
Ti
V
Cd
Cr
Cu
Zn
Zr
Fe
Al
23,000
3,800
40
1,800
70
625
200
260,000
18,000
TABLE 6. EFFECTS OF INDUSTRIAL WASTES ON GROWTH OF Skeletonema costatum (EC50
= calculated percentage that would inhibit growth by50%;SC20 =
calculated percentage that would stimulate growth by 20%. STP
sewage treatment plant)
Category
Whole Waste Inorganic Fraction
Cations
Anions
EC
50
sc
20
-so
SC
20
EC
0
SC
20
Chemicals
Fibers
STP
8.
15.
14.
2
4
0
none
none
none
9.
22.
30.
6
1
9
none
0.4
none
16.
16.
20.
0
5
0
none
none
none
none
none
none
1.0
0.9
5.6
For the fiber plant, the inorganic fraction was highly stimulatory (SC2o
= 0.4%) when separated from the organic fraction. The anionic subfraction of
each waste was highly stimulatory to algal growth when separated from the
heavy metals. These observations show that effects of growth stimulants may
not be detected when toxicants or other substances are present. In receiving
waters, where toxicants and stimulants may be separated spatially and
transformed biologically or chemically, algal growth promotion, resulting from
eutrophic conditions, could be the most significant effect of a waste stream.
We suggest that chemical fractionation and algal and other bioassays
should be done in order to assess potential effects of a liquid waste stream
on receiving waters.
121
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SUMMARY
Industrial effluents are major sources of pollution in aquatic eco-
systems. Chemical analyses alone cannot predict effects of such complex
wastes (Walsh et al., in press), and it is necessary that bioassays be done on
algae, plants, and animals if the potential impact of a waste is to be esti-
mated. This is especially true when both algal growth stimulators and toxi-
cants are present in a single waste.
Bioassays of whole complex-waste effluent is useful for regulation by law
enforcement agencies. However, effects of stimulants or toxicants may be
reduced by the presence of other substances. Since components of bioactive
complex wastes may become separated in receiving water, their organic and
inorganic components should be fractionated in the laboratory and bioassays on
fractions should be performed to estimate the total potential bioactivity and
to identify the bioactive substances. After these substances are identified,
technology may be applied to the industrial process for their removal.
REFERENCES
Bahner, L. H. , and J. L. Oglesby. Models for predicting kepone accumulation
and toxicity in laboratory exposures and natural ecosystems. In Envi-
ronment Risk Analysis for Chemicals, R. A. Conway (ed.) Van Notrand Co.,
New York. In press.
U.S. Environmental Protection Agency. 1977. IERL-RTP procedures manual:
level 1. Environmental assessment biological tests for pilot studies.
EPA-600/7-77-043. Industrial Environmental Research Laboratory, Research
Triangle Park, North Carolina.
U.S. Environmental Protection Agency. 1978. Bioassay procedures for the
ocean disposal permit program. EPA-600/9-78-010. Ocean Disposal
Bioassay Working Group, Cincinnati, Ohio.
Walsh, G. E., W. B. Horning, and L. H. Bahner. Toxicity of textile mill
effluents to freshwater and estuarine algae, crustaceans, and fishes.
Environ. Pollut. In press.
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ECOLOGIC ASPECTS OF USING CHEMICAL AGENTS FOR ELIMINATING THE
RESULTS OF OIL SPILLS IN THE OCEAN
M. P. Nesterova
P.P. Shershov Oceanology Institute
USSR Academy of Sciences
Eliminating oil pollution in the ocean is one of the major ecologic
problems in environmental protection.
Petroleum and petroleum products are the most common pollutants of the
world's oceans. It is sufficient to say that emulsified and dissolved petro-
leum products have been found in half of the 28,000 samples taken in different
areas of the Pacific and Indian Oceans by ships from the AN SSSR Institute of
Oceanology. During an international MOK and VMO experimental project to
detect oil films on the surface of the seas and oceans.
Of the 3500 miles covered in the northwest Pacific, 315 contained an oil
film, as determined by a remote procedure conducted from the side of a ship.
Studies (Simonov et al. 1974) have shown there is significant pollution from
petroleum products in the northern part of the Atlantic Ocean. This has been
confirmed by the studies of the AKADEMIK KURCHATOV research ship (January-June
1978) in the Sargasso Sea and in the region of the Gulf Stream, where 83 trawl
nets took samples for determining the concentration of tar lumps; only six of
the 75 trawl nets contained no tar lumps in the Sargasso Sea. Eight trawl
nets were used for cutting through the Gulf Stream, and only two contained tar
lumps.
The external appearance of the lumps differed strongly; some were sticky,
fairly fresh petroleum products, but most were dense, structurized coagulates;
there were brown films in some places. Young goose barnacles, Bryozoa, algae
and other organisms were often found inside the lumps
The presence of different types of oil and petroleum products in the
waters of the world's oceans indicates the predominance of accumulation over
biochemical degradation.
Petroleum pollution must have an effect on the physical, chemical and
biologic processes of the seas and oceans. The destructive effect of oil on
hydrobionts is well known. Reports of the effect of oil on the inhabitants of
the ocean began to appear in the literature in the 19th century. Many
investigators probed this question at the beginning of the 20th century, and
extensive literature appeared. At the beginning of the '70s, Nelson-Smith
reported more than 800-studies on the different aspects of the biologic effect
of oil (Nelson-Smith 1970, 1972). The oil film on the surface of the ocean
also disturbs exchange of energy, heat, moisture and gases between the ocean
123
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and the atmosphere. We found experimentally that evaporation had decreased to
45 t after one hour in one square mile of ocean surface, i.e., by almost two
times. The presence of an oil film over a large surface can apparently also
affect weather conditions (Buynitskiy 1976).
The marine fleet is the basic source of petroleum pollution of the world's
oceans (approximately 33%), and a particularly important fact is that this
pollution occurs in normal operating conditions. Eliminating this source of
pollution would only be possible by improving technological processes and
control methods. However, even the introduction of new technology and
improved operating conditions cannot totally prevent accidents and
catastrophes similar to the AMICO-CADIZ and hundreds of others. Oil spills,
sometimes very large ones, can naturally also occur in underwater fields.
Accidents in offshore drilling and tankers are the most dangerous and
seriously damage the ecological systems in individual regions of the world's
oceans. It has been found that 20-30% of the total pollution of the marine
environment is caused by oil spills.
The oil which enters the sea as a result of spontaneous spillage, and the
effect of waves and wind result in the formation of films of different thick-
nesses on the water surface. This is a primary, unstable form of oil in sea
water, since it is constantly changing under the effect of the surrounding
environment. First, the volatile components of petroleum and petroleum pro-
ducts evaporate very intensively under the effect of strong winds, swells and
temperature, and during evaporation crude oil can lose up to 50% of its com-
ponents, diesel fuel up to 75% and mazut from 5-9%. Evaporation begins when
the oil first enters the ocean, proceeds at a decreasing rate, and then
becomes insignificant and is almost completely over in a few days. However,
it is not true that evaporation stops the toxic effect of the oil on ecologic
associations. Conversion of the components of the spilled oil to the gas
phase results in pollution of the atmosphere, primarily the layer adjacent to
the water, and does not prevent the components from re-entering the water.
Very little spilled oil is dissolved in sea water (up to 5%); the largest
amount undergoes emulsification and dispersion. Depending on hydrodynamic
processes and meteorologic conditions, physicochemical dispersion can rapidly
(in a few hours) eliminate 15% of the spilled oil. Although dispersion
accelerates biochemical oxidation, the capacity of a body of water for self-
cleaning is not infinite. And, as the emulsified and dispersed oil decom-
poses, compounds which are more toxic to the inhabitants of the ocean than the
oil itself can be formed. In the absence of turbulence, dispersed oil can
again form oil films on the ocean surface. Dispersed oil can be absorbed on
mineral suspensions and plankton and enter the deeper layers of the ocean with
these carriers.
Dispersed oil is particularly rich in surface-active substances and forms
"oil in water" emulsions; however, in many cases where oil accumulates in
large quantities on a limited surface, it forms an emulsion in mixing with the
water, "water in oil," since the high-molecular compounds it contains (tars,
asphalts, etc.) stabilize emulsions of this type. Such highly viscous cross-
linked formations ("chocolate mousse," tar lumps) can remain on the water
surface for a long time; when the mineral suspension condenses and is absorbed,
124
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they can sink to the bottom or be tossed on shore, causing considerable damage
both to benthic organisms and coastal fauna and flora. Some of the methods
proposed for eliminating oil spills are based on this principle, i.e., sinking
the oil with specially introduced mineral hydrophobic additives (hydrophobic
sand, carbonaceous materials, etc.). Naturally, this method is not suitable.
The quantitative ratio of migratory forms of petroleum hydrocarbons in
the different regions of the seas and oceans also varies within very broad
limits, but can be used as some indication of the processes which result from
an oil spill. For example, the 26th passage of the research vessel AKADEMIK
KURCHATOV in the central part of the Baltic Sea in 1978 revealed that 3.4% of
petroleum hydrocarbons was found in the film, 15.6% in residue and 81% in the
water; 12% was dissolved, 10% was in a colloidal-disperse state, and 78% was
emulsified. This indicates redistribution of the petroleum hydrocarbons under
the effect of the environment (a significant effect) even when there were no
changes in the chemical composition of the oil.
However, conversion of oil from one form to another cannot be considered
utilization or elimination of the consequences of oil spills in the sea. Each
migration form of petroleum hydrocarbons will have a negative effect on the
ecology of the region. For this reason, methods and agents to achieve maximum
removal of the oil pollution resulting from spills in the marine environment
are necessary. The AN SSSR P.O. Shershov Institute of Oceanology adhered to
these concepts when developing methods and agents for eliminating the con-
sequences of oil spills.
Chemical agents are being developed to enclose an oil spill to prevent it
from spreading, and methods and agents for removing the oil and increasing the
natural biochemical decomposition are being devised. An agent has been
created that involves forming a dense barrier of the foam-plastic type which
prevents a spill from spreading and simultaneously keeps the oil from the
surface of the water. After mechanical removal from the surface of the water,
the foam-plastic could be squeezed out and reused as a sorbent.
In cooperation with the L'vov Polytechnic Institute, oil sorbents based
on plant, mineral and synthetic substances have been developed and investi-
gated. One of the basic requirements for such substances is unsinkability and
the potential for recovering the oil. For example, some plastic foams based
on complex polyesters which absorbed an amount of oil 18-20 times greater than
their weight in 5 minutes have been tested in the laboratory; oil occupied 90%
of their volume, and they could be used many times (Anufriyeva and Nesterova
1976). A sorbent was prepared on a mineral base from distended hydrophobic
volcanic rock-perlite, which can be used for production of bitumen and other
construction materials after the petroleum is absorbed from the surface of the
water.
The tests on hydrophobic perlite and thermally treated peat showed that
they could clean an area of water in a port of 98-99% artificially produced
oil pollution. After most of the oil has been collected mechanically (if a
large quantity is involved), the relatively thin residual film is treated with
substances which can disperse and convert it to a thin dispersion (emulsion of
the "oil in water" type). By eliminating the previously dense oil film,
125
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oxygen exchange between the water and the atmosphere is restored and gradual,
natural biochemical oxidation of the highly dispersed, isolated droplets of
oil suspended in the water layer begins. If the dispersing agent has a high
enough stabilizing capacity with respect to these drops and persists even when
the initial dispersion is greatly diluted by the water, and if it is economic-
ally practical and—more important—not toxic to marine organisms, then this
method of eliminating oil pollution should be a very promising and effective
method for eliminating the consequences of spills. We used surface active
substances as spontaneously active dispersing agents which could destroy a
dense oil film and convert it into a stable emulsion when aqueous solutions
were introduced in the film.
This condition of spontaneous emulsification is necessary, since the
usual methods for forming emulsions, especially highly dispersed emulsions by
mechanical fractionation into emulsified drops, are not suitable in this case.
For this reason, we used one of the properties of surfactants from the group
of nonionic compounds related to their solubility in aqueous and oil
(hydrocarbon) phases of emulsions. In contrast with most ionic surfactants
whose solubility in any phase is a function of their composition, nonionic
surfactants are usually readily soluble in both phases; thus, when they con-
tact an emulsion system, redistribution of the surfactant takes place by
diffusion through the interface. This process, as Taubman previously demon-
strated (1969), causes strong local decreases in the interphase tension and
consequently very pronounced disturbances in the hydrodynamic stability of the
easily moved interface. The intensive turbulence in the boundary layers
results in the appearance of highly dispersed emulsions of both types;
however, only the "oil in water" emulsion survives (the hydrophilic emulsifier
only stabilizes emulsions of this type) and a highly dispersed emulsion
appears—a microemulsion.
The microemulsion is stabilized by the adsorption layer of the surfactant
and is a surface structure which in turn stabilizes the "karli" in the basic
microemulsion; as a consequence, the entire system is highly resistant to
coalescence.
If we consider the oil film on the surface of the ocean as a separate
phase of relatively very low volume, then it follows from the above that
introducing an aqueous solution of the corresponding nonionic surfactant with
a jet apparently produces the following effect—decomposition of the film into
drops which are stabilized to such a degree that the microemulsion formed will
retain sufficient stability even with "infinite" dilution with water and the
effect of swells and winds.
In basing the emulsion stabilization mechanism on the above, it was
necessary to obtain evidence of the possibility of obtaining this effect in
actual natural conditions with ordinary dispersing agents with respect to the
features of methods of mixing solutions of the agents with the oil film on the
surface of the water, the presence of their electrolytes in the water, the
effect of strong dilution of the solutions, and several other factors.
The results of laboratory studies and field trials confirmed this possi-
bility.
126
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Dispersing agents have been developed in our laboratory--DN-75 oil dis-
perser and EPN-5 oil-film emulsifier (together with the Ufimskiy Institute of
Petroleum Transport and Storage) (Taubman and Nesterova 1977).
The laboratory studies and field trials on these products and on
Correxite-7664 (U.S.) and Berol-198 (Sweden) demonstrated the high efficiency
of the dispersing agents. In trials in the Baltic Sea, Berol-198 decreased it
to 0.06 mg/liter, Correxite-7664 to 1.22 mg/liter and DN-75 to 0.23 mg/liter.
An artificially created petroleum spill (50,000 m3) in the port of
Vladivostok was destroyed by DN-75 in 15 minutes (Nesterova ejt al. 1977).
Using dispersing agents in an aqueous environment naturally has some
effect on the hydrobionts; for this reason, in developing dispersing agents,
the Analytic Laboratory of the AN SSSR Institute of Oceanology first concen-
trated on the toxicologic studies of the products and their constituent parts.
The effect of EPN-5 and DN-75 dispersing agents was studied on hydrobionts of
different trophic levels both in the Institute of Oceanology and in a number
of scientific research institutes in our country: In the Laboratory for
Research on Water Pollution of the Saratov Branch of the State NIORKh, the
Department of Hydrobiology of the biological faculty at MSU, the Laboratory of
Radiation and Chemical Ecology of the VNIRO, the Georgian Branch of the VNIRO
and the DVNTs Institute of Marine Biology.
We shall briefly summarize the results of the studies.
The effect of the dispersing agents on aquatic microflora was traced
based on changes in the number of saprophytic bacteria in the water containing
different concentrations of EPN-5. It was found that EPN-5 undergoes strong
bacterial oxidation in the water which facilitates the vigorous development of
saprophytic bacteria. The presence of EPN-5 in the water, up to a concentra-
tion of 10 g/liter, had no inhibiting effect on multiplication of the bacteria
and the number of saprophytic bacteria in the water containing EPN-5 was many
times higher than the control. These studies showed that EPN-5 in concentra-
tions over 0.1 mg/liter promotes intensive development of saprophytic micro-
flora.
The effect of the dispersing agents on zooplankton was studied using the
water flea. We examined the condition of the water fleas exposed to different
concentrations of EPN-5. The results of the experiments showed that 2 mg/
liter is the inactive concentration of EPN-5, i.e., EPN-5 is moderately toxic
to water fleas.
The effect of solutions of the individual surfactants was investigated—
oxyphos and diproxamine 157 in concentrations of 0.1-10 mg/liter—on the
survival and growth of the young of a Black Sea isopod. All concentrations of
diproxamine 157 tested had no effect on the survival of the animals. The data
confirmed the low toxicity of diproxamine 157; the previously established
maximum admissible concentration (MAC) for fish breeding waters was 3.2 mg/
liter. A 10 mg/liter concentration of oxyphos increased the survival of
idothea by 1.3 times and had a positive effect on the growth of the young
animals.
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The effect of the dispersing agents EPN-5, DN-75, Correxite-7664 and
Berol 198, produced by the MoDoKemi AB Co. , Sweden, and the surfactant diprox-
amine 157 and oxyphos on blue-green algae included the following: Inhibition
of growth of the algae was observed with a concentration of 100 mg/liter of
EPN-5, the nontoxic concentration was 10 mg/liter; for DN-75, these concentra-
tions were one order-of-magnitude higher, respectively. Of the two foreign
samples investigated, Correxite-7664 was less toxic than Berol 198, whose
toxicity to blue-green algae was the same as the domestic dispersing agent
DN-75.
The toxicity of the other surfactant, oxyphos, was the same as that of
EPN-5, i.e., a 100 mg/liter concentration caused inhibition of growth in the
algae, and a 10 mg/liter concentration was nontoxic. In the range of concen-
trations investigated, diproxamine 157 had no toxic effect up to a concentra-
tion of 10 g/liter.
The studies conducted by the Laboratory of Radiation and Chemical Ecology
of the VNIRO showed that unicellular algae were the most sensitive hydrobionts
to low concentrations of toxic substances. Their photosynthesis was inhibited
by minute quantities of toxic substances which caused no pronounced physio-
logic anomalies in other organisms. The experiments conducted on Black Sea
and Caspian species of unicellular algae showed that a concentration of 100
mg/ liter of EPN-5 caused coagulation of the protoplasm in diatomaceous algae,
i.e., actually the death of the cells. The Caspian Ankistrodismus convolutus
and the Black Sea Coscinodiscus granii were most resistant to EPN-5; 10 mg/
liter concentrations did not harm these algae. Peridium, green and one of the
diatomaceous algae were most sensitive to EPN-5. Only the 0.5 mg/liter con-
centration of EPN-5 produced optimal results for these three species of algae.
Similar studies with the same species of algae were conducted with solutions
of Berol 198. The harmless concentrations of this preparation were an order-
of-magnitude higher than for EPN-5.
Acute, subacute and chronic experiments were conducted on Chironomid
larvae at the Saratov branch of the Gos NIORKh to determine the maximum admis-
sible concentration of EPN-5. The experimental studies showed that the 2.5
nig/liter concentration had no negative effect on metamorphosis of the larvae.
The effect of EPN-5 on fish eggs and larvae was investigated at the same
institute. Eggs and larvae of one species of Neopterygii—pike—and one
species of cartilaginous ganoids—Russian sturgeon—were used as the test
objects in investigating the effect of EPN-5 on fish in the embryonic and
postembryonic period of development. The studies showed that the maximum
admissible concentration of EPN-5 had no negative effect on the physiologic
condition or growth rate of pike larvae and could be considered inactive for
pike in the early postembryonic stage of development.
The results of the observations of Russian sturgeon eggs and larvae
showed that only a concentration of 0.09 mg/liter had a negative effect on the
physiological condition and growth rate of sturgeon and can be considered
inactive for sturgeon in the early postembryonic stage of development.
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The effect of EPN-5 on adult fish was determined in this year's brood of
rainbow trout. The following liminal concentrations which were harmless to
the fish based on the following indices were determined: survival 5.6 mg/
liter; clinical and pathologic condition 2.8 mg/liter; growth rate 1.4 mg/
liter; blood 0.8 mg/liter. Thus, the inactive concentration of EPN-5 for
trout is 0.8 mg/liter.
The toxicologic effect of EPN-5 and the dispersing agent DN-75 was inves-
tigated at the Georgian branch of the VNIRO on another species of fish—horse
mackerel. It was found that DN-75 was less toxic than EPN-5. The MAC for
DN-75 was 0.005 mg/liter.
An analysis of the data obtained in studying the effect of EPN-5 on the
physiocochemical properties of water, the toxicity for feeding organisms,
eggs, larvae and fish showed that fish larvae are the most sensitive com-
ponent. A maximum admissible concentration of 0.09 mg/liter, the boundary
toxicological index, was established for EPN-5 in fish breeding waters.
Studies on the hygienic basis of the maximum admissible concentrations of
dispersing agents in bodies of water were conducted at the same time as the
studies of the effect of the dispersing agents on hydrobionts. The effect of
dispersing agents on the organoleptic characteristics of water and the overall
sanitary regime of the reservoirs was investigated.
In generalizing all of the available data on the effect of dispersing
agents on hydrobionts, we can conclude that the effect of dispersing agents is
both a function of the chemical composition of the substance and the species
of biologic object exposed to the dispersing agent. For this reason, only
comprehensive studies of the toxicological effect on different levels of the
trophic chain will make it possible to recommend chemical agents, particularly
dispersing agents, for use on a scientific basis. The scientific and techni-
cal studies should be extended with respect to creating agents and methods for
eliminating the consequences of oil spills. Correct recommendations and
measures for eliminating the consequences of oil spills can only be developed
on the basis of fundamental, complex studies of the processes which take place
in the sea and on a precise and continuous evaluation of the state of pollu-
tion of marine waters.
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REFERENCES
Anufriyeva, N. M. , and M. P. Nesterova. 1976. Study of the plastic foam
polyurethane as an agent for removing oil from the surface of water.
Vodnyye resursy 4:149.
Buynitskiy, V. Kh. 1976. Some socioeconomic problems in environmental
pollution related to scientific-technologic progress. Vestnik LGU 12:7.
Nesterova, M. P. , 0. S. Mochalova, and N. M. Antonova. 1977. Chemical
agents for eliminating oil pollution of the world's oceans. In Proc. of
the 1st Conference of Soviet Oceanologists, Moscow.
Nelson-Smith. 1973. Petroleum pollution of the sea. Gidrometeoizdat,
Leningrad.
Simonov, A. I., S. G. Oradovskiy, and A. A. Yushchak. 1974. The current
state of chemical pollution of Northern Atlantic waters. Meteorologiya i
gidrologiya 3:6-69.
Taubman, A. B. , S. A. Nikitina, and V. I. Prigorodov. 1969. The role of
quasi spontaneous emulsification in stabilizing emulsions. Kol. zhur.
(27)2:291-292.
Taubman, A. B. , and M. P. Nesterova. 1977. Physicochemical properties of
the emulsifying effect of surfactants applied to problems of marine
ecology. Jji Physicochemical principles of using surfactants, Tashkent.
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CHEMICAL POLLUTION OF THE FILM LAYER OF
THE PACIFIC OCEAN
A. I. Simonov and V. I. Mikhaylov
State Oceanographic Institute
Moscow
The first survey of studies on the chemical pollution of Pacific Ocean
waters was completed in 1975. As a result of this effort, Soviet ocean-
ologists were able to draw a number of general conclusions concerning the
nature of marine pollution (Oradovskiy et a1. 1975; Simonov et al. 1974).
These conclusions have now been confirmed by many studies conducted in
different countries which have greatly facilitated determination of strategies
for monitoring pollution on national and international levels. We shall
mention three basic conclusions. First, it was found that the pollution of
the Pacific Ocean, primarily pollution from petroleum hydrocarbons (PH), i.e.,
petroleum, petroleum products and the products of their decomposition in sea
water, unsaturated hydrocarbons, is planetary in nature. Second, the vast
role of oceanic circulation in transferring and distributing PH, including in
relatively clean areas of the Pacific, for example, in the northern Arctic
Ocean, was demonstrated. Third, it was shown that chemical pollution
significantly affects primary production, and accelerates decomposition of
living organic substances.
However, there were gaps in the survey stage of the studies, basically
caused by the lack of methods and technical means for observations and
determinations. One of these lacunae consisted of the fact that the role of
boundary surfaces (the surface of the ocean, the bottom, the layer of density
discontinuity) in the accumulation and evolution of pollutants was not
investigated for these reasons.
The results of studies of pollution in the surface layer of the Pacific
Ocean by petroleum hydrocarbons and other pollutants are communicated in the
present article. The studies were conducted in the Northern Atlantic on
scientific-research weather ships from the State Oceanographic Institute in
1976-1977. These studies were important both with respect to the prediction
of the level of pollution in ocean waters and with respect to evaluating its
effect on the basic physical, chemical and biological processes and energy-
heat-moisture and gas exchange between the ocean and the atmosphere.
The observations of pollution in the surface microlayer (abbreviated SML)
of the water were conducted in broad areas of the North Atlantic, including
the Canary and Northern Tradewind currents, the Gulfstream, the Sargasso Sea
and the Northern Atlantic current system. The SML consists of the thin sur-
face layer 100 micrometers thick. Samples were taken in this microlayer
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with a grid sampler constructed according to the principle proposed by Garret
(1972). We note that the samples of water from the SML were taken only in
those cases where signs of petroleum hydrocarbons in the visibile frequency
range were not observed in the surface.
The grid sampler is a capron sieve with capron thread thickness of
approximately 0.2 mm and cell area equal to 1 mm2. The sieve is mounted on a
frame which is not subject to corrosion. Grid samples with a total area of
0.35-1.20 m2 were used in the study.
In taking samples of water from the SML, special precautions were taken
to prevent the ship from becoming a source of pollution. Thus, samples of sea
water were taken immediately after the ship stopped and discharge of all waste
and ballast water was terminated 15 minutes before sampling began and was not
resumed until sampling ended; washing using all types of discharge devices was
prohibited.
The water taken from the SML was analyzed for PH, synthetic surfactants
(SS) and organochlorine pesticides (OCR) in ship and shore laboratories using
the procedures described in the "Handbook of Methods ..." (1972).
Samples of air were simultaneously taken at the individual stations for
determination of PH using devices which allow both taking air samples from the
atmosphere and preparing an extract in which the concentration of PH is
subsequently determined from the side of the drifting ship and during move-
ment. The apparatus consists of the following: a pump device—a VK-I
microcompressor with output of 2 liters/min; equipment for removing water
vapor from the air samples and eliminating different types of impurities--
U-shaped calcium chloride tubes with different filters and extractor vessels.
Drexel flasks with atomizers were used for extraction. The microcompressor
was installed on the windward side of the ship to exclude the possibility of
petroleum hydrocarbon products from the ship itself. The air was pumped by
the microcompressor for 0.5-1 hour to concentrate the amounts of PH in the
extract; this could be determined by the currently used methods of IR
spectrometry and gas chromatography. The air pumped by the microcompressor
first passed through a desiccation tube filled with desiccated and PH-free
sodium sulfate, and then a filtering calcium chloride tube containing calcined
and purified aluminum oxide. Impurities were trapped in this tube, including
polar hydrocarbons. The PH were extracted in the Drexel flask with 40 ml of
OSCh brand carbon tetrachloride (CC14). The air stream containing nonpolar
petroleum hydrocarbons was sprayed on the layer of carbon tetrachloride and
subsequently passed into solution. Complete extraction was obtained by
passing the air stream through a second Drexel flask also filled with 40 ml of
CC14. After sampling and CC14 extraction had ended, the hydrocarbon-saturated
petroleum from the air was poured from the two Drexel flasks into one flask
for extracts. Standard solutions were prepared, calibration curves were
plotted, and the CC14 extracts were analyzed using the method for determina-
tion of petroleum products in sea water (Handbook of Methods ... 1977).
Half of the water sampled with the grid sampler in the SML and with a
vessel in the sub-superficial layer was filtered through SYNPOR No. 2 membrane
filter with pore size of 2.5 microns and total diameter of 65 mm for studying
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the suspended and dissolved PH in the SML. We assumed that filtration through
a filter with this pore size would trap most of the suspended substances, both
living and dead. The concentration of dissolved PH was determined based on
the difference between the total concentration in the surface microlayer and
in the suspended form.
In taking samples from the 0-1.0 m subsurface layer, water from the
surface microlayer directly in contact with the atmosphere was prevented from
entering the bottle. The sample of water from this layer was taken in a
heavy-based 5-liter bottle which was sealed with a stopper before it was
submerged; the bottle was lowered and then opened.
The extracts of the water samples taken from the subsurface layer, the
SML and also the filters with adsorbed suspended substances from the SML were
prepared for determination of PH immediately after sampling ended based on
practical recommendations (Methods ... 1977). The extracts were analyzed on
the ship 1-1.5 hours after extraction began; the concentrations of PH in the
extracts of the samples of water and air and on the filters were measured by a
model OIL-102 (Yanagimoto, Japan) IR spectrometer. This device allows eval-
uating the total concentration of petroleum hydrocarbons without identifying
their constituent components.
POLLUTION FROM PETROLEUM HYDROCARBONS (PH)
The analysis of the numerous data (Table 1) indicated that the surface
microlayer is a powerful concentrator of petroleum hydrocarbons and also of
synthetic surfactants and organochlorine pesticides. The concentration of PH
in the SML was generally one to two orders-of-magnitude higher than in the
subsurface layer. The high concentration of PH in the SML is due to their
physical and chemical properties, primarily the slightly lower specific
gravity in comparison to the specific gravity of the sea water and their
insignificant solubility. The fact that the sources and channels of PH entry
in sea water gravitate to the surface of the sea and that their entry into the
marine environment is apparently sufficiently regular so that the run-off of
PH from the SML is constantly compensated (in the atmosphere by evaporation of
the light fractions and with light spray in the water in dissolution or
discharge of suspended lumps) plays a significant role.
The range of changes in the concentrations of PH in the SML is rather
broad: from tenths of mgl-1 to 15 mgl-1 with average concentrations from 1
mgl-1 to 5.5 mgl-1. To describe the size of these concentrations it is suf-
ficient to recall that the value for the maximum admissible concentration
(MAC) for PH in reservoirs used as fisheries in the Soviet Union is a total of
only 0.05 mgl-1.
The second conclusion which can be drawn from the analysis of the data is
that there is a decrease in total concentration of PH in the SML and a simul-
taneous absolute and relative increase in the suspended PH and a decrease in
the dissolved PH fraction (Table 2) as the distance increases from the shore
and the shelf where "the primary sources and input channels of PH are located.
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TABLE 1. FEATURES O.F PETROLEUM HYDROCARBONS IN THE SURFACE MICROLAYER AND
IN THE 1-METER LEVEL IN DIFFERENT AREAS OF THE NORTHERN ATLANTIC
Level
No. of
Determinations
Range of Changes
in Concentrations
mgl-1
Mean
Concentration
mgl-1
Mean Square
Deviation
mgl-1
20 passages "Tradewind" in the Canary current region
(August 1976)
SML
1 m
38
38
1.2-15.0
0.0-0.20
5.49
0.06
0.04
0.001
15 passages "Monsoon" in the Northern tradewind current region
(Winter 1975)
SML
1 m
121
121
0.2-2.60
0.0-0.15
1.40
0.03
0.08
0.001
23 passages "Monsoon" in the northeastern region of the North Atlantic
(Winter 1977)
SML
1 m
99
99
0.20-2.87
0.00-0.05
0.95
0.001
0.06
0.0001
This is due chiefly to evaporation of PH with low molecular weights in the
atmosphere and conversion of part of the dissolved fraction into a suspended
fraction in the water.
Table 2 indicates that the mean concentration of the soluble PH fraction
changed from 2.7 mg/liter at the shelf to 0.4 mg/liter away from shore, and
the mean relative change was 58% (from 83% to 25%). This was due to an
absolute decrease in the total concentration of PH in the SML in the open
ocean. The mean relative decrease in the total concentration of PH in the SML
away from the shelf in the open ocean was approximately 40%. The mean total
concentration of PH on the shelf was actually 3.1 mg/liter and only 1.7
mg/liter at the boundaries of the sections in the open sea (data from the last
column in the second row in Table 2). The mean concentration of the suspended
PH fraction simultaneously increased from 0.4 to 1.3 mg/liter due to con-
version of part of the dissolved PH fraction into the suspended fraction, and
the mean relative concentration of the suspended PH fraction increased from
17% at the shelf to 75% in the open sea.
As the distance from the shelf increased, the decrease in the concentra-
tion of the dissolved PH fraction in the SML caused a decrease in the total
concentration and simultaneous increase in the concentration of the suspended
fraction.
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TABLE 2. CHANGE IN CONCENTRATION OF SUSPENDED AND DISSOLVED
FRACTION OF PETROLEUM HYDROCARBONS IN THE SURFACE MICRO-
LAYER IN THE DIRECTION OF THE SHELF ZONE
Section
(Conditional
Numbering)*
Concentration of suspended (top line) and dissolved
(bottom line) fractions of petroleum hydrocarbons
mg/liter
No.
No.
No.
No.
No.
No.
No.
Interpretation of the sections:
% of Total Concentration
1 0.2-2.0
8.0-1.0
2 0.4-1.5
3.0-1.0
3 0.6-1.2
1.6-0.2
4 0.2-1.4
1.2-0.0
5 1.0-1.6
1.8-0.6
6 0.0-0.4
1.6-0.2
7 0.2-0.6
0.8-0.1
Mean 0.4-1.3
2.7-0.4
2-67
98-33
18-60
82-40
27-86
73-14
14-100
86-0
36-73
64-27
0-67
100-33
20-86
80-14
17-75
83-25
1
2
3
4
5
6
7
long.
from the Straits of Gibraltar to the Cape Verde Islands.
from the Africa shelf 25° north lat. to 40° west long.
from the South America shelf 32° north lat. to 45° west
from the Scotland shelf to the Greenland shelf
from the Island shelf to a point 60° north lat. and 25° west
from the Norway shelf to the Faeroe Islands
from the Straits of Gibraltar to a point 36° north lat. and 35° west
long.
long.
It was suggested above that the decrease in the concentration of the
dissolved PH fraction in the SML was caused by their partial evaporation in
the atmosphere, dissolution in the underlying layers during dispersion and
transfer of part of the dissolved fraction into the suspended fraction in the
SML; the latter is supported by the data in Table 2 to a significant degree.
For this reason, it was extremely important to determine if there are quali-
tative or even quantitative correlations between the concentration of PH in
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the SML on the one hand, and their concentration in the adjacent layer of the
atmosphere and the subsurface layer of water, respectively, on the other hand.
It is necessary to remember that the concentrations of PH in the water are 1
to 2 orders-of-magnitude lower, and the concentrations in the air are 2
orders-of-magnitude lower than in the SML.
The analysis of the observations conducted in parallel on the SML at a
depth of under 1 m and in the air at a height of 0.5 m showed that there is no
correlation between the total concentration of PH in the SML, the air and the
subsurface layer. This could be due to the different fractional composition
of the PH in these media (predominance of the suspended fraction in the water
and volatile components in the air), the different additional routes by which
they can enter these media (advection of PH in water, in addition to their
entry to the SML), etc. In addition, a detailed analysis resulted in the
conclusion that there are definite direct correlations (Figure 1) between the
concentration of the dissolved PH fraction and the concentration of PH in the
atmospheric layer adjacent to the water. These correlations are not identical
for different areas of the ocean, however. It is characteristic that the
tangent of the angle of inclination of these relations decreases as the
distance from the shelf increases and has a smaller value in the open ocean,
in the area of oceanic weather station "C" with coordinates of 52°45'N lat.
and 35°30'W long. This correlation essentially loses its significance in the
open part of the ocean.
mg/liter IN THE SML DISSOLVED PART
Figure 1. Graph of the correlation oetween the concentrations of petroleum
hydrocarbons in the atmospheric layer adjacent to the water and the
dissolved petroleum hydrocarbons in the surface microlayer of the
water based on observations: I - northeast section of the Atlantic
(near Ireland); II - North Sea; III southern Iceland shelf; IV -
oceanic weather station "C". *,.
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The reason for the decrease in the degree of the effects of the PH
concentrations in the SML on their concentration in the atmospheric layer
adjacent to the water as the distance between the observation area and the
shore increases is based on the fact that freshly discharged PH in which
volatile fractions predominate usually predominate in the coastal regions,
while heavy, non-volatile fractions predominate away from coastal regions.
It is believed that the difference in the correlation is caused by other
factors: anemobaric conditions, moisture in the air which increases the
partial pressure of the PH in the air, the difference in the PH in the SML and
the air, turbulent conditions in the boundary layer, etc.
Three circumstances have been determined: A significantly lower con-
centration of PH in the air than in the SML, a decrease in the soluble PH
fraction in the SML as the distance from the shore increases, and the presence
of a correlation between the concentration of PH in the air and the concentra-
tion of the soluble PH fraction in the SML have resulted in a quantitative
evaluation of the process in which part of the PH is transferred from the SML
to the atmospheric layer adjacent to the water. This does not mean that the
process cannot proceed in the opposite direction, e.g., intake of PH from the
atmosphere with precipitation.
In concluding this section, we note that establishing the fact of the
high concentration and universal high concentration of PH in the SML will
result in an orientational quantitative evaluation of the concentration of PH
on the surface of the Pacific Ocean. We hypothesize that the mean concentra-
tion of PH in the SML in Pacific Ocean waters will fluctuate from 2 to 3
mg/liter. Of course, it could be significantly lower in certain areas.
However, this hypothesis is based on the presence of petroleum films in many
areas containing a concentration of PH which is many times higher than the
concentration indicated above. Based on these petroleum films, we hypothesize
that the average thickness of the SML varies from 250-750 urn. On the basis of
these hypotheses, the concentrations of PH in the SML in all water areas in
the Pacific Ocean could be estimated at 0.54-2.20 million tons.
Approximately 5.5 million tons of petroleum and petroleum products enter
the Pacific Ocean each year (Goldberg 1976) as a result of anthropogenic
activity. The value for PH contained in the SML calculated above represents
10-35% of the annual discharge into oceans and seas. Of course, it is dif-
ficult to give preference to any extreme value for the relative estimation of
the concentration of PH on the surface of the Pacific Ocean. In any case, the
amount of PH in the SML seems to be very large. This circumstance emphasizes
the role of the boundary surfaces in the powerful concentration of PH and
indicates the need to study their concentration in the other two boundary
surfaces: in the bottom sediment and in the density discontinuity layers.
This in turn would facilitate the development of strategies for monitoring
petroleum pollution in the marine environment. Knowing the amount of PH
contained in the SML (M = 540,000-2,200,000 tons) and their rate of entry (y =
5.5 million tons a year) would make it possible to calculate the estimated
time (T) the PH remain in the SML. We advance two hypotheses: First, all or
a large part of the PH which enter the sea necessarily pass into a stage of
concentration in the SML. The time they remain in SML can only be correctly
137
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calculated in this case. If we consider the phase of petroleum film
formation, this hypothesis is essentially confirmed. Second, the concentra-
tion of PH in the SML does not change or changes very little with time (a
number of years). In this case, we could use a simple ratio for calculating
the time the PH remain in the SML: T = M:^. The second hypothesis is
apparently not always plausible: One to two decades ago, the concentration of
PH in the SML primarily tended to increase; now. however, the concentration
apparently tends to decrease due to the water conservation measures instituted
in many countries. The time that PH remain in the SML can also be calculated
for the case where the mean concentration of PH changes in time, but the
formula is more complex. However, the quantitative features of the dynamics
involved in the change in PH are not known at the present time.
The time that PH remain in the SML could be from 1-5 months based on the
initial parameters and hypotheses indicated above. The time calculated in
this way is not only in agreement with respect to the order-of-magnitude, but
is also close to the value of the experimentally determined half-life of
petroleum dissolved and dispersed in water (Simonov et a]. 1978). The half-
life of petroleum in sea water at temperatures of O-llPC is equal to approxi-
mately 1.5 months; as the temperature increases, the value decreases: At
18-20°C, the half-life is equal to approximately 20 days, and at 25-30°C—7
days. We again emphasize that the half-life of PH coincides with the time PH
remain in water and that the time segment which is sufficiently close to the
total decomposition time will be greater than the values calculated for T.
Based on the comparisons of the values for T and T, we can assume that
only approximately half of the PH found in the SML can be suspended and dis-
solved in the water at low water temperatures, and again become concentrated
in the density discontinuity layers or in the bottom sediment, while the very
volatile components enter the atmosphere.
The other half of the PH can decompose in the SML. With high water
temperatures, a significant part of the PH, except for the very volatile
components, can decompose directly in the SML. Both cases (together with
turbulent diffusion) would completely explain the decrease in the concentra-
tion of PH in the SML as the distance from a shelf (pollution foci) to the
open ocean increases, and would also indicate the possibility of significant
accumulation of PH at mean and high latitudes and, on the contrary, the low
accumulation in equatorial, tropical, and subtropical zones.
Comparision of the values for T and T also leads to the conclusion that
the SML is not only a powerful PH concentrator, but is also a PH filter, since
a significant part of the PH can decompose in this layer, protecting the mass
of the ocean from intensive pollution. This conclusion is extremely important
in correctly calculating the basic components in the accounts of PH pollution
of the marine environment reflecting the dynamics of the levels of pollution.
Up to now, the concentration of a significant part of PH in the comparatively
small volume of Pacific Ocean water contained in the SML, consisting of less
than 1% of the total volume of water, and their degradation in the SML have
not been calculated in the expenditure part of the accounts. The equation for
138
-------
case. If we consider the phase of petroleum film formation, this hypothesis
is essentially confirmed. Second, the concentration of PH in the SML does not
change or changes very little with time (a number of years). In this case, we
could use a simple ratio for calculating the time the PH remain in the SML: T
= M:y. The second hypothesis is apparently not always plausible: One to two
decades ago, the concentration of PH in the SML primarily tended to increase;
now, however, the concentration apparently tends to decrease due to the water
conservation measures instituted in many countries. The time that PH remain
in the SML can also be calculated for the case where the mean concentration of
PH changes in time, but the formula is more complex. However, the quanti-
tative features of the dynamics involved in the change in PH are not known at
the present time.
The time that PH remain in the SML could be from 1-5 months based on the
initial parameters and hypotheses indicated above. The time calculated in
this way is not only in agreement with respect to the order-of-magnitude, but
is also close to the value of the experimentally determined half-life of
petroleum dissolved and dispersed in water (Simonov et al. 1978). The half-
life of petroleum in sea water at temperatures of O-llPc is equal to approxi-
mately 1.5 months; as the temperature increases, the value decreases: At
18-20°C, the half-life is equal to approximately 20 days, and at 25-30°C--7
days. We again emphasize that the half-life of PH coincides with the time PH
remain in water and that the time segment which is sufficiently close to the
total decomposition time will be greater than the values calculated for T.
Based on the comparisons of the values for T and T, we can assume that
only approximately half of the PH found in the SML can be suspended and dis-
solved in the water at low water temperatures, and again become concentrated
in the density discontinuity layers or in the bottom sediment, while the very
volatile components enter the atmosphere.
The other half of the PH can decompose in the SML. With high water
temperatures, a significant part of the PH, except for the very volatile
components, can decompose directly in the SML. Both cases (together with
turbulent diffusion) would completely explain the decrease in the concentra-
tion of PH in the SML as the distance from a shelf (pollution foci) to the
open ocean increases, and would also indicate the possibility of significant
accumulation of PH at mean and high latitudes and, on the contrary, the low
accumulation in equatorial, tropical, and subtropical zones.
Comparision of the values for T and T also leads to the conclusion that
the SML is not only a powerful PH concentrator, but is also a PH filter, since
a significant part of the PH can decompose in this layer, protecting the mass
of the ocean from intensive pollution. This conclusion is extremely important
in correctly calculating the basic components in the accounts of PH pollution
of the marine environment reflecting the dynamics of the levels of pollution.
Up to now, the concentration of a significant part of PH in the comparatively
small volume of Pacific Ocean water contained in the SML, consisting of less
than 1% of the total volume of water, and their degradation in the SML have
not been calculated in the expenditure part of the accounts. The equation for
139
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the accounts (Simonov et a_1. 1978) should apparently be made more accurate by
introducing an additional component—chemical and biochemical decomposition of
PH in the SML.
POLLUTION FROM SYNTHETIC SURFACTANTS (SS) AND
ORGANOCHLORINE PESTICIDES (OCP)
As in the case of petroleum hydrocarbons, an analysis of the data allows
the conclusion that significant quantities of synthetic surfactants, which
enter the marine environment with discharge of wastes from the shore, river
waters and ships, are concentrated in the surface microlayer. The concentra-
tion of SS in the SML could be 1 to 2 orders-of-magnitude higher than in the
subsurface layer. This can be seen from the features of SS shown in Table 3.
The range of changes in the concentration of SS in the SML reaches 1000
ug/liter with an average value of 820 ug/liter. Similar to the case of PH,
the mean concentration of SS is many times higher than the MAC, which is set
at 100 (jg/liter. In addition, the concentration of SS in the subsurface layer
is significantly lower than the MAC.
TABLE 3. CHARACTERISTICS OF SS IN THE SML AND AT THE 1-METER LEVEL IN THE
ATLANTIC OCEAN, ON THE NORTHWESTERN COAST OF AFRICA (in g/liter; 20
passages of the TRADEWIND in August 1976)
No. of Range of Mean Mean Square Interval of
Level Determinations Changes Concentration Deviation Significance
SML
1 m
38
38
180-1250
15-60
820
34
10.1
5.1
820±12.4
34±6.3
The same features are observed in the spatial distribution of SS as for
PH: The concentration decreases as the distance from the shore increases.
Thus, the concentrations of SS were over 1000 ug/liter in the region from the
Straits of Gibraltar to the Canary Islands (longitude 17-18°W), and the
maximum reached 1220 ug/liter. The significant concentration of SS in this
region is related to the effect of the polluted waters in the Straits of
Gibraltar, effluence from the coast of Africa and intensive navigation which
was also found for the high concentrations of PH in the SML discussed above.
In moving to the open ocean, the concentration of SS decreases to 200
pg/liter in the region 20-21°W longitude. The decrease in the concentration
of SS is basically due to their dynamic dispersion. The concentration in the
subsurface layer in going from the Straits of Gibraltar to the open ocean
decreases from 50-60 to 20 ug/liter and less.
All of the characteristic features of the concentration of PH in the SML
and their spatial distribution also apply to SS; this is not only due to the
same sources and channels of PH and SS input, the similarity of some of their
140
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physical and chemical properties (lower specific gravity with respect to sea
water, low solubility), but also to the fact that the concentration of PH
increases due to their emulsification in surface films.
In turning to the characteristics of the concentrations and spatial
distribution of organochlorine pesticides, we emphasize that most of them are
difficult to dissolve in water, are easily soluble in organic solvents and
ethers, and can be adsorbed in marine organisms and suspended substances. In
light of what was shown for PH in the SML and since the suspended matter,
living and dead organisms are concentrated in the SML and since the atmosphere
is the basic means by which OCP enter the marine environment, we can hypothe-
size that the concentration of OCP in the SML should be high, especially in
areas with significant concentrations of PH. The results of studies in 23
passages of the "Monsoon" in the winter of 1977 in the northeastern section of
the Atlantic Ocean confirm this hypothesis. The concentration of OCP in the
SML is a characteristic feature in comparison with the concentration in the
subsurface layer (Table 4).
TABLE 4. MEAN CONCENTRATIONS (top figure) AND RANGE OF VARIABILITY (bottom
figure) IN THE CONCENTRATIONS OF PESTICIDES (in ng/liter based on
data from 23 voyages of the NISP MONSOON, Winter 1977)
Level
SML
1 m
DDT
33.7
0.2-132
2.04
0-17.6
DDE
7.6
0.0-51.5
0.53
0-2.0
ODD
7.8
0.0-32.0
1.22
0-4.4
T-HCCH
26.0
0.0-148
0.89
0-2.0
Judging by the mean concentrations, the concentration of all pesticides
in the SML is an order-or-magnitude higher than in the subsurface layer. The
highest concentrations of DDT were found over the Ireland shelf, where they
reached 80 ng/liter (Figure 2a), while they fluctuated from 2-4 ng/liter in
the subsurface layer. As a function of the distance from the North Altantic
current streams, the concentration of DDT sharply decreased and was approxi-
mately 10 mg/liter in the North Atlantic water mass. The decrease in the
concentrations of DDT is not only due to their dynamic dispersion in dissemin-
ation from islands. It is also related to conversion of part of the DDT into
metabolites. Figure 2b shows that the spatial distribution of the total
metabolites (DDD+DDE) is inverse to the distribution of ODD. The total
concentrations of metabolites increases with distance from islands, and
attains maximum values in the waters of the North Atlantic current and the
southern approaches to the Danish Straits.
We note that although approximately 69% of the total DDT, ODD, and DDE is
represented in the mean concentration of DDT in the SML, it decreases to
40-50% in the waters of the North Atlantic current.
141
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50
10 W
50
30
10
60
50
B
60
50
Figure 2. Spatial distribution of DDT (a) and total ODD
northeastern part of the Atlantic Ocean, in
observations in the winter of 1977.
and DDE (b) in the
ng/liter, based on
The fact that the maximum values (20-50 mg/liter) are observed in the
southern approaches to the Danish Straits is characteristic in the spatial
distribution of T-HCCH in the SML; the concentrations of T-HCCH in the SML in
the region of the Faeroe Island shelf and the Faeroe-Scotland straits falls to
zero values. The concentration of T-HCCH in the subsurface changed insignifi-
cantly (from 0 to 2 mg/liter).
Certain general features thus exist in the concentration of PH, SS, and
OCR in the SML and in their spatial distribution. However, in contrast with
the other pollutants, the concentration of OCR in the SML can decrease to zero
in moving away from shelves.
1.
CONCLUSIONS
Based on multiple two-year observations by the COIN NISP in the North
Atlantic:
1.1. A significant concentration of petroleum hydrocarbons, synthetic
surfactants and chlorinated hydrocarbons was found in the surface
microlayer of water; these substances entered the Pacific Ocean in
different ways;
1.2. It was found that the concentrations of these substances in the
surface microlayer was 1 to 2 orders-of-magnitude higher than the
concentrations in the subsurface layer and the atmospheric layer
adjacent to the water;
142
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1.3. The spatial changes in the concentrations of chemical pollutants
were investigated in the surface microlayer of water under the
effect of external physical-chemical factors and current systems; it
was found that they generally decreased as the distance from the
mainland or shelf zones increased, but still remained comparatively
high except for chlorinated hydrocarbons, whose concentration
decreased to traces. The qualitative composition of the pollutants
simultaneously changed as the distance from the shore increased:
The suspended portion absolutely and relatively increased and the
dissolved portion of the petroleum hydrocarbons decreased; the
absolute and relative fractions of DDE and ODD (DDT metabolites)
also increased and the DDT fraction decreased;
1.4. It was shown that high concentrations of petroleum hydrocarbons and
synthetic surfactants in the surface microlayer of the water were
characteristic of all water areas in the North Atlantic, and high
concentrations of chlorinated hydrocarbons were characteristic of
significant expanses in the shelf zones of the North Atlantic.
On this basis, we drew a conclusion concerning the global nature of the
appearance of stable, high concentrations of chemical pollutants in the
surface microlayer of water in the Pacific Ocean and the global nature of
the disturbances in the naturally combined physical-chemical features of
the surface microlayer of the Pacific Ocean under the effect of pol-
lution.
These conclusions indicate the need for the following:
2.1. Organizing systematic observations of the pollution in the surface
microlayer of water both in oceans and in the seas in the Soviet
Union; this would allow more precise calculation of the pollutants
contained in this layer;
2.2. Calculating the amounts of pollutants contained in and decomposing
in the surface microlayer (primarily petroleum hydrocarbons) in
calculating the amounts in seas and predicting the level of pol-
lution in seas using the balance method.
143
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REFERENCES
Garret, W. D. 1972. Impact of natural and surface film on the ocean. Nobel
Symposium 20.* Almqvist and Wiknell, Stockholm, pp. 75-81.
Goldberg, E. D. 1976. The health of the oceans. UNESCO Press, pp. 117-128.
Handbook of methods for the chemical analysis of marine waters. 1977.
Gidrometeoizdat, pp. 118-127, 131-136, 145-155.
Oradovskiy, S. G. , A. I. Simonov, and A. A. Yushchak. 1975. Study of the
distribution of chemical pollutants in the Gulfstream zone and their
effect on primary production of oceanic waters. Meteorologiya i
gidrologiya, 2:48-58.
Simonov, A. I., S. G. Oradovskiy, and A. A. Yushchak. 1974. The current
state of pollution in the North Atlantic. Meteorologiya i gidrologiya,
3:61-69.
Simonov, A. I. , N. A. Afanas'yeva, T. A. Bakum, and B. M. Zatuchnaya. 1978.
Self-purification processes in sea water with respect to chemical pol-
lutants. Trudy GOINa, 128:96-104.
144
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STUDIES ON THE BIOLOGICAL TRANSPORT OF MATERIALS FROM
SURFACE TO DEEP OCEAN WATERS: I. FLUXES OF CARBON,
NITROGEN AND PHOSPHORUS II. FLUXES OF TRACE ELEMENTS
by
G. A. Knauer and J. H. Martin
Moss Landing Marine Laboratories
Moss Landing, California
INTRODUCTION
It has long been recognized that the plankton play an important role in
the biogeochemical cycles of various materials. For this reason, a great deal
of research has been devoted to the analyses of plankton remains (oozes) on
the sea floor (e.g., Arrhenius 1963) and plankton/sediment interaction
(Manheim et al. 1972). The plankton living in the surface waters of the
oceans have also been analyzed for trace elements (Martin and Knauer 1973;
Martin et al. 1976), hydrocarbons (Conover 1971), etc., under the assumption
that much of this material will sink from the surface mixed layer and even-
tually transport associated elements, compounds, etc. to mid-depths and to the
bottom. Thus, considerable data exist on the plankton in the surface waters
and their remains on the sea floor.
In addition, the types of material carried to various depths by particles
are becoming well documented through sinking rate studies; considering fecal
pellets alone, for example, there are convincing data with respect to trace
metals (Lowman et al_. 1971; Boothe and Knauer 1972; Small et al_. 1973; Small
and Fowler 1973; Benayoun et aK 1974; Bishop et aj. 1977; Fowler 1977;
Spencer et al. 1978); skeletal structures and frustules (Schrader 1971;
Ferrante and Parker 1977; Roth et al. 1975; Honjo 1976); alpha emitters and
transuranic elements (Cherry et al. 1975; Bacon et al. 1976; Beasley et al.
1978; Higgo et al_. 1977); chlorinated hydrocarbons (Elder and Fowler 1977);
petroleum hydrocarbons (Conover 1971).
However, there are few data for the most important measurement needed,
i.e., the actual flux of detritus and associated materials as they leave the
surface and sink through the water column. The lack of this type of informa-
tion has hindered our ability to deal with specific problems such as removal
of inorganic and organic pollutants from the sea surface, understanding
nutritional aspects of the mid-water column and deep-sea benthos, determining
the fate of "excess" industrially produced C02 and even unraveling the basic
biogeochemistry of the world ocean.
145
-------
During the past two or three years, the means have become available to
assess fluxes of materials, not only to the sea floor but also through various
portions of the water column (e.g., Wiebe et a/h 1976; Soutar et aJL 1977;
Bishop et aJL 1977; Knauer et ah 1979). Recent advances in methodology and
instrumentation have also made possible the accurate measurement of elements
and compounds at the very low levels at which they exist in sea water (e.g.,
Boyle and Edmond 1975; Moore and Burton 1976; Boyle et aj. 1977; Bruland et
al_. 1978ab; Bruland et al. 1979). The combination of developments now enables
oceanographers to measure amounts added or removed (rates of change) by
comparing fluxes at various depth intervals. In other words, we can now
study the processes and measure the rates that determine the distribution of
elements and/or compounds in the water column.
However, because of the complexity of the vertical transport problem in
terms of interacting oceanographic variables (e.g., rates of primary produc-
tivity, particulate vs. dissolved components, season, sample depth, and
particle interceptor trap design, etc.), we do not feel that isolation of one
single topic for discussion (e.g., fluxes of trace elements from surface to
deep ocean waters) would present the clearest picture in a dynamic sense.
Therefore, we would like to present various aspects of the vertical flux
problem that we are actively working on at the present time. This will be
done through the use of tables and figures which will be used to illustrate
various points. The tables and figures are organized according to related
content and will be preceded by a topic phrase or sentence followed by a brief
discussion of their relevance.
We have included no major discussion in the text at this time, since many
of the data here have been recently obtained and many samples are yet to be
analyzed. However, even now the data suggest some exciting results and
trends, and we are looking forward to feedback and general discussion from our
Soviet counterparts engaged in similar research.
METHODS
Many of the methods pertaining to MULTIPIT design and sample processing
can be found in Knauer et a|. 1979 (Fluxes of particulate carbon, nitrogen,
and phosphorus in the upper water column of the Northeast Pacific, Deep Sea
Research 26(1 A):97-108).
In terms of metal analysis, procedures involving clean techniques such as
the use of portable ship-going laboratories, clean-up, etc. can be found in
Bruland et al_. 1979 (Sampling and analytical methods for the determination of
copper, cadmium, zinc, and nickel at the nanogram per liter level in sea
water, Anal. Chem. Acta, 105:233-245).
Most of the data presented here were obtained using a free-floating
MULTIPIT system set approximately 60 km off the Central California coast for 6
days (Figure 1). The MULTIPITs were placed at 35, 65, 150, 500, 750, and
1,500 meters. The vessel used was the R/V WECOMA, December 1978.
146
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PIT RECOVERED
1000 DEC 16
7o330 DEC 16
37°-
PIT
LAUNCHED
0500
DEC 10,1978
DEPTH CONTOURS
IN METERS
8
I22°4O' CM
I V
'..-. .36°20-
Figure 1.
-------
RESULTS
I. Mechanics
Figure 2. MULTIPIT collector assembly. The cross is of polyvinyl
chloride consisting of 8-12 individual acrylic tubes. The tubes are filled
with a density gradient (e = 1.07 g/cm-3) to insure retention of the
particles. We use 5% formalin as a preservative.
Table 1. There are zooplankton in the area of each MULTIPIT as well as
vertically migrating zooplankton populations. We have found that some of
these organisms actively swim into the MULTIPITs and die upon encountering the
gradient-formal in solution. These must be removed by hand (dissecting scope)
or considerable contamination can result. Table 1 illustrates this point.
For example, in the 1,500 meter MULTIPIT, removed swimmers account for 4,893,
624, and 1.16 ug of C, N, and Zn, respectively. This would contribute 310,
330, and 30 percent more of these elements respectively to the MULTIPIT
particulates if not removed.
TABLE 1. SWIMMER EFFECTS: POTENTIAL CONTRIBUTION OF REMOVED SWIMMERS TO
TOTAL CARBON, NITROGEN, AND ZINC IN MULTIPITS (from Knauer and
Martin)
MULTIPIT
(m)
35
65
150
500
750
1500
Removed
Swimmers (ug)
C
5485
5257
5370
1059
1546
4893
N
1123
939
1067
230
245
624
Zn
1.21
0.97
0.78
0.70
1.16
MULTIPIT
Parti culates (ug)
C
7328
3384
1819
493
473
1578
N
1474
525
258
67
54
189
Zn
7.28
5.96
3.81
1.87
1.64
3.82
Swimmers
Parti culates
C
75
155
295
215
327
310
N
76
120
414
340
455
330
( 100
Zn
17
16
20
37
--
30
Table 2. It is reasonable to assume that particles sinking through the
water column will lose some of their associated contents during their descent.
This is illustrated in Table 2. For example, at the end of two weeks (some of
our sampling periods are this long), 77, 89, and 97% of the initial carbon,
nitrogen, and phosphorus respectively were lost under the "decomposition
treatment," while with formalin 32, 40, and 58% were lost. Thus, it is
important to retain these "dissolved" compounds, and this is accomplished
using the density gradient which is also analyzed. For example, to get total
carbon, you must analyze the carbon in both the "particulate" and "dissolved"
fractions.
148
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TOP VIEW CROSS
122 CM
POLYPROPYLENE
HYDROLINE
HANDHOLD
STABILIZING
LANYARD
7.5 CM \
LINE CLAMP.
15 CM
RETAINING
'CORD 76 CM
HIGH IMPACTPVC
53 CM
REMOVEABLE
BAFFLE SYSTEM
35 CM
TOP VIEW
BAFFLE
GRID SYSTEM
MULTI-REPLICATE COLLECTOR
27CM
RETAINING
^COLLAR
COLLECTION
CUP
SINGLE COLLECTOR
Figure 2.
149
-------
en
o
TABLE 2. DECOMPOSITION: MIXED PLANKTON TOW COLLECTED AND PLACED IN METAL STRIPPED SEA WATER AT 35%
SALT. DECOMPOSITION ALLOWED TO PROCEED FOR TWO WEEKS AT ROOM TEMPERATURE (from Knauer and
Martin)
Treatment
Initial
1
2
3
4
X
Decomposition
1
2
3
4
X
5% Formalin
1
2
3
4
X
Azide (10 g/1)
1
2
3
>. 4
X
Initial
Wet Wt.
(g)
1.57
1.78
1.64
1.70
(1.67)
1.62
1.70
1.67
1.68
(1.67)
1.48
1.70
1.65
1.63
(1.62)
1.52
1.66
1.67
1.78
(1.66)
Final
Wet Wt. % Wt.
(g) Loss
0.55
0.56
0.57
0.61
(0.57) (66)
1.08
1.10
1.14
1.09
(1.10) (32)
0.67
0.75
0.66
0.83
(0.73) (56)
Total CNP (mg) % Loss of Initial
C
51
57
57
37
(51)
11
11
11
13
(11.5)
34
33
40
31
(34.5)
14
18
14
19
(16)
N
13
14
14
9
(13)
1.3
1.3
1.4
1.6
(1.4)
8
7
9
7
(7.8)
3.0
5.2
3.8
5.3
(4.3)
P C N
1.1
1.1
0.9
0.9
(1.0)
0.04
0.03
0.03
0.03
(0.03) (77) (89)
0.43
0.41
0.43
0.42
(0.42) (32) (40)
0.07
0.10
0.07
0.10
(0.09) (69) (67)
P
(97)
(58)
(91)
-------
Table 3.
field samples
17% of the Cd
gradient. Copper,
fraction.
This table
for cadmi urn
is retained
on the
illustrates the above effect with actual MULTIPIT
, copper, manganese, and P04-3. For example, only
on the particles, while 83% is lost to the density
other hand, is strongly bound to the particulate
TABLE 3. PARTICULATE/DISSOLVED: SEPARATE ANALYSIS OF MULTIPIT PARTICULATE
AND DISSOLVED FRACTIONS (from Knauer and Martin)
MULTIPIT
Depth (m)
35
65
150
500
750
1,500
Cd
Part.
43
25
11
4
3
5
(ng)
Di
ss.
248
1
39
60
19
14
18
Cu
Part.
274
324
314
261
230
410
(ng)
Diss.
<22
<22
<22
<22
<22
<22
Mn
Part.
3
5
6
3
3
4
,760
,377
,020
,926
,285
,855
(ng)
Di
1,
1,
ss.
633
329
743
192
232
214
P04
Part.
124
51
26
7.3
5.5
12.7
(M9)
Diss.
196
82
54
10.5
12.2
33
II. Fluxes of Carbon, Nitrogen, and Phosphorus
One of the major questions we are concerned with is: Of the total
organic material produced in the euphotic zone by the phytoplankton (primary
productivity), what fraction reaches various depths in the water column? This
is important in terms of biogeochemical cycles as well as pollutant transfer.
Data pertaining to this question are presented in Tables 4 and 5.
Table 4. These data represent the total amount of carbon fixed through-
out the euphotic zone over the duration of the MULTIPIT set (i.e., 6 days).
The method used for primary productivity measurements was a new metal-free
technique developed by us which can be discussed at the Symposium. It is
obvious from these data that most of the carbon fixed is recycled in the upper
150 meters. For example, of the total carbon fixed (i.e., 4,105 mg C m-2/6
days), 956 mg m-2 or 23% reached 65 meters (= the 1% light level or "bottom"
.of the euphotic zone). This supports earlier assumptions that most of the
material produced in the upper layers of the ocean is regenerated in the upper
levels of the ocean (Votinez 1953; Menzel 1974). Note that the percentage
appears to increase at 1,500 meters. We think this is an example of the
ladder effect suggested by Vinogradov (1961) and can be discussed in more
detail at the Symposium.
Table 5. .This table presents the amount of C and N collected in each
MULTIPIT over the 6-day set. Notice the rapid decrease culminating at the
oxygen minimum zone (750 m). Also notice the increase of C and N at 1,500 m
(ladder effect?). In terms of C/N ratios, the obtained values seem reason-
able. The values 6.0 and 6.9 found over the euphotic zone (0-65 m) were not
unexpected since most of the C and N is ultimately plankton-derived and should
yield values close to the Redfield et al. ratio of 6.6 (Redfield et al. 1963).
151
-------
TABLE 4. CARBON LEAVING THE EUPHOTIC ZONE AS PERCENT OF CARBON (PRIMARY
PRODUCTIVITY) FIXED OVER THE 6-DAY MULTIPIT SET (R/V WECOMA,
December 1978; from Knauer and Martin)
Date
1978
10 December
11 December
12 December
13 December
14 December
15 December
Integrated Primary
Productivity
mg C m-2 day-1
780
810
695
690
650
480
I 4,105
(mg C m-2/6 days)
Total Average
Carbon Collected
Depth
(m)
65
150
500
700
1,500
in MULTIPIT
mg m-2
956
530
107
97
290
As % of Primary
Productivity
23
13
2.6
2.3
7.1
TABLE 5. PARTICULATE CARBON AND NITROGEN AND C/N RATIOS / MULTIPIT (R/V
WECOMA, December 1978; from Knauer and Martin)
Depth
On)
35
65
150
500
750
1,500
Carbon
(mg)
7,880
6,775
3,745
3,152
2,080
1,500
420
580
380
575
1,145
2,475
X
7,330
3,450
1,790
500
480
1,810
Nitrogen
(M9)
1,590
1,270
600
570
315
230
58
86
49
65
130
316
X
1,430
586
270
72
57
223
C/N
(by Atoms)
5.8
6.2
7.3
6.5
7.7
7.6
8.5
7.8
9.1
10.3
10.3
9.1
X
6.0
6.9
7.7
8.2
9.7
9.7
106C:16N from Redfield et al. (1963) = 6.6
152
-------
Increases in this ratio with depth are also plausible, since the more nutri-
tious nitrogenous substances (N = protein) are being preferentially used while
the more refractory carbon is left behind (Knauer and Ayers 1977).
III. The Rain of Detritus
'rain1
In 1888, Agassiz proposed that "... deep-sea organisms are nourished by a
of organic detritus from overlying surface waters." We are also exam-
ining this question with our MULTIPITs. The data presented in this section
reflect this.
Figure 3. Shows the decrease in material collected in sextuplicate.
Again, it would seem that most of the material produced at the surface is
regenerated above the oxygen minimum. However, there appears to be an
increase in total weight collected at 1,500 m. That this is not an artifact
is reinforced in the data presented in Table 6.
Table 6. We have been examining and analyzing the MULTIPIT material for
exoskeletons, chitin, fecal pellets, etc. These data represent the major size
classes of fecal pellets found in our collectors over the 1,500 meter sampling
depth. Notice the general decrease in the cylindrical size class. However,
the 0.05 and 0.10 mm size class increases at 1,500 meters. We feel this may
be an example of repackaging.
TABLE 6. FECAL PELLET DATA, NORTHEAST PACIFIC (WECOMA, December 1978; from
Urrere, Knauer, and Martin 1979, in preparation)
Size Class (mm)
Depth
(m)
35
65
150
500
750
1,500
Elliptical
0.05
124
181
132
159
313
219
154
175
540
481
790
864
0.10
341
401
116
133
80
105
140
131
388
364
759
767
0.15
230
269
125
109
8
16
7
16
10
10
77
63
0.25
11
11
7
5
1
4
4
2
7
4
3
5
cyi
0.05
5,013
4,871
4,188
3,831
3,737
3,127
299
363
193
229
135
198
indrical
0.10
902
1,259
733
847
636
724
123
141
37
33
42
64
0.15
29
35
20
36
28
30
7
8
13
3
10
13
Round
725
674
472
395
359
289
78
70
58
41
76
63
Coiled
281
251
119
107
62
54
7
4
3
3
0
0
Total Flux
m-2 day-1
322
238
215
35
49
84
,660
,440
,990
,722
,966
,614
153
-------
0
100
300
500
- 700
x
i-
0.
900
MOO
1300
1500
mg DRY WT / COLLECTOR
10 20 30
40
Figure 3.
154
-------
Figure 4. This is a plot of In fecal pellet number (0.10 mm) with depth.
There appears to be a good fit to an exponential curve to 750 meters, although
this size class at the 1,500 meter depth falls off the line. This may be the
result of the ladder effect (see below, Table 7).
Table 7. Analysis of the 0.10 mm size class reveals that the primary
in these pellets were Pseudoeunotia doliolus and Coscinodiscus
In general, all pellets from this size class contained the same
these cells except for the pellets analyzed at 750 m. This
some of the organisms from 1,500 meters are migrating up through
zone to feed.
food items
fragments.
quantity of
suggests that
the oxygen minimum
TABLE 7. DOMINANT FOOD ITEMS IN THE 0.10 mm DIAMETER CYLINDRICAL SIZE CLASS
(from Knauer and Martin)
X Number/Depth (m)
Food Item
35
65
150
500
750
1,500
Pseudoeunotia doliolus 13
Coscinodiscus fragments 14
Prorocentrum sp. 2
Nitzschia-navicula (?) 2
Nitzschia sp. 1
Dinoflagellate parts 1.!
Zooplankton parts 1
(n pellets analyzed) (14)
9
10
1
0
0.1
0.6
0.8
(7)
1
17
8
0
0
0
0.2
0.2
(8)
10
7
1
0
0
0
0.2
(7)
1
2
2
0
1
0
0
0.2
(6)
1
12
9
0
2
0.4
0
0.3
(7)
% Dominant Phytoplankton Species (Partial List)
Schroderella delicatula
Bacteriastrum sp.
Nitzschia sp.
Skeletonema costatum
Chaetocerus af finis
19
17
10
12
7
Pseudoeunotia doliolus
Coscinodicus (9 species)
Ceratium sp.
Prorocentrum sp.
1.2
0.4
0.02
0.15
IV. Metal Fluxes
We are presently analyzing our MULTIPIT samples for Al, Ag, Cd, Cr, Cu,
Fe, Mn, Ni, Pb, and Zn. However, we are only in a position to discuss a few
of these elements at this time. Three examples (Mn, Pb, and Cd) are presented
below.
Figure 5. Shows oxygen distribution (ml 02/1) and dissolved Mn (ng/1).
[Dissolved Mn is defined here as Mn remaining after passing a seawater sample
(10-30 1) through acid washed 0.4 u nucleopore filters.] It can be seen that
there is a rapid decrease in dissolved Mn from surface waters (100-200 ng/1)
to depth (40-60 ng/1). It is interesting to note that there is a slight
maximum in the oxygen minimum zone which may suggest a redissclution of Mn02.
155
-------
en
CTl
8
Q
OJ
EG
UJ
d
Q.
c
0
0.10 mm DIAMETER SIZE CLASS
CYLINDRICAL FECAL PELLETS
400
800
DEPTH (m)
1200
1600
Figure 4.
-------
0
0
20
ng Mn/ liter
40 60 80 100
200
400
800
h-
CL
LU
Q
1200
1600
2000
2400
/02
- \
\
\
\
\
\
\
0
o Sta63 DEC 1976
• Sta 64 APR 1977
• Sta 65 JUL 1978
a Sta 66 DEC 1978
mliter 02/ liter
246
8
Figure 5.
157
-------
Figure 6. Illustrates the distribution of three Mn fractions found in
the MULTIPITs with depth: the particulate, the residual particulate digested
in HF (inorganic), and the soluble. In the case of Mn, most of this element
resides on the particulate phase. Comparison of the dissolved Mn (Figure 5)
with the total MULTIPIT Mn (this figure) suggests a loss of Mn from the latter
phase, while the dissolved phase increases in the oxygen minimum zone (750 m).
Table 8.
total Mn
At this time, we are unable to distinguish what fraction of the
flux is a result of biological activity or inorganic input (e.g.,
runoff, resuspended sediment, etc.). For this reason, Mn fluxes are reported
for "Total" (all phases of Mn/MULTIPIT), "Mud Corrected" (i.e., hand waving
using Al/Mn ratios derived from marine sediment analysis which can be used to
theoretically correct for sediment inclusions—probably not too accurate) and
"Soluble Mn" (the amount of Mn found dissolved in the density gradient).
"Soluble" Mn fluxes were used to compute residence times since we assumed that
this fraction most probably represents the biological contribution. In this
table, "Total Mn" represents the total dissolved Mn in a 35-meter water column
(e.g., Average Mn concentration in 35-meter water column = 157 ug/m3 x 55 =
5,500).
TABLE 8
Depth
(m)
35
65
150
500
750
1,500
Total
Mn Flux
|jg/m2/day
240
330
340
190
160
250
Mud Corr.
Flux
ug/m2/day
95
70
67
68
76
0
Sol. Mn
Flux
ug/m2/day
69
55
31
8.0
9.7
8.9
(•^9.0)
Depth
Interval
(m)
0-35
0-65
0-150
0-500
0-750
0-1,500
0-3,500
Total Mn
ug/m2
5,000
11,000
21,000
40,000
58,000
110,000
210,000
(Sol.)
Res. Time
(Years)
0.22
0.55
1.9
14
16
34
•v-64
Table 9. From the data presented above, we have calculated "Soluble" and
"Mud Corrected" Mn fluxes to compare with Bender et aj. (1970) Mn fluxes
needed for "excess Mn in open-ocean sediments."
Table 10. This table presents total ng Pb/individual MULTIPIT (i.e.,
particulate + dissolved), Pb concentration in this fraction (ug/g dry wt), Pb
fluxes, and rates of change. As with Mn, total Pb appears to increase rapidly
with depth to the area of the oxygen minimum zone followed by a marked
decrease. It is interesting to note that except for the 35-meter collection
depth, Pb concentrations associated with the total particulate flux remain
essentially constant. This appears to be at odds with the literature in terms
of Pb210 activity, since there is evidence that no regeneration of this
nuclide occurs in the water column (Bacon et al. 1976). If this is so, then
as particle mass decreases (as it does here—see Figure 3) and if Pb is
retained on the particles (i.e., not regenerated), then the concentration
158
-------
FRACTIONAL ng Mn/ COLLECTOR
1000 2000 3000 4000 5000 6000
O
2 6
x
E
Q_
LJ
Q
8
10
12
14
1000 3000 5000 7000
TOTAL ng Mn
9000
Figure 6.
159
-------
TABLE 9. MANGANESE FLUXES (from Knauer and Martin)
pg Mn/cm2/!,000 Years
Flux Needed for Excess Mn in Open-Ocean Sediments
(Bender et afL 1970)
Soluble Mn Flux at 1,500 m
x Mud Corrected Flux (35, 65, 150, 750 m)
Total Uncorrected Flux at 1,500 Meters
800
320
2,600
9,100
TABLE 10. TOTAL LEAD (PARTICULATE AND DISSOLVED) COLLECTED / MULTIPIT,
CONCENTRATION (SALT REMOVED). FLUXES AND RATES OF CHANGE (from
Knauer and Martin)
Depth
(m)
35
65
150
500
750
1,500
Total P
(M9)
529
503
526
520
500
643
606
573
261
401
339
240
322
168
248
264
420
Concentration Fluxes
(ug g-1) ug m-2 day-1
19.0 21
17.9
39.8
26.1 21
27.0
35.9
36.9 25
36.9
32.2
31.0 14
38.9
48.0
35.4 10
24.7
20.3
39.4 13
32.3
Ranges of Change
ng I-1 yr-1
(35-65)
(65-150)
(150-500)
(500-750)
(750-1,500)
0
16
12
5.5
1.4
should increase with depth. Certainly, the common Pb isotope should not be
expected to act differently from Pb210. We expect to have Pb210 data for the
MULTIPIT set in the future. It should be noted that in Knauer et al_. (1979),
Pb210 activity did increase with depth to 700 m during the coastal upwelling
episode.
160
-------
0
ngCd/LITER (DISSOLVED)
25 50 75 100
125
0
50
100
150
^ 200
300
Q_
LJ 500
0
700
900
1100
1300
1500
0
ngCd/COLLECTOR
100 200 300
A
AM
At C
* j
1
j
'
yip O
'
j
'
•KID
n
i
i
4-
1
1
1
1
**^>x* — — — ••"
^ O Q4| *^<\O
/ \
/ \
/ \
/ \
•my oa • \
7 \
/ \
f' \
\ ^
\
> "
\
\*-
\
\
L
1
1
1
• (DISSOLVED) 1
0 MULT/ PIT SOLUBLE i
A PARTICULATE ,
a TOTAL (PART+SOLUBLE) I
r
i
, J
Figure 7.
161
-------
15
ng Cd = 32(jjM PO*})- 0.43 (Dissolved + Particulate)
ng Cd = 34.9 (jjM P04) -3.6 (Sea water)
ro
10
ro
g
X
o
o •
0
0
100
200 300
P04/g DRY WT.
400
Figure 8.
-------
Figure 7. This graph compares total Cd (particulate + soluble) with
"dissolved" Cd in sea water. (Again, "dissolved" Cd is defined here as that
fraction left in solution after passage through a 0.4 n acid washed nucleopore
filter). The two phases appear to be related inversely, as might be expected.
For example, as total Cd (from MULTIPITs) decreases with depth (squares),
"dissolved" Cd (solid circles) increased with depth. This relationship is
reinforced in Figure 8.
Figure 8. This is a regression of total Cd (particulate + soluble)
against total P04-3 (particulate + solublej from the MULTIPIT collections.
The equation of the line [ng Cd = 32(uM P04-) - 0.43] is very similar to that
published by Bruland et a\_. (1979) for the relationship between these two
elements "dissolved" in sea water, which tends to support the relative
accuracy of the system.
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CORRELATION BETWEEN DOSE LOADS IN FISH AND THE BIOGEOCHEMISTRY
OF ARTIFICIAL RADIONUCLIDES IN A MARINE ENVIRONMENT
by
I. A. Shekhavona and A. P. Panarin
All-Union Scientific Research Institute
of Marine Fisheries and Oceanography
VNIRO, Moscow
The entire process of establishing life on Earth has taken place under
the effects of ionizing radiation, caused by cosmic radiation and natural
radionuclides from the planets. The relative constancy and stability of
natural radiation has helped contemporary animals and plants adapt to this
factor; however, we can assume that this type of adaptation is limited.
Scientific-technological progress, the development of the atomic energy
industry, and the wide use of nuclear energy since the middle of the current
century have caused the appearance of new, anthropogenic sources of ionizing
radiation. In contrast with natural radiation, the concentration of arti-
ficial radionuclides is increasing in some areas of the biosphere, and the
intensity of irradiation of living organisms is also changing as a conse-
quence.
Like any other factor, after reaching a certain intensity which exceeds
the limits of the tolerance developed in the course of evolution, ionizing
radiation can become a limiting factor and can have a negative effect on the
most radiosensitive organisms. This is also true of aquatic biocenoses, since
almost all radioactive substances of an artificial origin which enter the
atmosphere finally become concentrated in the world's oceans.
The ban on nuclear tests in three environments as a result of the Moscow
pact of 1963 sharply decreased the intensity of global radioactive fallout,
but did not totally stop artificial radionuclides from entering the hydro-
sphere. This was related to a number of causes. First, the existence of
stratospheric reserves of long-lived products of nuclear explosions which,
based on many estimates, will fall out until the year 2000 and the continuous
migration of accumulated reserves of these products from land into the seas
and oceans with drainage from rivers. Second, the continuation of nuclear
arms tests in some countries (China, France). Third, the practice of dumping
radioactive wastes in the oceans, seas, rivers, and inland bodies of water in
some countries (U.S., England, Japan, Belgium, etc.) (Nelepo 1970; Patin 1970;
Gromov and Spitsyn 1975).
166
-------
An extensive program has recently been established for constructing
nuclear power plants (NPP) on the coasts of seas and oceans, and the various
aspects of building floating NPP on special platforms in coastal waters around
large cities has been discussed (Gusev 1975). Projected annual production of
atomic energy will increase to 2-1012 W (el.) by the year 2000, and the number
of operating NPP will increase to 5,000 versus 187 in operation in 1976
(Anonymous 1977). In these circumstances, it is logical also to expect a
corresponding increase in the amount of radioactive waste. This creates the
basis for local pollution of individual bodies of water by radionuclides.
However, as a result of the large-scale circulation and integral biological
structure of the world's oceans, regional anomalies in any one part can be
reflected in the radiation situation of neighboring regions and the entire
system as a whole. For this reason, any type of pollution in surface waters
has become one of the most acute international problems and requires rigorous
regulation.
There currently are no unified common standards for admissible concentra-
tions of artificial radionuclides in surface waters based on sanitary-hygienic
and ecological aspects in our country and abroad.
Based on current sanitary-hygienic standards for regulating the concen-
tration of radionuclides in water, only criteria for estimating the signif-
icance of contamination of drinking water have usually been established and
the basic biological chains for migration of radionuclides from the water into
the human body have been taken into consideration. Such an approach evidently
meets all requirements for radiation-hygienic safety, but does not include the
effect of ionizing radiation and its consequences on the inhabitants of an
aquatic environment, for example, fishes. The urgency of the problem of
ecological, piscicultural standards for artificial radionuclides in open
bodies of water has been emphasized repeatedly (Polikarpov 1964; Egami 1973;
Anonymous 1976), since fish breeding more significantly suffers from deterior-
ation of the quality of the aquatic environment. For this reason, it is of
particular interest in solving the scientific and practical problems related
to protecting bodies of water from pollution.
This position is probably complicated by the long-standing opinion that
if hygienic-sanitary standards are satisfactory for man, then they are also
suitable for fish. This is probably true in conditions of equivalent dose
loads on humans and fish with some limit to the concentration of radionuclides
in the water. However, in reality, the radiation loads on fish, which
constantly live in water, and man, who is exposed for limited periods of time.
differ. This is primarily related to the specific properties of the aquatic
environment in which intensive concentrations of radionuclides up to exceed-
ingly high levels are observed not only in fish, but also in important sources
of external irradiation of fish—bottom sediment and algae—parallel to
dilution and dispersion of radionuclides.
The behavior of radioactive elements in a body of water is much more
complex than in the atmosphere. Their migration and concentration are a
function of both the physical-chemical and biogeochemical properties of the
aquatic environment which contains a large amount of dissolved and colloidal
167
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organic substances, hydrobionts and suspensions which differ in composition
and origin. In addition, the rate of radionuclides in a body of water is also
a function of the properties of the radionuclides themselves.
Most of the artificial radionuclides which enter an aqueous environment
are isotopes of stable macro- or microelements which form a structural base
for living organisms or are included in the composition of biologically active
compounds which participate in all vital processes. Because of this, they are
actively exchanged, assimilated, and transported by hydrobionts and their
detritus, thus causing biocirculation of radioactive substances in the water.
Plankton organisms whose biomass is most significant in the world's
oceans play a particularly large role in the biogenic transport of radio-
nuclides. These inhabitants of an aquatic environment very effectively remove
radioactive elements from the water due to their large adsorbing surface.
During migration, plankton organisms transport the radioisotopes which they
have accumulated from the surface to the lower layers of a body of water;
after they die, they sink to the bottom and become concentrated in the
sediment. Biogenous sediment of radionuclides in the composition of plankton
residue, characterized by continuous and frequently changing generations, can
result in a marked increase in their concentration in the upper layer of
bottom deposits with all of the consequences for the living inhabitants of the
bottom area of a body of water.
Fish also participate in the biological migration of radionuclides, but
their role is insignificant. They disperse radionuclides in feeding and
digestion. At the same time, in contrast with land animals, fish can accum-
ulate and disperse dissolved radionuclides not only along the food chain, but
also directly from the water. The nature and rhythm of the uptake varies
significantly as a function of abiotic and biotic factors, the species, age,
and the physiological condition of the fish (Shekhanova 1978). For example,
it has been found that an increase in the temperature of the water involved a
proportional increase in the rate and even the degree of accumulation of
radionuclides by fish (Kulikov et al_. 1978; Katkov et aJL 1978). Such factors
should apparently be considered in using heated water from NPP in breeding
fish.
Particularly intensive uptake of radioactive isotopes by fish is observed
when there is a shortage of some necessary components in a region and they are
present in the surrounding environment. Both of these processes—feeding and
digestion on the one hand, and interaction with the elements in the surround-
ing environment on the other—do not exclude but only supplement each other.
As a result, fish accumulate artificial radionuclides in amounts which are
significantly higher than their concentration in the water. Due to these,
weighable dose loads are formed in their tissues and organs even when the
concentrations of radionuclides are low. We emphasize that almost all arti-
ficial radionuclides are selectively accumulated and are often localized in
tissues which are not used for standardization in sanitary-hygienic evalua-
tions. However, the same tissues and organs can be critical (most radiosensi-
tive) for the fish themselves.
168
-------
In addition to plankton, the suspended matter in the world's oceans
includes a significant amount of mineral particles of terrigenous and volcanic
origin; the specific surface of the suspended matter is significantly greater
than in living organisms. This is the reason for its excessively high
adsorbing capacity. It is believed that migration of such microelements as
Zn, Pb, Bi, Cu, Hg, Ag, and Mo (Kranskopf 1956) into the depths of the oceans
is basically due to sorption of suspended matter by mineral components; this
is apparently also true of the radioactive isotopes of these elements. Abio-
genetic suspended matter together with adsorbed radionuclides is gradually
transported to the lower layers of the ocean and deposited on the bottom.
This process takes place with particular intensity in coastal zones where
formation of sediment occurs much more rapidly than in pelagic areas. We also
note that solid particles and colloids are very effectively assimilated by
living organisms and are thus part of the biosedimentation system.
As a result, both sedimentation and suspended matter of mineral origin
and biosedimentation affect the enrichment of the upper layer of bottom
deposits with radioactive nuclides. The bottom deposits of oceans and seas in
turn are good natural sorbents with inherent ion exchange, chemosorption,
physical (or molecular) sorption, etc. For this reason, radionuclides are
intensively and sometimes irreversibly trapped by these deposits.
Bottom sediments consisting of argillaceous minerals, sludge character-
ized by high dispersity and sorptive capacity, are distinguished by the
strongest absorptive capacity (Gromov and Spitsyn 1975; Ryndina 1970; Carroll
1959). In addition, benthos organisms and attachment of algae which intens-
ively assimilate them in life and convert them to soil in dying also promote
the process of radionuclide accumulation in the upper layer of bottom
sediments (Ketchum 1960; Schafer 1960). The accumulation coefficients for
radionuclides in such soils can reach hundreds and thousands. As a result,
demersal eggs and bottom fish will be exposed to very appreciable dose loads
due to gamma and even beta stratification. Aquatic plants, which accumulate
radionuclides up to levels which significantly exceed their concentration in
the water, cause an additional dose load for phytophilic eggs.
The radioactive substances which enter bodies of water are thus very
rapidly involved in the hydrological, physical-chemical, and biogeochemical
processes occurring in the water. In the final analysis, this results in the
distribution and redistribution of radionuclides among the abiotic and biotic
components of the aquatic environment. The role of these factors in the
distribution and concentration of the individual radionuclides differs and is
primarily a function of their individual properties. Sedimentation and
biological factors apparently have no significant effect on the distribution
of 90Sr and 137Cs in the ocean, since these radionuclides are primarily
present in a marine environment in a dissolved state and are almost unrelated
to inorganic suspensions and hydrobionts. At the same time, for 144Ce, 95Zr,
95Nb, 106Ru-106Ru and particularly the isotopes of biogenous elements (55Fe,
65Zn, 54Mn, 60Co and others), the role of sedimentation and biogeochemical
processes in the migration and spatial distribution can be comparable to the
effects of hydrological factors (Svedov and Patin 1968).
169
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Some authors (Timofeyeva-Resovskaya and Timofeyeva-Resovskaya 1960)
believe that artificial radionucTides should be divided into four groups as a
function of their behavior and distribution in a body of water: (1) hydro-
tropic, persisting in high concentrations in the water; (2) biotropic, inten-
sively adsorbed by hydrobionts; (3) subtropic, accumulating primarily in
bottom sediment; and (4) eurytropic, equally distributed among the individual
components of the body of water.
This type of division should be useful in determining the dose loads in
fish, since they are not only determined by the value of the internal irradia-
tion from incorporated radionuclides, but also by a series of external sources
of ionizing radiation which are active in the body of water--the water, water
flora, the soil. In turn, the individual role of each of these sources of
external irradiation in forming the integral dose load in fish will be deter-
mined by the ecology, more precisely the biotope, in which the fish live. For
example, all other conditions being even, the irradiation to which pelagic and
phytophilic (or bottom) eggs are exposed is far from equivalent.
As a consequence, in developing standards for monitoring radioactive
contamination of bodies of water with respect to ecology and fish breeding, it
is insufficient to limit the concentration of radionuclides in the water
alone, but their concentration in all of the components of the body of water
and particularly their concentration in aquatic plants and bottom sediments
should also be taken into consideration.
We evaluated the hygienic studies of the liminal concentrations proposed
by Gusev (1975) with respect to their admissibility for fish breeding to
illustrate a possible ecological approach to standardization of the concentra-
tion of artificial radionuclides in sea water. In solving this problem, we
analyzed the level of internal beta radiation of fish from the radionuclides
accumulated in their tissues, and the external gamma and beta radiation from
radionuclides accumulated in bottom deposits.
The doses were calculated according to the commonly used equations
(Aglintsev et aJL 1962; Khayn and Braunell 1958):
Pp = 2.13 • E - C, (1)
where PD is the tissue dose rate, rad/hr,
P
E is the mean energy of beta radiation per disintegration, Mev,
C is the concentration of radionuclides in the tissue, c/g,
170
-------
2n -C • K
(2)
where P is the gamma radiation dose rate on the surface of the bottom, r/hr,
o
C is the concentration of radionuclides in the bottom sediment, c/cc,
K is the gamma constant for the radionuclide, r-cmVg'ms,
jj is the linear attenuation coefficient for gamma quanta (broad beam)
in the emitter material, cm-1,
1 is the thickness of the irradiated layer, cm,
<)> is King's function.
Equation (2) is recommended for an extended plane source of finite thick-
ness. In our case, with a bottom thickness containing gamma- irradiating
radionuclides equal to 20 cm, <|>(ul) •* 0 and expression (2) with an error of +1
to +10% (as a function of the energy of the gamma irradiation) approaches the
expression for a semi -infinite 2n source:
2n • C - K
(3)
The dose of beta irradiation on the surface of the bottom was set as
I to 0.5D6 (Khayn and Braunell 195
in the interior of the bottom sediment.
equal to 0.5D0 (Khayn and Braunell 1958), where D0 is the corresponding dose
p P
The dose rate of gamma radiation in the bottom sediment is a function of
the level of accumulation and the nature of the distribution of the radio-
active substances in the groups which form the bottom, and this in turn is
determined by an entire series of conditions (Marey 1976; Ryndina 1970). To
simplify the calculations, we assumed that the radionuclides were evenly
distributed in the 20-cm surface layer of bottom sediment. This assumption
could apparently be correct (at least for some long-lived radionuclides) in
estuaries and coastal zones of oceans and seas (Patel et al. 1975). The fact
that the concentrations of 60Co and 137Cs at a depth of 0.5 m did not differ
from the concentrations on the surface in bottom deposits in the Columbia
River, where nuclear reactor wastes have been dumped for a long time (Haushild
et al. 1973), supports this assumption.
We used the upper limits of the concentrations of radionuclides in sea
water (Gusev 1975) and the limiting coefficients for accumulation of radio-
active and stable isotopes in fish tissues for calculating the mean doses
absorbed by the tissues of fish (Bakunov et al_. 1973; Gusev 1975; Patin and
Petrov 1973; Pertsov 1973; Rozhanskaya 1970JT Since we were interested in the
dose loads in the fish itself and not in its consumer, we did not divide the
171
-------
accumulation coefficients into bone and muscle tissue, but took the maximum
value for both types of tissue (Table 2). The expediency of this approach is
supported by the fact that the actual dose of some radionuclides e.g., in the
kidneys of fish, is significantly higher than the calculated dose based on our
experimental measurements using thermoluminescent dosimeters (Shekhanova
1976). The effect of irradiation from radionuclides assimilated by the spine,
which is directly adjacent to the kidneys, is present here.
On the other hand, it is not possible to prevent fish from entering the
massive accumulation zone and the commercial marine invertebrate and algae
business. As a consequence, the division of working limits of concentrations
according to the different industrial zones loses its meaning in real condi-
tions (see Table 1). Based on the above, it was fo^nd that the tissue dose
TABLE 1. WORKING LIMITS FOR CONCENTRATIONS OF RADIONUCLIDES IN SEA WATER
(c/liter) WITH PROLONGED CONTAMINATION OF REGIONS INVOLVING
DIFFERENT INDUSTRIAL APPLICATIONS (Gusev 1975)
Radionuclide
Chromium51
Manganese54
Iron55
Iron59
Cobalt57
Cobalt58
Cobalt60
Zinc65
Strontium89
Strontium90
Yttrium90
Zirconium95
Niobium95
Molybdenum99
Ruthenium103
Ruthenium106
Cadmium109
Cadmium1 15m
Cadmium115
Antimony124
Antimony125
Iodine131
Cesium134
Cesium137
Cerium141
Cerium144
Polonium210
Plutonium239
Fishing
Zone
7-10-10
1-10-9
4. 10-io
3-10-11
5-10-9
MO-9
4-10-10
3-10-11
MO-9
4-10-11
2-10-9
4.10-io
MO-9
2-10-9
MO-9
l.lO-io
4-10-8
5-10-9
7-10-9
2-10-10
7.10-10
l.lO-io
3-10-10
6-10-10
3-10-9
4.10-10
5-10-11
2-10-10
Mollusks
MO-8
6-10-11
4.10-io
3-10-11
MO-8
2-10-9
g.lO-io
5-10-11
6-10-9
2-10-10
5-10-9
MO-8
4.10-9
2-10-9
5-10-8
6-10-9
l.lO-io
MO-11
2-10-11
MO-9
5-10-9
l.lO-io
MO-9
MO-9
4.10-9
6-10-10
l.lO-io
MO-9
Production Zone
Crustaceans
MO-8
8-10-11
2-10-9
l.lO-io
3-10-9
y.lO-io
3.10-io
l.lO-io
6-10-9
2-10-10
MO-9
3-10-9
5-10-9
2-10-9
5-10-8
6-10-9
MO-8
1-10-9
2-10-9
2-10-9
7-10-9
2-10-10
2-10-9
3-10-9
4.10-9
6-10-10
l.lO-io
5-10-10
Nutritive
Algae
7-10-9
5.10-io
4.10-io
3-10-11
4-10-9
MO-9
4.10-10
5«10-10
MO-9
4.10-10
l.lO-io
3-10-10
3-10-10
MO-8
3-10-10
3-10-11
6.10-io
8-10-11
l.lO-io
5-10-10
2-10-9
MO-11
6-10-10
MO-9
6-10-10
l.lO-io
4.10-n
MO-11
Mixed
Industrial
Zone
7.10-10
MO-10
4.10-io
3-10-11
3-10-9
7.10-iO
3-10-10
3-10-11
MO-9
4.10-n
l.lO-io
3-10-10
3-10-10
2-10-9
3-10-10
3-10-11
MO-10
MO-11
2-10-11
2-10-11
7.10-iO
MO-11
3-10-10
6-10-10
6-10-10
VI O-io
4.10-11
MO-11
172
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TABLE 2. DOSE RATE OF BETA IRRADIATION OF FISH FROM INCORPORATED RADIO-
NUCLIDES*
Radionuclide
Upper Limits
for concen-
trations in
sea water,
c/liter
Accumulation
coefficients
in fish
tissues
Limits of
concentration
in fish tissue,
c/kg
Tissue dose
rate,
rad/year
Iron59
Cobalt58
Cobalt60
Strontium89
Strontium90
Yttrium90
Zirconium95
Niobium95
Molybdenum99
Ruthenium103
Ruthenium106
Cadmium115m
Cadmium115
Antimony124
Antimony125
Iodine131
Cesium134
Cesium137
Cerium141
Cerium144
1-10-10
2-10-9
9-10-10
6-10-9
2-10-10
5-10-9
MO-8
5-10-9
1-10-8
5-10-8
6-10-9
5-10-9
7-10-9
2-10-9
5-10-9
2-10-10
2-10-9
3-10-9
4-10-9
6-10-10
3,000
560
560
200X
200X
10
150
100
20
100
100
5
5
140
140
15
230
230
100
100
3.00-10-7
1.12-10-6
5.04-10-7
1.20-10-6
4.00-10-8
5.00-10-8
1.50-10-6
5.00-10-7
2.00-10-7
5.00-10-6
6.00-10-7
2.50-10-8
3.50-10-8
2.80-10-7
7.00-10-7
3.00-10-9
4.60-10-7
6.90-10-7
4.00-10-7
6.00-10-8
0.66
0.63
9.09
12.84
0.85
0.87
3.33
0.40
1.45
6.04
16.12
0.30
0.23
2.00
1.14
0.01
1.66
2.45
1.08
1.46
* Translator's note: Note to table illegible.
rate for the individual radionuclides reaches 16.12 rad/year with the recom-
mended upper working limits for their concentration in sea water (see Table
2). The highest dose loads are formed by such radionuclides as 60Co, 89Sr,
103 106Ru
For comparison, we note that 40K, which is usually
radioisotope which forms the beta activity in the body of
of flora and fauna (Pertsov 1973), creates a dose load in the tissues of fish
which does not exceed 2.8 • 10-2 rad/year (Polikarpov 1964).
the basic natural
any representative
Radionuclides contained in bottom deposits can be one of the most impor-
tant sources of radiation for hydrobionts. The particularly appreciable
effect of this source of ionizing radiation will be exercised not only on the
comparatively radioresistant benthos organisms, but also on the much more
sensitive bottom eggs, the young and adult specimens of marine fish which lead
to a benthopelagic type of life. This is not always related to gamma radia-
tion alone. The range of the effect of beta radiation from some radionuclides
173
-------
in water and the biosubstrate extends over distances comparable to and some-
times exceeding the dimensions of fish eggs, larvae, and a number of small
hydrobionts, e.g., more than 1.5 cm for 106Ru (E = 3.54 mev) (see Table 3).
(113.X
TABLE 3. LENGTH OF THE MAXIMUM RANGE OF BETA-PARTICLES IN WATER AND IN THE
BIOSUBSTRATE (Pertsov 1973)
Energy
mev
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
Water
(Biosubstrate)
mm
0.14
0.45
0.84
1.29
1.77
2.27
2.78
3.31
Energy
mev
0.9
1.0
1.2
1.4
1.6
2.0
3.0
5.0
Water
(Biosubstrate)
mm
4.10
4.80
5.47
6.56
7.76
9.84
15.3
25.8
In addition, as our direct measurements using thermoluminescent dosim-
eters showed, the dose field on the bottom of natural bodies of water contam-
inated with beta-radiating radionuclides exhibits no sharp difference at the
boundary of the water-soil section, and changes much more smoothly than
calculations indicate. We can assume that this is related both to the
presence of a peculiar "diffusion layer" with a high concentration of radio-
nuclides in the water itself, and to the presence of a highly active suspen-
sion within the water (soil particles, detritus, plankton organisms).
The latter could play a particularly significant role in the shallow
regions of the open ocean and the shelf zone where intensive entrapment of
soil particles from the bottom and intermixing of water masses take place
during swells and where most aquatic animals live. In turn, suspended
particles which are deposited on their skin could also be an additional source
of external irradiation.
In an ecological evaluation of the admissible levels of radioactive
contamination of an aquatic environment, it is thus also necessary to consider
the presence of external sources of irradiation of fish in this environment
and primarily the dose loads created by radionuclides concentrated in the
bottom sediments (see Table 4).
In extreme conditions, fish will receive both a maximum dose of external
gamma radiation from the bottom and a very high dose of internal beta radia-
tion from incorporated radionuclides. The total dose rate from both sources
of ionizing radiation will then be close to a value in the order of 20-30
rads/year as a function of the individual radionuclides. In these conditions,
the bottom eggs, larvae, and young of fish can additionally be exposed to
174
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TABLE 4. DOSE RATE OF GAMMA AND BETA RADIATION FROM RADIONUCLIDES CONTAINED
IN BOTTOM SEDIMENTS
Radionuclide
Upper limits
of Concen-
trations in
Sea Water, c/kg
Limiting
Coefficients
for accumu-
lation in
Sand
Upper Limits
of Contamin-
ation of
Sand, c/kg
Dose Rate,
rads/year
^
Chromium51
Manganese54
Iron59
Cobalt60
Zinc65
Zirconium95
Niobium95
Ruthenium103
Ruthenium106
Cadmium109
Antimony124
Cesium134
Cesium137
Cerium141
Cerium144
MO-8
MO-9
MO-10
g.-l O-io
5-10-10
MO-8
5-10-9
5-10-8
6-10-9
4-10-8
2-10-9
2-10-9
3-10-9
4.10-9
6-10-10
1,000
400
1,000
1,000
60
150
150
600
600
100
300
300
300
1,400
1,400
1.0-10-5
4.0-10-7
1.0-10-7
9.0-10-7
3.0-10-8
1.5-10-6
7.5-10-7
3.0-10-5
3.6-10-6
4.0-10-6
6.0-10-7
6.0-10-7
9.0-10-7
5.6-10-6
8.4-10-7
1.16
2.13
0.71
19.22
0.11
6.40
3.20
21.35
4.66
0.09
4.85
4.27
2.75
0.59
0.19
___
—
0.11
0.81
—
1.67
0.30
18.18
48.38
—
2.15
1.08
1.58
7.65
10.28
1.16
2.13
0.82
20.03
0.11
8.07
3.50
39.53
53.04
0.09
7.00
5.35
4.33
8.24
10.47
external beta radiation from bottom sediments which attains 48.38 rads/year
for 106Ru, for example, 53.04 rads/year (~ 145 mrads/day) together with the
gamma components.
It is known that other aquatic fauna and flora (Chipman 1972; Anonymous
1976) as organisms representing the highest degree of development with respect
to fish are the most radioresistant link in the set of hydrobionts. If we
take 10-15 mrads/day (approximately 5 rads/year) (Shekhanova 1975) as the
maximum admissible dose rate for irradiation of the gonads and kidneys of
fish, the upper limits of the concentrations of radionuclides in sea water
(Gusev 1975) will be too high from a radioecologic viewpoint not only due to
the total accumulation of external (gamma) and incorporated (beta) sources of
radiation, but also even in exposure to one of these sources, e.g., 60Co,
89Sr, 95Zr, io3.io6RUj i24Sb> i34Cs If we consider the total contribution of
these sources plus the external beta radiation from the bottom, then the
following must be added to this series of emitters: 137Cs, 141Ce, and 144Ce.
The admissible levels of irradiation for fish for the individual radio-
nuclides can be obtained with the concentrations in sea water shown in Table
5. These concentrations were calculated with the condition that each radio-
nuclide should not create a total dose rate of external and internal radiation
exceeding 5 rads/year and this apparently corresponds to the radioecological
requirements more precisely.
175
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TABLE 5. WORKING LIMITS FOR CONCENTRATIONS OF ARTIFICIAL RADIONUCLIDES IN SEA
WATER (c/liter)
Limits of Concentrations in Sea Water
Radionuclide
Chromium51
Manganese54
Iron59
Cobalt60
Zinc65
Stronti urn89
Strontium90
Zirconium95
Niobium95
Ruthenium103
Ruthenium106
Cadmium109
Antimony124
Cesium134
Cesium137
Cerium141
Cerium144
Hygienic
(Gusev 1975)
MO-8
MO-9
MO-10
9-10-10
5-10-10
6-10-9
2-iQ-io
1-10-8
5-10-9
5-10-8
6-10-9
4-10-8
2-10-9
2-10-9
3-10-9
4-10-9
6-10-10
Ecologic
(our data)
1.0-10-8
1.0-10-9
1.0-10-10
1.0-10-10*
5.0-10-10
2.0-10-9*
2.0-10-10
3.5-10-9*
5.0-10-9
4.3-10-9*
3.4-10-10*
4.0-10-8
8.8-10-10*
1.0-10-9*
5. MO-10*
1.7-10-9*
2.0-10-10*
Radionuclides limited by an ecologic criterion.
The calculation of the possible dose loads for fish is naturally approx-
imate. The addition of gamma radiation from the water and incorporated radio-
nuclides and the gamma and beta radiation from algae were not taken into
consideration. External radiation from the radionuclides contained in the
gastrointestinal tracts of fish can have a significant effect on the gonads
and kidneys of the fish (Orlov et al. 1978). In addition, as we observed
above, the doses of external radiation from the bottom shown in Tables 4 and 5
were determined according to limiting coefficients for accumulation of radio-
nuclides in the sand on Baltic Sea beaches. However, argillaceous soil in the
sea (Marey 1976; Ryndina 1970), whose sorption capacity is approximately 15
times higher than that of sand (Zlobin 1965) with all other conditions being
equal, has the highest capacity to accumulate radionuclides. For example,
Black Sea sludge accumulated 3.3 time? more 106Ru, 7.8 times more 137Cs, and
10 times more 144Ce than sand (Ryndina 1970). Higher doses of gamma and beta
radiation from the bottom should naturally be anticipated in this case.
In actual conditions, the presence of any one radionuclide in wastewater
from nuclear power plants, as well as the presence of the entire set of radio-
nuclides examined, are not very probable. In this case, the working limits of
the concentrations should be standardized not only for each individual radio-
isotope, but also for actual mixtures of isotopes present in the wastes from a
given plant.
176
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On the whole, the analytical evaluation of the possible doses of radia-
tion of fish in the presence of artificial radionuclides in sea water in
concentrations corresponding to those shown in Table 1 indicates the need for
making these limits more precise with respect to the interests of the fish
industry.
SUMMARY
The intensity of irradiation is dependent upon the concentration of
artificial radionuclides in organs and tissues of fish (inner sources) and
their distribution in water, vegetation, and bottom sediments (outer sources).
The value of the dose of outer irradiation is associated with biogeochemistry
of radionuclides in the aquatic environment. In those cases when radio-
nuclides are accumulated in much greater quantities in bottom sediments than
in water, the outer irradiation is responsible for the integral dose. The
admissible concentration of some radionuclides in water bodies is recommended
to be determined in relation to biogeochemical regularities of the behavior
pattern of radionuclides and intensity of fish irradiation.
REFERENCES
Aglintsev, K. K. , V. M. Kodyukov, A. F. Lyzlov, and Yu. V. Sivintsev. 1962.
Applied dosimetry. Moscow, Gosatomizdat Press, 248 pp.
Anonymous. 1976. Effects of ionizing radiation on aquatic organisms and
ecosystems. Tech. Rep. Ser. No. 172, IAEA, Vienna, 131 pp.
Anonymous. 1977. Production of atomic energy. Report of the OON Scientific
Committee on the Effect of Atomic Radiation. A/AS, 83/343. 107 pp.
Bakunov, N. A. , A. P. Panarin, and L. V. Fedotova. 1973. Accumulation of
Cs137 in commercial species of fish in the Caspian Sea. In Radioecology
of Aquatic Organisms, 2. Riga, Zinatne Press, pp. 229-234.
Chipman, W. A. 1972. Ionizing radiation. Jji Marine Ecology, a Comprehensive
Integrated Treatise on Life in Oceans and Coastal Waters, 1:3. Wiley-
Interscience, London-New York-Sydney-Toronto, 1578 pp.
Egami, Nabuo. 1973. Radioactivity and Fish (Japanese), Tokyo, 398 pp.
Garroll, D. 1959. Ion exchange in clays and other minerals. Bull. Geol.
Soc. Amer., 70:749.
Gromov, V. V., and V. I. Spitsyn. 1975. Artificial radionuclides in the
marine evironment. Moscow, Atomizdat Press, 224 pp.
Gusev, D. I. 1975. Hygienic criteria for evaluating contamination of coastal
marine waters by radionuclides. Reprinted from Impacts of Nuclear
Releases Into the Aquatic Environment, IAEA, Vienna, pp. 363-373.
177
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Haushild, W. L. , H. H. Stevens, et a_L 1973. Radionuclides in transport in
the Columbia River from Pasco to Vancouver, Washington. Geol. Surv.
Proceedings IV:43.
Katkov, A. Ye., D. I. Gusev, A. V. Dzekunov, M. I. Grachev, Ye. N. Lyapin, and
V. D. Stepanova. 1978. Effect of water temperature on accumulation of
radionuclides in fish. |ri Problems in the Radioecology of Cooling
Nuclear Power Plants. Svertlovsk, (UNTs AN SSSR), pp. 70-75.
Ketchum, B. H. 1960. Oceanographic research required in support of radio-
active waste disposal. Disposal of radioactive waste. Vienna,
2:285-291.
Khayn, J. and T. Braunell (eds.). 1958. Radiation Dosimetry. Moscow, IL,
758 pp.
Kranskopf, K. B. 1956. Factors controlling the concentration of thirteen
rare metals in sea water. Geochim. et cosmochim. acta, 9(1/2):1
Kulikov, N. V., L. N. Ozhegov, M. P. Chebotina, and V. F. Bochenin. 1978.
Accumulation of radionuclides by freshwater hydrobionts at different
water temperatures. In Problems in the Radioecology of Cooling Nuclear
Power Plants. Svertlovsk (UNTs AN SSSR), pp. 65-69.
Marey, A. N. 1976. Sanitary conservation of water from contamination by
radioactive substances. Moscow, Atomizdat Press, 222 pp.
Nelepo, B. A. 1970. Hyclear hydrophysics. Moscow, Atomizdat Press, 224 pp.
Orlov, E. V. , A. P. Panarin, and I. A. Shekhanova. 1978. Formation of the
dose load and effects of chronic irradiation of tilapia in strontium90
solutions. Moscow, Works of the VNIRO, Ecologic Aspects of Chemical and
Radioactive Pollution of the Aquatic Environment, 134(2):94-104.
Patel, B., C. D. Mulay, and A. K. Ganguly. 1975. Radioecology of Bombay
harbour—A tidal estuary. Estuarine and Coast. Mar. Sci., 3(l):13-42.
Patin, S. A. 1970. Radioactive contamination of the marine environment.
Moscow, TsNITEIRKh, pp. 1-42.
Patin, S. A., and A. A. Petrov. 1973. Artificial radioactivity in sea water
and commercial hydrobionts in the world's oceans. Ln Radioecology of
Aquatic Organisms, 2. Riga, Zinatne Press, pp. 200-209.
Pertsov, L. A. 1973. Ionizing radiation of the biosphere. Moscow, Atomizdat
Press, 288 pp.
Polikarpov, G. G. 1964. Radioecology of marine organisms. Moscow, Atomizdat
Press, 296 pp.
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Rozhanskaya, L. I. 1970. Manganese, copper, and zinc in water and organisms
in the Sea of Azov. In Marine Radioecology. Kive, Naukova dumka Press,
pp. 182-208.
Ryndina, D. D. 1970. Sorption and desorption of radionuclides by marine
sediments, algae, and detritus. Ibid., pp. 46-62.
Schafer, M. B. 1960. New research required in support of radioactive waste
disposal. Disposal of radioactive wastes. Vienna, 2:267-282.
Shekhanova, I. A. 1975. Biologic and piscicultural aspects of standardiza-
tion of the concentration of radioactive substances in the aquatic envi-
ronment. TsNIITEIRKh, 32 pp.
Shekhanova, I. A. 1976. Biologic evaluation of the effects of radioactive
contamination of the aquatic environment of fish. Moscow, VNIRO, 57 pp.
Shekhanova, I. A. 1978. The biologic role of artificial radionuclides in the
ontogenesis of fish. Auth. Abst. Doct. Diss., Moscow, 53 pp.
Svedov, V. P. , and A. A. Patin. 1968. Radioactivity in oceans and seas.
Atomizdat Press, 288 pp.
Timofeyeva-Resovskaya, Ye. A., and N. V. Timofeyeva-Resovskaya. 1960. Tr.
Ural'sk. fil. AN SSSR, No. 12, Sverdolvsk, p. 194.
Zlobin, V. S. 1965. Some features of the mechanism of contamination of
bottom sediments and soils by radioactive substances. In Problems in the
Radiation-Hygienic Investigation of the Sea. Moscow, Atomizdat Press,
pp. 93-108.
179
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THE CONSEQUENCES OF CHEMICAL POLLUTION OF THE "WATER-
BOTTOM SEDIMENT" CONTACT ZONE IN THE SEA
A. Bronfman and Z. B. Aleksandrova
Institute of Economics, Ukrainian SSR Academy of Sciences, and
Azov Scientific Research Institute of Fisheries
It has become obvious that the direct toxic effect of pollutants is not
the only channel through which they impact marine biocenoses. Indirect
biologic effects caused by transformations of the chemical system in polluted
waters also play an important role in this process.
Naturally, the relative importance of direct and indirect (intermediate)
effects will differ in each case. In the overall context of this question,
these aspects are of undisputed interest. However, the methodologically
correct organization of studies of the problem should be based on biogeocen-
otic principles, which suggest integral consideration not only of the
reactions in systems of the "pollutant-organism, population, biocenosis" type,
but also changes in the chemical system of the sea which could have a direct
effect on the degree of toxicity of the various pollutants or the level of
resistance of marine organisms to the toxins.
Objectively, there are various circumstances which stimulate orientation
of studies in this direction. They primarily include the fact that the corre-
sponding changes in the hydrochemical system of a body of water either signif-
icantly attenuate or, on the other hand, potentiate the result of the toxic
effect of pollution. This important concept is not new, but it has still been
inadequately considered in experimental studies on stimulation and prediction
of the biologic effects of marine pollution.
The transformation of chemical processes in polluted marine waters has
been studied comparatively little, especially with respect to the contact
zones with bottom sediment and the atmosphere. Vernadskiy (1960) called these
specific marine biotopes "zones of condensation of life11 and attributed an
enormous role to them in the biologic and geochemical structures of bodies of
water.
In the present report, the authors examine some aspects of this question
for the "water-bottom sediment" contact zone. The oxygen regime of bottom
waters and processes of the interchange of biogenic elements between bottom
residue in the marine waters covering them were selected as the parameters to
be examined. The result of many years of complex studies of the Azov Sea and
special laboratory experiments served as the empirical basis.
180
-------
Deterioration of the oxygen regime due to inhibition of photosynthesis
and to significant consumption of oxygen for oxidation of organic components
of wastewaters is one of the ways in which toxic substances affect biota.
According to the calculations of Oertzen (1972), 3.6 million tons of oxygen
are consumed annually in the Baltic Sea in the oxidation of organic matter
from domestic wastewaters.
More pronounced deterioration in the oxygen regime should be expected in
the bottom layers of the sea where pollutants accumulate and where aeration of
the water is poor. This situation is particularly dangerous because the
bottom layers of the sea are the habitat for zoobenthos and populations of
many commercial fish, and because important physical-chemical processes
affecting the chemical parameters and productivity of the pelagic zone occur
here.
There are currently only isolated data on the consequences of pollution
of the contact zone for the oxygen regime of bottom waters (Bronfman and
Gorstko 1978). The urgent need for reliable information on this subject has
become increasingly obvious.
Quantitative information on the effect of different pollutants on the
concentration of dissolved oxygen could undoubtedly be obtained by using
laboratory models of the process, but extrapolation of the experimental
results to actual conditions in a real body of water involves the risk of
significant error. More objective evaluations can be made on the basis of the
analysis of direct natural measurements which permit consideration of a number
of important physical-chemical factors whose simulation is difficult in
experiments or is not possible in general.
The problem was solved for petroleum products and detergents in our
study. The results of 114 synchronous measurements of different hydrochemical
parameters in the Azov Sea, conducted within a broad range of fluctuations,
were the basis of the calculations. Multiple regression analysis was used to
determine the simultaneous effects of many parameters as well as the relative
effect of each parameter.
The following empirical models were investigated:
V = f/E.CBOCj)^, HB, Hg/
K02 = f/E' V' V V
where Kn is the concentration of dissolved oxygen in the layer 0.5 m from the
2 bottom (range of values studied from 0.21 to 9.80 mg/liter);
E is the total vertical stability of the water in the surface-bottom
layer (range from 19.72 to 3.85 thousand arbitrary units);
(BOCT) is the daily biochemical oxygen consumption by the surface layer of
gr the bottom sediment (range from 10.09 to 0.50 g 02/m2 per day);
181
-------
Hg, H is the concentration of petroleum products in the bottom layers
y (range from 0.30 to 6.0 mg/liter) and the surface layer of the
sediment (range from 0.03 to 9.35 mg/g);
DR, D is the same for detergents (range from 0.02 to 0.75 mg/liter and
y from 0.03 to 6.27 mg/g).
The models satisfactorily approximate the phenomenon investigated—the
multiple correlation coefficients consisted, respectively, of 0.627, 0.643,
and 0.874, while the numerical values of Fisher's dispersed ratio (F) were
8.77, 11.11, and 6.67; with these degrees of freedom, the critical distribu-
tion values for F exceeded 99%.
Based on the values of Fisher's criterion and the level of significance a
(Table 1), the arguments for the models based on their effect on the oxygen
regime of bottom waters were distributed in the following sequence: vertical
stability of the water, biochemical oxygen consumption by the contact zone of
the sediment, detergents in the bottom layers of the sea, detergents and
petroleum products accumulated in the surface layer of the sediment. The
petroleum products contained in the bottom layers have an insignificant effect
on the concentration of oxygen dissolved there, and are not a significant
factor according to the criteria selected.
TABLE 1. VALUES OF FISHER'S CRITERION (F), LEVELS OF SIGNIFICANCE (a), AND
REGRESSION COEFFICIENTS (a) FOR THE ARGUMENTS OF THE MODELS STUDIED
Arguments of Model F a a
E
(BOCl)gr
DB
V
V
HB
6.63 -
6.57 -
6.49 -
3.68 -
1.04 -
0.02 -
9.85
6.82
9.49
5.34
4.42
0.03
^ 99%
^ 98%
£ 98%
^ 90%
> 75%
--
-7.0 x 10-5
-8.4 x 10-5
-0.20
-0.23
-5.00
-5.66
-0.30
-0.40
-0.10
-0.15
-0.06
182
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Judging by the value of the regression coefficient (a), each mg of petro-
leum products and detergents entering the bottom layers of the sea with other
conditions being equal reduces the concentration of oxygen in these layers by
0.06 and 5.0-5.7 ml/liter, respectively. Petroleum products and detergents
accumulated in the surface layer of the bottom sediment were represented,
respectively, by 0.10-0.15 and 0.30-0.04 ml/liter.
Thus, even "moderate" pollution of bottom water and sediment by organic
components from waste waters results in a significant decrease in the concen-
tration of oxygen in the benthoic zone leading to asphyxiation and death of
benthoic organisms. This is most evident in the case of detergents. The
effect of petroleum products on the oxygen regime in the "water-bottom sedi-
ment" contact zone is relatively weak, possibly due to the entry of heavy,
slowly oxidized petroleum fractions in the bottom layers of the sea.
The development of an oxygen deficiency and organic pollution of the
benthoic zone is a relatively rapid process; even in the absence of any
anthropogenic impurities, oxygen consumption due to biochemical oxidation of
organic matter in the surface layer of bottom sediments usually takes place
intensively. In particular, in clean or slightly polluted zones of the Azov
Sea in the warm season of the year, oxygen consumption for oxidation of 1 g of
organic matter in the sediment consists of 25-60 mg of oxygen per day
(Aleksandrova and Romova 1977). The value of (BOC,) is equal to an average
of 3.7 here, attaining 10.5 g 02/m2 per day. ^r
All of the above increase the need for a quantitative evaluation of the
potential risk of an oxygen deficiency due to pollution of the benthoic zone
of the sea. The time (T) during which the oxygen reserve in the bottom layers
is completely exhausted for biochemical oxidation of organic matter accum-
ulated in the "water-bottom sediment" contact zone on cessation of vertical
water exchange can be used as the criterion. The oxygen reserve (RQ ) is
calculated as: 2
R02 ~ (K02)100% " (K02)60%'
where (Kn )inn U2 DlM>
60%, respectively, saturation of the water in the bottom layer.1
In view of the above:
" (BOCI}gr r = (K02}60%'
1 The value of 60% was selected as the upper limit for asphyxiation of organ-
isms inhabiting the "water-bottom" contact zone.
183
-------
hence T -
nence, i
100%
0.7(BOCj)(
where h is the height of the bottom layer used in the calculations, and 0.7 is
the transition factor from weight to volume units of oxygen.
Experimental determination of time T for a given body of water can be
made for "critical" phases of an oceanographic regime characterized by the
maximum frequency and duration of stagnation of most of the indigenous organic
matter entering the benthoic zone and the highest activity for its biochemical
oxidation.
This approach was first implemented in developing a method for preserving
the natural complex of the Azov Sea. The calculations (Aleksandrova and
Romova 1977) were made for 20 basic oceanographic stations based on averaged
data from special complex observations made in the summers of 1969-1977. The
value of the biochemical oxygen consumption in the bottom layers of the sea,
consisting of approximately 10% of the (BOCT) was not considered in the
calculations.
(BOCj)
The results obtained allowed zoning the Azov Sea according to the time of
potential possible depletion of the oxygen reserve in the bottom layer (Figure
1), and also establishing that the majority of the sea is characterized by a
value of T < 0.75 days; it thus exhibits minimum natural resistance to organic
pollution.
Figure 1. Zoning of the Azov Sea according to the time for potential possible
depletion of the oxygen reserve in the bottom layer. 1 = T < 0.5
cyt. ; 2 = 0.5 < T < 0.75 cyt. ;3 = 0.75
-------
Based on an analysis of all of the data systems obtained in conducting
the studies, a number of recommendations were developed to improve the system
for protecting the Azov Sea from pollution. Certain additions were also
introduced into the existing programs for ecological monitoring.
In the overall context of the question examined, the fact that disturbing
the oxygen regime in the bottom layer of the sea, organic pollution of the
bottom can result in a decrease in the exchange processes in the "water-bottom
sediment" system and significant corrections can be made in the chemical
balance of the bottom, is of significant interest.
Unfortunately, these studies are still in the initial stage. At the same
time, the available data indicate the need to activate them, also within the
framework of solving problems related to the consequences of pollution of the
sea.
These findings, obtained during limnological (Hatchinson 1969; Martynova
1973; Lars 1974; Fillos and Biswas 1976; Lech 1977) and marine studies
(Pirogova 1953; Yrukovskiy 1972; Fonselius 1972; Aleksandrova and Bronfman
1975; Rowe et aj. 1975), basically can be interpreted in the form of the two
major positions stated below.
1. The direction and degree of migration of biogenic elements in the
"water-bottom sediment" system are decisively regulated by the concentration
of oxygen, the pH value, and the oxi dative- reductive conditions in the contact
zone, i.e., those physical-chemical parameters which are most strongly altered
under the influence of pollution.
2. Exchange of biogenic elements in the direction of the bottom
pelagic zone takes place relatively intensively, and in this respect should be
considered an important element in the biogenic balance of the photic zone of
the sea. Thus, according to the evaluations of Rowe et al. (1975), desorption
of ammonium nitrogen from bottom sediments is equal to 14 mg/m2 per day on the
continental shelf of the ocean under conditions close to anaerobic. According
to Yurkovskiy's observations (1972), desorption of phosphate phosphorus in
stagnating basins in the Baltic Sea consists of 9 mg/m2 per day.
The geochemical mechanism of exchange and its role in the biogenic
balance has been more completely studied in the Azov Sea, which is the most
appropriate natural model for investigating this process based on a number of
morphological -structural features and the features of the physical-chemical
regime. The almost annual summer stagnation of the water, characterized by
extremely high indexes for the vertical stability of the layers (up to 96,000
arbitrary units), and the development of an oxygen deficiency and reducing
conditions in the contact zone of the sea, is one of the important features.
In individual years, iso-oxygen of 60% satuation and zero isovolts are mapped
here for areas of up to 20,000 km2, and the oxi dative- reductive potential of
the bottom sediments is altered in the range of 300 to -200 mV (Aleksandrova
and Bronfman 1975; Bronfman 1976).
Observations showed that sorption processes, which result in impoverish-
ment of the bottom layers and biogenic elements, primarily develop in condi-
tions of sufficient aeration of the contact zone of the sea. In conditions of
185
-------
stagnation, the active desorption in the water through the reducing surface of
the bottom sediment is observed. Typical situations for phosphates are illu-
strated in Figure 2.
0
2
"E
4
1C
t
LU
Q 6
8
in
I I ..... .Q.
Eh = + 0.286 /
/ a
/
/
- /
t
•/•0.268 «
1
1
1
1
/+0.258
J +0.084 ^
—
-
-
-
-
"
-
-
9 02 ml/LITER
p-P04mg/m3
"Eh = +0.158*: <
1
•2
0
E
-' 4
a.
i"J e
o 6
•20 40 60 8.0 '
9 02 ml/LITER
p-P04 mg/m3
0.6
'0*0. • • ..'.-...•.•.•.-
"-'• •:•:•': '-Eh = -0.048
i.'.".:~ .". r .•.•.•*•.•!". jr.*. . .
V- +0.026 ~
Figure 2. Vertical distribution of dissolved oxygen (1), phosphate (2), and
oxidative-reductive potential with different degrees of aeration of
the water in the Azov sea: (a) conditions of adequate aeration;
(b) conditions of stagnation.
The results of natural observation, special laboratory studies conducted
in a broad range of oxi dative-reducing conditions, and also the data of
Hatchinson (1969), Ayvazova and Fedosov (1972), Yurkovskiy (1973), Shippel and
Hallberg (1973) and Lech (1977) permit drawing the following conclusions with
respect to the basic mechanisms of exchange of nitrogen, phosphorus, and
silicon compounds in the "water-bottom sediment" system.
Desorption of phosphorus compounds from the bottom sediment primarily
occurs in the form of phosphates whose transition to the aqueous phase and
reducing conditions of the medium is basically due either to concentration
diffusion or an increase in the solubility of the phosphoric acid potassium
and iron salts accumulated in the sediment. Biochemical dephosphorylation of
the organic matter in the sediment also plays a determining role in this
process. The identity of the physical-chemical properties of the phosphate
and silicate anions forms the basis for hypothesizing that the changes in
solubility and the ion exchange reactions in the Si03 "-Ca-Fe" system are also
the basis for desorption of silicon. The yield of nitrogen is basically
determined by anaerobic ammonification of organic matter with subsequent
dissolution of ammonium salts.
186
-------
The data on special vertical zoning of the "water-bottom" system obtained
over many years allowed conducting a statistical study of the functions AP =
f(Eh) and AP = f(Kn ), where AP is the difference in the concentration of
U2
phosphate in the 5 cm contact layer of water and at a distance of 0.5 m from
the bottom. This index has been used as an indirect criterion of the nature
and intensity of exchange.
As indicated by the results obtained, in the 20-25°C temperature range,
the limit values of the oxidative-reducing potential of the surface layer of
sediment and the concentration of oxygen in the contact zone of the sea which
cause a change in the direction of migration of biogenic elements are, respec-
tively, equal to 20 mV and 4.5 ml/liter (60% saturation).
The results of laboratory simulation of exchange processes have confirmed
these limit values for Eh and Kn (20 mV and 3.5 ml/liter). This also allowed
U2
estimation of the diffusion coefficients for biogenic elements from the bottom
sediment of the bottom layers of the sea. Based on average data, the latter
consist of 4'10-3 mVday for phosphates and silicates, and 10-2 mVday for
ammonium.
According to the calculations performed, the mean rates of desorption of
phosphate, silicon, and ammonium nitrogen from bottom residue in the Azov Sea
were equal to 7, 190, and 150 mg/m2 per day, respectively.
Together with the materials on the duration of stagnation in the areas of
distribution of the oxygen deficiency in bottom layers, the data cited allowed
establishment of the fact that desorption of phosphates, ammonium nitrogen,
and silicic acid from the sediment in the bottom layer of the Azov Sea
consisted of an average of 1.5, 28, and 35 thousand tons/year, respectively,
for the period investigated (1970-77).
The values cited consist of approximately 27 and 20% of the total intake
of nitrogen and phosphorus compounds from other external sources (river
drainage, precipitation, conversion of shores, exchange with the Black Sea),
and thus objectively demonstrate the exclusive role of desorption processes in
formation of the chemical bases of marine productivity. Characteristically,
the amount of desorbed nitrogen and phosphorus for the period analyzed
consisted of approximatly 50% of their discharge with river water, and
drainage basins with an area greater than 500,000 km2 on the average.
The significant role of exchange processes in the biogenic balance of the
Azov Sea is not a regional exception. In particular, according to the calcu-
lations of Yurkovskiy (1973), the amount of mineral phosphorus desorbed from
sediment consisted of more than 58,000 tons only in the Silurian basin of the
Baltic Sea for the stagnation period of 1931-33 and 1935-59. Yurkovskiy
(1973), Fonselius (1972), and Sjoberg et aJL (1972) believe that diffusion of
phosphates from bottom sediments is a factor which enriches the entire aqueous
layer of the Baltic Sea.
Based on the above, it is possible to speak of the difference in the
chemical-biological reactions of the benthoic and pelagic regions of the sea
in pollution of the "water-bottom" contact zone.
187
-------
In the first case, the reaction takes place according to the scheme—
accumulation of pollutants, increase in BOC, depletion of oxygen reserve, and
formation of anaerobic or similar situations. The negative effects of this
situation for the biocenosis in the contact zone are obvious and require no
commentary.
In the second case, a decrease in the oxidative-reducing potential of the
contact zone, up to the onset of reducing conditions, sharply activate desorp-
tion of biogenic elements in the surface layer of the sediment and enrichment
of the bottom and overlying layers of the sea. The last circumstance is
certainly a factor which to some degree stimulates processes of primary
production of organic matter for the organic life and the pelagic zone in
general. In contrast to the benthoic zone of the sea, the possibility of
partial, total, or excessive compensation of the negative production effect of
the toxic influence of pollutants is also totally real in this case.
This interpretation of the facts presented in the report is still
inadequately supported by a system of empirical evidence. It is still to some
degree a hypothesis whose development and refinement would obviously be
facilitated by a more objective understanding of the complex biogenocenolog-
ical consequences of pollution of the sea.
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