SOURCES, FATES AND EFFECTS OF AROMATIC HYDROCARBONS IN THE
ALASKAN MARINE ENVIRONMENT WITH RECOMMENDATIONS
FOR MONITORING STRATEGIES
by
J. N. Anderson
Pacific Northwest Laboratory
Bat telle, Marine Research Laboratory
439 West Sequim Bay Road
Sequim, Washington 98382
and
J. M. Neff and P. D. Boehm
Battelle, New England Marine Research Laboratory
397 Washington Street
P.O. Box AH
Duxbury, Massachusetts 02332
Project Officer
James C. McCarty, Deputy Director
Environmental Research Laboratory
U.S. Environmental Protection Agency
200 S. W. 35th Street
Corvallis, Oregon 97333
Under
Interagency Agreement No. TD 1668
to the U.S. Department of Energy
Richland, Washington 99352
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SOURCES, FATES AND EFFECTS OF AROMATIC HYDROCARBONS IN THE
ALASKAN MARINE ENVIRONMENT WITH RECOMMENDATIONS
FOR MONITORING STRATEGIES
by
J. W. Anderson
Pacific Northwest Laboratory
Battelle, Marine Research Laboratory
439 West Sequim Bay Road
Sequim, Washington 98382
and
J. M. Neff and P. D. Boehm
Battelle, New England Marine Research Laboratory
397 Washington Street
P.O. Box AH
Duxbury, Massachusetts 02332
Project Officer
James C. McCarty, Deputy Director
Environmental Research Laboratory
U.S. Environmental Protection Agency
200 S. W. 35th Street
Corvallis, Oregon 97333
Under
Interagency Agreement No. TD 1668
to the U.S. Department of Energy
Richland, Washington 99352
-------
The information in this document has been funded wholly or In part by
the United States Environmental Protection Agency under EPA-82-D-X0533
to Pacific Northwest Laboratory. It has been subject to the Agency's peer
and administrative review and it has been approved for publication as
an EPA document. Mention of trade names or commercial products does not
constitute endorsement or recommendation for use.
ii
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CONTENTS
Figures Iv
Tables v
Acknowledgment vii
1. Introduction 1
2. Sources of Aromatic Hydrocarbons in the Alaskan
Marine Environment 3
Biosynthesis 3
Fossil Fuels 4
Pyrolysis 4
Rates and Routes of Entry of PAH into the
Marine Environment 10
Alaskan Perspective 13
3. Fate of Oil Spills and Chronic Effluents 14
The Geochemlcal Environment ............... 14
Fate of Oil In the Arctic Environment 15
General Considerations 15
Case Studies 23
Alaskan Perspective 29
4. Transformation and Degradative Processes 30
Photodegradation ..... . 30
Transformation by Bacteria, Fungi, and Algae 34
Transformation by. Marine Animals 37
Interaction of Degradative Processes in the
Removal of PAH from the Marine Environment 41
Alaskan Perspective 42
5. Accumulation and Release by Marine Animals 43
Bloaccumulation from Uater and Sediment 43
Distribution of Accumulated PAH in Animal Tissues .... 53
Release of PAH 53
Factors Controlling the Uptake and Release of PAH .... 59
Relationships Between Tissue Contamination by PAH
and Effects 60
Alaskan Perspective 61
6. Biological Effects of PAH 63
Toxicity of Specific Compounds 63
Toxicity of Oil Extracts 65
Toxicity of Oil to Different Life Stages 69
Acute Toxicity from Flow-Through Tests 74
Sublethal Effects . 74
Toxicity of Hydrocarbon-Contaminated Sediments 78
Effects on Individual Species 79
ill
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CONTENTS (CONTINUED)
Biological Effects of Oil Spills and Chronic
Low-Level Discharges of Hydrocarbons .... 81
Recovery of Benthic Communities 86
Summary 90
Alaskan Perspective 91
7. Information Needs and Recommended Monitoring Strategies .... 92
Information Needs 92
Monitoring Programs 94
8. Summary 97
References 99
FIGURES
Number
1 Transport and Interaction path of oil and suspended
particulate matter 17
2 Transport of oil to benthos due to sorption and sinking 20
3 Hypothesized methods by which oil sinks and remains on
the bottom 21
4 Transport of weathered oil to benthos in low-density
water 22
5 Mass balance results from the Amoco Cadi2 oil spill 24
6 Concentrations of oil along a transect oriented to the
northeast of the IXTOC I blowout* September. 1979 26
7 Summary of comparative fates of oil from the BIOS
experimental spill ..... 28
8 Annual variation of the near-surface half life (tso) of
benzo(a)pyrene at several northern latitudes 33
9 Concentrations of aromatic hydrocarbons in pink shrimp,
pink scallops, pink salmon fry, and king crab during
exposure to the water-soluble fraction of Cook Inlet
crude oil and during subsequent depuration 51
10 Tissue distribution of naphthalenes 54
11 Accumulation and depuration of naphthalene In early
life stages of coho salmon 58
iv
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FIGURES (CONTINUED)
12 Concentrations of specific petroleum compounds producing
50% mortality for selected marine organisms 64
13 Concentrations of No. 2 Fuel oil WSF producing 50%
mortality in 96-h exposures to various life stages
of crustaceans and polychaetes 72
TABLES
Number
1 Concentrations of PAH in two crude oils and two refined
oils 5
2 Typical concentrations or concentration ranges of the
PAH carcinogen, benzo(a)pyrene (BaP), in petroleum and
coal products 6
3 Typical concentrations of benzo(a)pyrene and total PAH
In gaseous and particulate emissions from burning of
different fuels 8
A Summary of the major oil spills for which extensive
post-spill fate and effects studies were performed 16
5 Computed half-lives for direct photochemical transformation
of several PAH 31
6 Summary of aromatic hydrocarbon uptake by marine organisms .... 44
7 The effect of molecular weight on the accumulation of aromatic
hydrocarbons by pink salmon 52
8 Summary of depuration kinetics for petroleum hydrocarbons 56
9 The concentrations of n-parafflns and aromatic
hydrocarbons in reference oils and the 10% water soluble
fractions (WSFs) prepared from them 66
10 Composition of extracts from crude and refined oils and
tanker ballast treatment effluent 67
11 Acute toxicity of petroleum to marine animals 68
12 Acute toxicity of water-soluble fractions of a crude and
refined oil to arctic marine species from four phyla 70
v
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TABLES (CONTINUED)
13 Concentrations of naphthalenes in specific oils producing
50% mortality for several marine Invertebrates .......... 71
1A Summary of effects of petroleum hydrocarbons on the
growth and reproduction of marine animals 75
15 Sublethal effects of the water-soluble fraction (VSF)
and Individual components of crude oil on Alaskan marine
organisms 77
16 Summary of effects of oiled sediments on Alaskan marine
invertebrates 80
vi
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ACKNOWLEDGMENT
The authors are grateful to the U. S. Environmental Protection Agency
for the funding required to prepare this document. We appreciate the
constructive comments of outside reviewers and those of our colleague,
Dr. James States in our Anchorage office. Thanks also go to Jan Engel and
Joan Pfeifer for the preparation of the manuscript and illustrations.
vii
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SOURCES, FATES AND EFFECTS OF AROMATIC HYDROCARBONS IN THE
ALASKAN MARINE ENVIRONMENT WITH RECOMMENDATIONS
FOR MONITORING STRATEGIES
1. INTRODUCTION
Representatives from 19 Federal and state agencies met In 1982 to
prioritize 17 problems being considered for funding by the Cold-Climate
Environmental Research Program of the U.S. Environmental Protection Agency.
Ranked fourth In Importance was the need for an overview of the current
state of knowledge on fate and effects of oil-derived hydrocarbons in cold
marine waters. Although a moderate amount of research has been done in this
area, many of the results have not yet been published or if published, have
been in obscure journals and the so-called "grey" literature. There was a
need to collect the most relevant literature and present It from the
viewpoint of professionals in the field as a basis for determining the
future directions of research.
The Importance of such an effort was underscored by Battelle Memorial
Institute's working session on "The Determination of Unreasonable
Environmental Degradation in Alaskan Marine Waters" held in Anchorage,
Alaska during March of 1984. Approximately 30 key individuals active In
offshore oil development were invited from the oil industry, state and
Federal agencies and other Interested organizations to seek ways of
resolving current Impasses.
The transport, fate and effects of petroleum-derived hydrocarbons,
particularly aromatic hydrocarbons that might be released in discharges of
produced waters (PW), was a major topic of discussion. Participants were
asked to develop lists of useful things to do and, using a computer-assisted
process, rank them from most to least useful. Among the agency and Industry
representatives there was remarkable consensus. From a list of 31 useful
things to do, 8 of the- 9 identified as most useful and 8 of the 10
identified as least useful by industry were so identified by the agencies
as well. On the other hand, the limited native representation ranked the
priorities almost the reverse. For example, items ranked 1 and 6 from the
native perspective were ranked 24 and 31 by the agencies and industry. The
need for more extensive dialogue between industry and Federal/state agencies
on one hand and native organizations on the other was one of the important
conclusions of the working session.
The PW activities seen as most useful by the agencies and Industry were
to:
1
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(1) Develop a management decision framework for PW discharges and the
criteria to be applied In Its use.
(2) Determine the behavior of PW discharge plumes and sedimentation
of contaminants from plumes, including the evaluation and
verification of existing dispersion models.
(3) Evaluate acute toxicity of components from PU discharges where the
toxicity of the components is not already known.
(4) Identify areas of high potential for contaminant persistence and
impacts from PW discharges.
(5) Evaluate potential chronic effects of components from PW
discharges.
(6) Examine the feasibility and consequences of alternatives to PW
discharge to the ocean.
The objective of this report is to critically review what is known
about the sources, fates, and effects of polycyclic aromatic hydrocarbons
(PAH) in the Alaskan marine environment. . Based on this review, we identify
several information needs and recommend the design of research and
monitoring strategies to fill these needs.
While a considerable amount of information has been generated regarding
the fates and effects of petroleum components, there are untested conditions
and unique areas in the Alaskan marine environment that require special
consideration. It is Important to first bring to light the findings in
these research areas that are relevant to Alaskan conditions and then to
identify gaps in ous- understanding which require further experimentation.
First, we will examine the natural and anthropogenic sources of
aromatic hydrocarbons in the Alaskan marine environment. Then we will
review the physical, chemical, and biochemical fates of these compounds in
marine ecosystems. Finally we will review the current knowledge on the
bloaccumulation and biological effects of aromatic hydrocarbons in marine
organisms. Based on this review, we will make recommendations concerning
research needed to better understand the processes affecting aromatic
hydrocarbons in the Alaskan marine environment.
Since its inception in 1975 through an interagency agreement between
the Department of the Interior, Bureau of Land Management (now Minerals
Management Service) and the Department of Commerce, National Oceanic and
Atmospheric Administration, the Outer Continental Shelf Environmental
Assessment Program (OCSEAP) has provided financial support and technical
guidance for a very large number of Investigations dealing with all aspects
of the impacts of oil and gas exploration and development on the Alaskan
marine environment. A recent comprehensive bibliography (OCSEAP, 1984)
lists titles of more than 3,000 documents dealing with research supported by
OCSEAP. We have made extensive use of these documents in preparing this
review, particularly those documents in the form of final reports and
publications in the hard literature.
2
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2. SOURCES OF AROMATIC HYDROCARBONS IN THE ALASKAN MARINE ENVIRONMENT
Polycyclic aromatic hydrocarbons (PAH) may be formed by a variety of
mechanisms. These include the very rapid, high temperature (e.g., 700°C)
Incomplete combustion or pyrolysls of organic materials, the very slow
(e.g., millions of years) rearrangement and transformation of organic
materials at moderate temperatures of 100-300°C to form fossil fuels, the
relatively rapid (days-years) transformation of certain classes of organic
compounds in soils and sediments, and possibly biosynthesis.
BIOSYNTHESIS
It is well-known that a wide variety of organic molecules containing
fused-ring polyaromatic systems are synthesized by organisms, particularly
bacteria, fungi, and higher plants. Many of these compounds are not true
PAH's since they contain oxygen, nitrogen or sulfur substituents. For
example, microorganisms, higher plants, and even a few animals synthesize
a wide variety of polycyclic qulnone pigments. Vitamin K2 is a
naphthoquinone. The marine marsh grass Juncus roemerianus synthesizes large
amounts of a pigment called juncusol which is a phenanthrene quinone.
Under oxygen-free conditions, as occur in many marine and freshwater
sediments, these quinones and phenols may be reduced to the parent PAH. A
four-ring PAH, perylene, frequently occurs at high concentrations in
anaerobic freshwater and marine sediments (Aizenshtat, 1973; Gschwend et
al., 1983; Louda and Baker, 1984).
It is thought that perylene is formed in anoxic sediments by the
reduction of extended qulnone pigments synthesized by fungi and other
mlcrooganisms living at the sediment/water Interface. Under oxidizing
conditions, certain chemicals such as abletlc acid from pine tree resin and
triterpenes (hopanes) from mosses and certain other plants can be oxidized
to PAH (Tan and Heit, 1981). Retene (methyl-lsopropyl phenanthrene) is a
common PAH found in coniferous forest soils. The PAH assemblages produced
by the early dlagenesis of suitable biogenic precursors will have a simple
composition.
All attempts to date to demonstrate the direct biosynthesis of PAH by
bacteria, algae and higher plants have yielded negative or equivocal results
(see review by Neff, 1979). If direct biosynthesis of PAH does occur, it is
of minor importance in the overall mass loading of PAH in the environment.
3
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FOSSIL FUELS
Transformations similar to those described above for the early diagenic
formation of perylene and alkyl phenanthrenes, if allowed to continue for
millions of years under the right temperature and pressure regimes, lead to
the formation of fossil fuels containing a great diversity of hydrocarbons,
Including PAH.
Coal generally is considered an aromatic material. Host of the PAH in
coal are tightly bound In the coal structure and cannot be leached out.
Total PAH concentrations tend to be higher in hard coal than in soft coals,
like lignite and brown coal. However peat deposits prevalent over vast
areas of the arctic tundra, may contain high concentrations of some PAH,
particularly perylene.
Crude petroleum, shale oil and most refined and residual petroleum
products are extremely complex mixtures of many thousands of organic
compounds. Typical crude petroleums may contain from 0.2 to more than
7 percent PAH. Kerosene, gasoline, and dlesel oil contain relatively low
concentrations of tricyclic and larger PAH, while heavy oil products such as
Bunker C oil (residual oil) and asphalt may contain several percent PAH
(Neff, 1979). Shale oil and coal-derived synthetic crude oils may contain
as much as 15 percent PAH.
The four API reference oils have been used extensively In biological
effects studies' and considerable information is available about their
chemical composition (Anderson et al., 1974a; Pancirov and Brown, 1975).
Concentrations of several PAH in these oils are listed in Table 1. Most
oils contain the same hydrocarbons, but the relative proportions of
different hydrocarbons vary widely from one oil to another. Nearly always,
alkyl homologues are present at higher concentrations in crude and refined
oils than are the parent PAH (Blumer, 1976). This is reflected in the
relative concentrations of naphthalene and Ci-C3 naphthalenes in the four
API oils.
Benzo(a)pyrene (BaP) often is used as a marker or indicator of the
presence of carcinogenic PAH in environmental samples. Typical
concentrations of BaP in crude, refined and residual oils and coal are
summarized in Table 2. Crude petroleum and products refined from crude
petroleum generally contain low concentrations of BaP. Shale oil and
coal-derived oils contain substantially higher concentrations. Residual
oils and coal contain Intermediate amounts. Concentration of BaP in motor
oil increases during usage through pyrolysis.
PYROLYSIS
A majority of PAH in the environment is formed during the incomplete
combustion of organic natter at high temperatures (Suess, 1976). When
heated to high temperature, organic molecules are broken into smaller
fragments (pyroly6ls) which then are rearranged into different
configurations, Including PAH (pyrosynthesls), as they cool. Optimal
conditions for PAH pyrosynthesls Include a fuel-rich flame (high fuel to air
4
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TABLE 1. CONCENTRATIONS OP PAH IN TMO CRUDE OILS AND TWO REFINED OILS
CWKUHlllttal fcl Wtt/la *—>
ConpMid
kutfi
Lwdsltvw Ckv
Kimh
tft Crudt
No. 2
PMl OU
KmMmI i
Nipfcthftknc
too
400
4000
1000
l-mvlhylnaphttoJcn*
too
500
1200
2S00
2-mct>))rin*phthalm
900
700
11900
4700
DlmcthylnafhthaltMS
MOO
2000
>1100
12300
Trimrthylnapbthftltms
MOO
1900
1M00
ssoo
PJuortnM
200
<100
5600
2400
Phtnanthrtne
70
26
429
412
l.mtthylpheumhrtfw
III
9
173
43
2>nt*tlqrlph>nun)ir«nc
I4«
19
7677
S28
Pluoranthtne
3.0
2.9
>7
240
Pyitrm
3.5
4.5
41
23
6cn*(a)anthraccne
1.7
2.3
1.2
90
Chrytcnc
17.56
6.9
2.2
196
Triphcnylcm
10
2.1
1.4
11
BcrunfghiXluortnthcne
1
<1
Beuto (bMviorwtttene
<0.9
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TABLE 2. TYPICAL CONCENTRATIONS OR CONCENTRATION RAMSES Of THE PAH
CARCINOGEN BENZO(t)PYRENE(BaP)f IN PETROLEUM AND COAL PROOUCTS
ItfCwcumtoi
ItaMil Of/Id fern)
Crude Petroleum
aio->.c
Shale Oil
10-30
Diesel Fuel
0.2-0.7
No. 2 Heating Oil*
0.01-0.05
Gasoline
0.0J-6.2
Meter Oil* (new)
0.06-0.23
Meter Oil* (toed)
0.07.35
No. » Residual Fuel Oil
2.1
No. 5 Residual Fuel Oil
1.1.).)
No. < Residual Fuel Oil (Bwker C)
x-u
Asphalt
0.1-27
Co*J-Derlvetf Oil*
1.MS0
Bittonlnous Coal
C-20
6
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ratio) and a flame temperature above about 500°C. Nearly all of the
airborne PAH produced by flame pyrolysls are associated with the particulate
fraction produced during combustion (soots and carbon blacks).
Unlike PAH assemblages in fossil fuels, PAH assemblages formed by
pyrolysls are dominated by the parent compound with a decreasing frequency
of PAH homologues with increasing numbers of alkyl carbons. These
differences between composition of PAH assemblages in fossil fuelB and
combustion sources may be used to aid in the identification of the sources
of PAH in environmental samples.
A great many domestic and industrial activities Involve pyrosynthesis
of PAH. The resulting PAH may be released to the environment in airborne
particulates or in solid or liquid by-products of the pyrolytlc process.
Domestic activities that produce significant quantities of PAH Include
cigarette smoking, home heating with wood or fossil fuels, waste
incineration, broiling and smoking of foods, and use of internal combustion
engines. Many industrial activities produce large quantities of PAH. These
include coal coking; production of carbon blacks, creosote, coal tar and
related materials from fossil fuels; petroleum refining; synfuel production
from coal; and uBe of Soderberg electrodes in aluminum smelters and iron
works, etc. (Neff, 1979).
The burning of various fuels aB a source of heat or energy is one of
the most important sources of PAH in the environment (Table 3). The
particulate fraction of the exhaust from gasoline and diesel engines
contains high concentrations of one to six-ring PAH (Stenberg et al., 1983).
Modern fossil fuel power plants and municipal waste incinerators emit
relatively small amounts of PAH in stack gases or fly ash per unit fuel
burned. However, because of the large amounts of fuel or waste burned, mass
emission rates may be quite high.
Home heating with wood, coal and peat in stoves, fireplaces or hot
water boilers, which is common in rural and even many residential areas in
Alaska, results in the emission of relatively large amounts of PAH in smoke.
Domestic wood-burning stoves produce the greatest PAH emissions,
particularly when the fire is banked or allowed to smolder. Air pollution
associated with use of wood- and coal-fired stoves for home heating has
risen to the point that this type of heating has been prohibited in parts of
several cities in Canada and Alaska, Including Juneau. Forest fires and
open burning of brush and plant wastes are also Important sources of
particulate emissions rich in PAH. Sullivan and Mix (1983) estimated that
as much as 1.3 kg of total PAH were generated per hectare during slash
burning of clear-cut logging sites in Oregon.
Solid wastes (ash and creosote) from coal and wood burning also contain
significant concentrations of PAH. The relative Importance of stack gas,
ash, and creosote as sources of PAH depends on the ratio of stack gas to ash
produced during combustion, which in turn depends on the type of wood or
coal burned, and the combustion conditions.
7
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TABLE 3. TYPICAL CONCENTRATIONS OF BENZO(a)PYRENE AND TOTAL PAH IN
6ASEOUS AND PARTICULATE EMISSIONS FROM BURNING OF DIFFERENT
FUELS •
sasBBaBaaBassaam
tmbalon Rate ftna/fca fuel toned)
Ml InaUMvm TMalPAH
Cm), Pmr Plant (Mpdem)
*>10»l4JQ2
<0X23
Oil, Fewer Plant
<0.04
RcIum, Municipal Incinerator
~*10-'
0.009
Gasoline, Automobile, Noneatalytic
0416
•
Gasoline, AutomeUlt, Catalytic
*«I0**
•
Coal, Residential Stove
25
•
Wood, ResJdentlaJStove
0.0V6J
2.6-230
Charcoal,. Residential Stove
Uil^
0.093
Wood, Residential Hot Water BoUar
0.02
1.9a
Peat, Residential Hot Water Boiler
0J6
1«.9
Wood, Residential Fireplace
0.00 V 1.9
•
Bnoh, Open Slash Burning
3-12
73.176
*D*u from Davie* et al., 1976j Aiiberg and Stenberg, 1979j DascK, 1982j Ramtfahl ct al.,
iW2j Ahiberg «t al, 1983; Sullivan and Mis, 191); Knight tt *1, 1983.
8
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Many industrial processes produce by-products or effluents rich in PAH.
The best known is the production of coke from coal. Coke is an important
high-energy fuel used extensively by the iron and steel industry. The
particulate fraction of stack gases from a coking oven nay contain up to
five percent PAH (Lao et al., 1975).
Catalytic cracking of crude petroleum fractions to produce hydrocarbon
fuels and other refined petroleum products results in the production of PAH.
These tend to become concentrated in the heavy residual fractions, Bunker C
or No. 6 fuel oil and asphalts. Significant quantities of PAH may be
released in flue gases and in waste water effluents. Most of the PAH are
removed by various methods before the waste streams are released to the
environment. The treated flue gas from an oil refinery may contain 1-10 ug
total PAH/m3 (Sawlckl et al., 1965).
Activated sludge treatment of liquid waste effluents from oil
refineries removes up to 95 percent., of the five-ring PAH and more than
99 percent of the four-ring PAH (Panclrov et al., 1980). However, the small
amount of PAH remaining in the effluent includes known carcinogens (e.g.,
methyl- and dimethyl benzanthracene, and benzo(a)pyrene). Jensen (1983)
reported that a refinery effluent discharged to a Danish estuary contained
an average of 0.08 parts per billion benzo(a)pyrene. Sediments and mussels
(Mytllus edulis) from the estuary contained 27-54 ppb and 5-87 ppb,
respectively of benzo(a)pyrene.
Several metal-smelting processes produce gaseous and liquid waste
streams rich in PAH. Soderberg electrodes are sometimes used in aluminum
smelters, ferrosilenium and iron works, and calcium carbide production.
Soderberg electrodes are made of anthracite coal, coke tar, pitch and
anthracene oil. During metal smelting the electrode is burned continuously
and the high boiling PAH escape in the fumes (Palmork et al., 1973). The
production of one ton of aluminum consumes half a ton of electrode
material. Particulate PAH may escape to the environment in the flue gas, in
water effluent from the gas scrubber system, or in the sludge resulting from
treatment and recycling of the scrubber water. Bjorseth and Eklund (1979)
identified more than 100 PAH and related heterocyclic compounds in the
working atmosphere of an aluminum smelter. Palmork et al. (1973) identified
and quantified several PAH in waste water and sludge effluents from
Norwegian aluminum smelters and In bottom sediments and associated benthic
biota from the fjords into which effluents were discharged. The scrubber
sludge contained about 700 mg/kg (ppm) low molecular weight PAH
(anthrace-pyrene). Contaminated fjord sediments contained more than
1700 ppm of these compounds.
There are, of course, many other natural and anthropogenic sources of
PAH. The extreme diversity of sources of environmental PAH and our limited
knowledge about them makes any estimate of the global flux of these organic
molecules very difficult.
9
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RATES AND ROUTES OF ENTRY OF PAH INTO THE MARINE ENVIRONMENT
PAH formed by the processes discussed above may reach marine and
freshwater environments by a vide variety of -routes. Major routes of entry
of PAH to marine waters include early diagenesis of suitable precursors,
spillage and seepage of fossil fuels, discharge of Industrial and domestic
waste water, fallout or washout in precipitation from air, and runoff from
land.
In situ formation of PAH will be restricted to sediments of anoxic
basins or other restricted environments. and is unlikely to contribute
significantly to the overall mass loading of PAH In Alaskan coastal waters
and sediments. However, peat moss, as Indicated above, can contain high
concentrations of certain early diagenic PAH (i.e., perylene and retene).
There are extensive deposits of tundra peat In northern and interior Alaska
(Schell, 1983). Riverine and coastal erosion results in substantial Inputs
of this material to Alaskan coastal waters. It has been estimated that
coastal erosion alone results in the Introduction of 700 metric tons of peat
per km of coast per year to the Alaskan Beaufort Sea. Major rivers such as
the MacKenzie and Yukon also Introduce massive amounts of peat to coastal
waters.
Spillage and seepage or erosion of fossil fuels into the ocean is an
Important source of PAH in this environment. World oil production was
estimated at about 3,100 million metric tons In 1961 (Miller and Connell,
1982). These authors estimated that from 4.5 to 6.1 million tons of
petroleum (0.15-0.20Z of production) enters the marine environment each
year. Since crude and refined oils may contain up to several percent total
PAH, this accidental spillage and natural seepage represents a
quantitatively Important Input of PAH to the marine environment. Erosion or
disposal of coal or coal dust also may be an important source of PAH in
coastal environments (Tripp et al., 1981).
Blasko (1976) described 18 oil and 7 tar seep areas in the northern
Gulf of Alaska. There are gas seeps in Norton Sound near Nome (Cline and
Holmes, 1977), and In the southern Bering Sea near the Pribllof Islands
(Kvenvolden and Ridden, 1980). There are coal outcrops along the Mead
River and an oil seep area near Smith Bay In the Alaskan Beaufort Sea (Shaw
et al., 1979). The MacKenzie River drains the Athabasca tar sands and
Normal Wells seep area of north-central Canada and discharges to the
Beaufort Sea a short distance west (up-current) of the Alaskan border.
These natural seeps or eroslonal sources may contribute significantly to
PAH assemblages detected in Alaskan coastal sediments (Shaw et al., 1979;
Stlch and Dunn, 1980; Atlas et al., 1983).
Two activities associated with oil production and transport in Alaska
which result In the discharge of large amounts of hydrocarbon-contaminated
waste water are produced water production and ballast water treatment.
Most petroleum reservoirs also contain fossil water. When the oil Is
removed from the formation, some of this water, called produced water or
oilfield brine water, comes with It (Neff, 1985). This produced water
usually has a salt concentration as high as or higher than that of seawater
10
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and also contains high concentrations of petroleum hydrocarbons. The
produced water, may be reinjected through another well back into the
reservoir or it may be treated to remove particulate oil and discharged to
the surface. Although some produced water may reach arctic wetland tundra
and coastal areas via its discharge into reserve pits and subsequent
discharge or leaching, the only.,place in Alaska where treated produced water
currently is discharged directly to the marine environment is in Cook Inlet.
Lysyj (1982) estimated that in 1982 an average total of 12.9 million liters
per day of produced water was' discharged to Cook Inlet from two on-shore
treatment plants (at Kenai and Trading Bay) and from one offshore production
platform (Dillon).
There are several discharges of produced water to Cook Inlet other than
those studied by Lysyj. These Include a treatment plant at Granite Point,
and platforms Anna, Bruce, Baker,, and the Phillips A,gas platform. Thus,
the rate of produced water discharge to Cook Inlet is substantially greater
than the 12.9 million liters/day reported by Lysyj (1982). A reasonable
estimate would be twice as much or 25.8 million liters/day. The treated
produced water effluent contained 5.8 to 42 ppm particulate oil (mean 21
ppm) and an average of 10 ppm volatile monocyclic aromatic hydrocarbons
(benzene, toluene, .xylene). Unfortunately, Lysyj (1982) did not measure
concentrations of PAH in the produced water effluent. Armstrong et al.,
(1979) reported that a produced water sample from the Texas coast,
containing about 10 ppm monoaromatlcs, contained a total of 1.74 ppm two-
and three-ring aromatics. If we used a value of 1.5 ppm as an average
concentration of lower molecular weight PAH in produced water, then a total
of 38.7 kg/day or 14.1 metric tons/year of low molecular weight PAH are
released to Cook Inlet with about 826 metric tons/year of total petroleum
hydrocarbons in produced water. Whereas reinjection often is a feasible
alternative to surface discharge in on-land oil fields, this usually is not
the case for offshore fields. Therefore,, if or when oil production occurs
in the Beaufort and Bering Seas, there may be a need to discharge treated
produced water to these water bodies. The current permit for development of
the Endicott Field in the Beaufort Sea calls for reinjection of produced
water.
Crude oil tankers arrive in ballast at Valdez to take on oil from the
southern terminus of the trans-Alaska pipeline. The tankers discharge their
ballast water to a treatment facility before taking on oil cargo. The
Valdez ballast water treatment facility discharged a total of 33.4 billion
liters of treated ballast water, containing an estimated 132 metric tons of
particulate oil and 170 metric tons of volatile hydrocarbons (mainly
benzenes) during the first two years of operation, August 1977-July 1979
(Lysyj et al., 1981). The effluent probably contained no more than 0.25 ppm
of low molecular weight PAH (Rice et al., 1981). Based on these figures,
approximately 4.2 metric tons/year of low molecular weight PAH were
discharged to Valdez Harbor in treated ballast water.
A massive oil spill may result in the introduction of massive amounts
of PAH within a relatively small area. For instance, approximately 250,000
metric tons of crude oil was released and spread to approximately 350 km
of the Brittany coast of France after the Amoco Cadiz spill. If this
oil contained two percent total PAH, a total of 5,000 metric tons of PAH was
11
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released to the marine environment within a few days. There have been no
major oil spills in Alaskan coastal waters. However, small spills (less
than 50 barrels: 7,900 liters), do occur occasionally during normal tanker
and offshore production platform operations. Should development occur
in the Norton Sound, Navarin, North Aleutian or St. Georges Basin areas of
the Bering Sea or in the Chukchi Sea, the produced oil probably would be
transported to refineries via tankers, Increasing the likelihood of spills
from this source.
Industrial and domestic sewage often contains high concentrations of
particulate and soluble PAH. Several of the industrial sources of these PAH
were discussed above. Storm sewage (runoff from roadways, etc.) contains
PAH arising from wear and leaching of asphalt road surfaces, wear of vehicle
tires which contain carbon black, condensation from vehicular exhaust, and
from the all too common practice in Alaska and elsewhere of oiling roadsides
and unpaved roadways with used crankcase oil (high in PAH). Several
investigators have identified used crankcase oil as a major source of PAH in
stormwater runoff from urban and rural areas (Hoffman et al., 1984).
Treated and sometimes untreated liquid sewage is nearly always
discharged to any available water body (river, lake, estuary). Solid
residues from activated sludge treatment of wastes may be disposed of in
the ocean or in land fills. Liquid dometlc sewage usually contains less
than 1 ug/liter total PAH, Industrial sewage from 5 to 15 ug/liter PAH, and
sewage sludge from 1 to 30 mg/kg (Neff, 1979).
Most of the PAH emitted to the atmosphere during the burning of organic
matter are adsorbed to microscopic particles (Suess, 1976). Atmospheric
processes that result in the deposition of airborne PAH on land and water
include washout in precipitation, dry fallout and vapor phase deposition
onto surfaces. For many years it was assumed that airborne particulate PAH
were degraded rapidly by photooxidation processes (NAS, 1972). There is
growing evidence, however, that airborne particulate PAH may be transported
in environmentally significant concentrations over thousands of kilometers
(Lunde and Bjorseth, 1977; Bjorseth et al., 1979).
There has been growing Interest in recent years in the source and
composition of seasonal atmospheric haze that occurs throughout much of the
Arctic each year. This haze, called the Arctic aerosol, contains
relatively high concentrations of several pollutant metals (Heidam, 1984)
as well as organic pollutants, including PAH (Dalsey et al., 1981). The
concentration of pollutants in the Arctic aerosol is highest in the winter.
The pollutantB are thought to come from major industrial sources in the
middle latitudes (i.e., the central U.S., Europe, China and Russia). At
Point Barrow, the mean concentration of total PAH was 1.25 ug/m3 in March
and 0.14 ug/m3 in August (Dalsey et al., 1981).
The March samples contained fly ash particles, suggesting industrial
coal and oil burning as as source of the PAH. In August, the PAH profile
resembled that of vehicular exhaust particulates, probably of local origin.
The concentrations of PAH in the Arctic aerosol at Point Barrow, Alaska,
are similar to those in other remote atmospheres, and much lower than
12
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concentrations typically reported for urban atmospheres (e.g., 6-9 ug/m3 in
New York: Daisey et al., 1981. and 2.9-155 ug/m3 in Fairbanks, Alaska:
Reichardt and Reidy, 1980). Shaw et al. (1979) suggested that long-range
transport of air-borne particulates from mid-latitude industrial sources was
a major source of PAH in coastal sediments of the Alaskan Beaufort Sea.
The estimated annual Input of PAH to the marine environment from the
Bources discussed above is approximately 230,000 metric tons total PAH/year
(Neff, 1979). Surface runoff from land and fallout from the air appear to
be the main sources of high molecular weight PAH in the marine environment,
while petroleum seepage and spillage is the main source of total PAH. This
reflects the relatively high PAH content (mainly bl- and trl-cycle PAH) of
petroleum in comparison to other major PAH sources.
ALASKAN PERSPECTIVE
Although several point sources of polycycllc aromatic hydrocarbons in
the Alaskan marine environment have been identified (i.e., produced water
and ballast water discharges), these probably do not contribute more than a
small fraction of the total PAH entering Alaskan coastal waters from all
sources. In non-lndustrlal areas of the world, including most of Alaska,
major sources of PAH in coastal waters and sediments are from aerial
deposition of particle-associated PAH derived from remote industrial and
other combustion sources (Neff, 1979; Hites et al., 1980; Johnson et al.,
1985). In Alaska, wood burning for home heating and in controlled or wild
forest fires are major sources of airborne PAH which may eventually reach
the marine environment. Erosion of peat and coal deposits probably are
major sources also, particularly to the Beaufort and Bering Seas.
13
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3. FATE OF OIL SPILLS AND CHRONIC EFFLUENTS
In the past twenty years there have been many massive spills of crude,
refined and residual petroleum products to the marine environment as a
result of oil tanker accidents and blowouts on offshore oil platforms.
Following the peak years of 1978 (Amoco Cadiz Oil Spill) and 1979 (IXTOC 1^
blowout) the amount of oil spilled annually in the ocean had declined until
1983* when it rose again (by 9-fold over 1982) to approximately 242 million
gallons (912 million liters) (OSIR, 1984). The large increase in oil
spilled in 1983 was due to oil well blowouts in the Arabian Gulf as a result
of the Iran-Iraq war, and six tanker accidents, each involving loss of 1
million gallons or more of oil. There were no major reported spills (more
than 10,000 gallons: 37,854 liters) in Alaska in 1983. Between 1971 and
1980, the average rate of spillage of oil in Lower Cook Inlet was 0.0001
percent of the total crude oil produced (Wondzell, 1981), which is a
considerably better record than that for the Gulf of Mexico.
THE GEOCHEMICAL ENVIRONMENT
Several recent studies have documented the distribution of hydrocarbons
and metals in Beaufort Sea sediments. (Shaw et al., 1979; Kaplan and
Venkatesan, 1981; Venkatesan and Kaplan, 1982; Naidu et al. 1981; Burrell et
al. 1981). Shaw et al. (1979) examined the hydrocarbon geochemistry of
nearshore sediment along a transect from Point Barrow to Barter Island.
Total hydrocarbon concentration in the nearshore ranged between 0.3 and
20 Ug/g dry sediment. The saturated hydrocarbons were dominated by
n-alkanes ranging in chain length from 23 to 31 carbon atoms with a strong
odd-even preference and no unresolved complex mixture (UCM) evident. This
distribution is consistent with a prevalent biogenic input of terrigenous
plant material, most likely resulting from transport of riverine suspended
particulate matter during the spring runoff. Shaw et al., also examined
sources of aromatic hydrocarbons in nearshore sediments using the alkyl
homolog distribution of selected aromatic hydrocarbon series determined by
GC/MS. Distributions characteristic of both pyrogenlc and petrogenic
origins were observed. The geographic distribution of pyrogenlc combustion
products was indicative of difuse non-point source Inputs (probably aerial
deposition of particulate PAH), rather than point source Inputs. Their
analysis also ruled out natural seepage or spills of Prudhoe Bay crude oil
as the source of aromatic hydrocarbons in the region.
The hydrocarbon geochemistry of the Beaufort Sea outer continent shelf
(OCS) has been studied by Kaplan and Venkatesan (1981) and Venkatesan and
Kaplan (1982). The range of total hydrocarbon concentrations reported was
20-50 Ug/g dry sediment, which is slightly greater than that found in the
14
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nearshore sediments. Whether these differences are due to differences in
the analytical methods employed or to a greater abundance of fine-grained,
organic-rich sediments in the OCS region was not investigated. As with the
nearshore sediments, the major source of saturated sedimentary hydrocarbons
was found to be higher plant debris, with no evidence of a UCM indicative of
anthropogenic inputs. A marine biogenic origin for some of the organic
matter was also indicated by the occurrence of the hydrocarbons prlstane
and n-heptadecane. The occurrence of several alkanes, together with
steranes, diterpanes and trlterpanes also attested to the biogenic origin
of the organic matter. Measurable concentrations of aromatic hydrocarbons
were found in almost all the Beaufort Sea OCS sediments examined. The
distribution of alkylated homologues determined by GC/MS was found to be
characteristic of pyrogenic origin. Thus, the available organic geochemical
data for the region indicate that hydrocarbons found In nearshore and
offshore sediments originate primarily through natural processes, with
little evidence of anthropogenic petroleum Inputs.
Data on the concentrations of hydrocarbons in bivalves from the region
are sparse. Shaw et al. (1981) reported very low levels of hydrocarbons in
bivalves from the area.
FATE OF OIL IN THE ARCTIC ENVIRONMENT
General Considerations
Although there have been few actual case studies concerning oil spills
in Alaskan waters, many of the studies conducted in sub-Arctic and temperate
marine environments have provided Information which is directly applicable
to addressing the fate of oil spilled in Alaskan waters, and on the fate of
polycyclic aromatic hydrocarbons in petroleum.
There has been a large number of studies of the fate of spilled oil.
The oil spills for which major fate and effects studies were performed
during the last two decades are summarized in Table A. In addition, there
have been several investigations of the marine environmental fate and
effects of chronic low-level discharges from point sources (e.g. produced
water effluents, ballast water effluents, oil-base drilling fluid cuttings
discharges, refinery effluents, etc). These studies have demonstrated that
PAH from petroleum may be quite persistent in some marine environments and
may produce massive immediate biological damage as well as long-term impacts
In some low-energy depositlonal marine environments.
The fate of oil spilled onto the surface of the water generally can be
depicted as In Figure 1. This generalized schematic addresses oil spilled
on the water's surface prior to landfall. A variety of processes can act on
the oil and partition it into various compartments in the marine environment
(i.e. water column, sediments, biota).
Each of these processes may act differentially on the various
components of oil. It is widely recognized that during the initial stages
of a spill or a crude oil discharge, the major transport mechanism of oil
15
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TABLE 4. SUWMRY OF THE NUOft OIL SPILLS FOR WHICH EXTENSIVE POST-SPILL FATE AND EFFECTS STUDIES
MERE PERFORMED
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16
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Photo Oxidation
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Evaporation
Slick
Surface
• Oil In
Water Emulsion
Desorptlon.
Dissolution
Oil Emulsion
Dissolution
• 'Particles
(Sorption, Agglomeration)
Resuspenslon and
Surface Return
Sedimentation
Figure 1 . Transport and interaction path of oil and suspended particulate natter.
-------
Into the water column Is by physical dispersion. This dispersion results in
a spectrum of whole oil droplets in the upper water column. Subsequently,
soluble aromatic hydrocarbons, the one- and two-ringed aromatics,
preferentially dissolve in the water aB a result of the increased surface
area of the droplets and due to the relatively high solubility of these
aromatic components. As a result, the chemical composition of hydrocarbons
in the seawater exclusive of the droplets themselves is distinctly different
than that of the whole oil droplets. The water-soluble and more toxic
aromatlcs dominate the dissolved fraction. Animals exposed to the water
mass will initially acquire these compounds. However, the total
concentration of these dissolved aromatlcs will seldom exceed 1 ppm and
usually is in the 0.01 to 0.1 ppm range. The instantaneous concentration of
aromatlcs and for that matter of the whole oil droplets themselves, is
directly dependent on the physical energy of dispersion, i.e., the result of
wind and wave Induced turbulence.
The worst case of water column exposure is actually represented by
either, (1) the subsurface discharge of oil from an oil well blowout (see
Brooks, et al., 1981, Boehm and Fiest, 1982) to be discussed below, or by,
(2) a chemically dispersed oil plume, also to be discussed below. In these
cases, maximum petroleum hydrocarbon levels seldom exceed 50 ppm, with these
concentrations rapidly diluted by mixing with seawater by factors of 103 to
105.
The potentially more serious environmental effects related to petroleum
hydrocarbon transport mechnanisms and specifically aromatic hydrocarbons
relate to: (1) the sorption of oil droplets on suspended particulate
material (SPM) and transport to the bottom; and (2) the landfall or beaching
of an oil slick, followed by sorption onto beach sediments and subsequent
erosion of these sediments and deposition in the nearshore zone.
Suspended particulate matter Interacts with spilled oil through two
major processes. The first process which is considered to be quantltively
minor Is the molecular sorption of dissolved oil species from the water
phase onto suspended particles. The second process is the physical
collision and adherence of dispersed drops of oil with SPM. The resulting
oil-loaded particulate may lose sorbed hydrocarbon components through
desorptlon or microbial biological processes or may ultimately be deposited
on and Incorporated into the seabed. In Alaska, SPM distributions from
riverine runoff and other coastal sources, are often highly concentrated in
the nearshore zone (Brltch, 1976; Barnes and Fox, 1979; Strakov, 1967) and
are transported parallel to the coastline. Thus, the sorption efficiencies
of SPM for petroleum hydrocarbons, the mechanical loading potential of SPM,
along with distribution patterns and abundance of SPM, are potentially
important parameters of the transportation and fate of spilled oil in
Alaskan coastal waters.
Although numerous studies have been performed on the adsorption of oil
to SPM (Gearing et al., 1980; Zurcher and Thuer, 1978; Payne, et al.,
1984), the detailed dynamics of this process have not been described
in a quantitative manner. In general, the variability in bulk oil
characteristics such as density and viscosity must be considered, but the
18
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key controlling variable appears to be the amount of SPM, and specifically
the percent clay in the SPM (Zurcher and Thuer, 1978; Meyers and Quinn,
1973; Payne et al., 1981; Meyers and Oas, 1978; Bassin and Ichlye, 1977).
These Investigators generally agree that, at equilibrium, between 120 mg and
300 mg of hydrocarbons are incorporated Into each kilogram of clay sediment.
Studies of oil spills have shown that In order for significant
quantities of oil to reach the bottom, oil must sorb to suspended sediment
and sink. The Santa Barbara oil spill Is a good example (Kolpack et al.,
1971), where transport and deposition of large quantities of flood runoff
material occurred during the spill. This appears to be the major transport
route of oil to the benthos (Figure 2), although other mechanisms can become
significant In certain cases (Figure 3). Most notably, the water column to
benthlc transport of oil can occur by fecal pellet transport (Boehm, et al.,
1982; Johanssen, et al., 1980), by sinking due to Langmlur circulation, or
by direct sinking of.. dense (cold, weathered) oil in areas of low density
water (Figure 4) (fresh water input at river mouths or near ice melting).
Studies by Juszko, et al., (1983) have suggested that the weathering of oil,
typically of an initial density of 0.9 g/ml can create a residue having a
density of 1.02 to 1.03 g/ml. Once oil achieves such a density, which is
greater than fresh Ice melt, oil can sink. The behavior of oil, once it
sinks, will be determined by the vertical density structure of the water
column. In coastal Alaskan waters, oil may sink to the bottom by this
mechanism.
Under conditions of low suspended particle concentrations in the water
column (1 to 10 ppm), no significant quantitative transport of
particle-sorbed oil to the benthos will occur. Under conditions of
moderate suspended particle concentrations (10 to 100 ppm), significant
quantities of oil may be sorbed to particles if sufficient mixing of oil
and available particles occurs. Under extreme conditions of Influx of
particles (100 ppm) and mixing of oil and particles in the water column,
massive transport of sorbed oil to the offshore benthos can occur with
possible severe offshore Impact (see Figure 2).
The transport of biologically pelletized oil (zooplankton fecal
pellets) to the bottom certainly has occurred and has resulted in
analytically detectable oil residues In the bottom with some biological
consequences (Johanssen et al., 1980). In some spill scenarios, this may
represent the major transport path to the benthos with regards to the
overall fate of the mass of spilled oil.
Sedimentation of petroleum hydrocarbon residues directly to the
seafloor may occur where discharge of organic-laden drilling fluids occurs.
This process depends on the nature of the associated hydrocarbons. A light
distillate such as diesel oil or a light fuel oil will largely (80%)
partition into the aqueous phase or evaporate (Breteler et al., 1985).
Heavier oil and hydrocarbons associated with the formation during drilling
are more apt to remain associated with particles and settle with drilling
mud particles.
19
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Offshore
¦j^Rlvtr Moot*
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Relative Degree of Pollutant
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Moderate
Low'
Figure 2. Transport of oil to benthos due to sorption and sinking.
-------
ENTRAINMENT
I IN /
CIRCULATION CELLS
MIXES WITH SEDIMENT
AND STAYS ON BOTTOM
©
OIL MIXEO WITH SEDIMENT ON BEACH AND TRANSPORTED
SEAWARD BY BOTTOM CURRENTS
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Figure 3. Hypothesized methods by which oil sinks and remains on the bottom.
-------
K)
K)
reshwater
Ice Melt
nput
V \
Pycnocline
Figure 4. Transport of weathered oil to benthos In low-density water.
-------
The amount of petroleum hydrocarbons reaching the sediments by any of
the possible transport paths depends on water depth (and hence dilution) and
the offshore settling regime (i.e., waves, currents and tidal energy).
It is also widely accepted that significant transport of petroleum
components including toxic aromatics to nearshore sediments will occur
after beaching of oil, followed by beach erosion and/or ice scour. Studies
In the Canadian Arctic (Baffin Island Oil Spill Program, Boehm et al., 1985)
and in the Amoco Cadiz spill (Marchand and Caprais, 1981; Boehm et al.,
1981; Atlas et al., 1981; Gundlach et al., 1983) have shown that large
amounts of oil (100 ppm in sediments) can be expected to impact nearshore
(subtidal sediments) if shoreline Impact Is allowed to occur.
The interaction of oil and sea Ice represents a specialized case which
is particularly of significance in Alaskan waters. Numerous studies of the
physical Interactions of crude oil and sea ice have been undertaken (e.g.,
Lewis, 1976; N0RC0R Engineering, 1975). When oil is released Into the water
column, the oil droplets rise and upon coming in contact with ice, spread
under ice and accumulate in depressions. As additional ice freezes under
the ice sheet, the oil is effectively trapped. The trapped oil penetrates
only slightly Into the ice sheet until thawing begins at which point oil
penetrates into brine channels within the thawing ice. During this period
of time the oil is cut off from evaporation, the mechanism that allows
volatile aromatics to "escape" the water column. The effects on under-lce
biological communities of exposure to the relatively toxic unweathered oil
are likely to be severe. Eventually the trapped oil, which may have been
transported over considerable distances while entrapped, escapes to weather
normally.
Case Studies
Of particular relevance vis-a-vis the fate of petroleum and petroleum
aromatics in Alaskan waters are results from the Amoco Cadiz Oil Spill; the
IXTOC _I blowout; the TsesiB Oil Spill and the Baffin Island Oil Spill (BIOS)
experiments.
The fate of the Amoco Cadiz cargo of crude oil has been extensively
studied by Atlas et a!. (1981), Boehm et al. (1981), and Gundlach et al.
(1983). Gundlach et al. (1983) have conducted perhaps the most rigorous
mass balance of spilled oil ever undertaken. Their results are summarized
in Figure 5. In this spill, a massive beaching, and other direct
introduction of oil into the nearshore sediments and a mixing of oil into
the anoxic sediment layers resulted in a long term (3+ years) Impact of
oil. Biodegradation did not affect oil in the sediment's anoxic zone
(Winfrey et al., 1982) and the oil remained toxlcologlcally potent for long
periods of time. The most persistent hydrocarbon components, both in the
sediments and In benthic animals, were the abundant three ringed PAH
compounds (i.e., phenanthrenes) and the polycycllc aromatic sulfur
compounds, the dlbenzothlopbenes, present at 1-10 ppm out of a total
aromatic hydrocarbon concentration of 1000 ppm. Most of the 1000 ppm
consisted of uncharacterized PAH compounds.
23
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Subtldal Sediments
18,000 tons (82)
Water Column
30,000 tons (13.52)
Blodegraded
10,000 tons (4.52)
Evaporation
67,000 tons (302)
First Month Total Spilled: 223,000 tons
Onshore
€2,000 tons (282)
Unaccounted For*
46,000 tons (20.52)
~Probably surface slicks
and tar balls.
B
Maximum 011
No 011
liTflT'V-Abers In Water Column (tons)
: y\ Onshore (tons)
•V Onshore (kilometers of shoreline)
ft / * t-W Subtldal Sediments (tons)
fc u \ \ Nearshore
V
'\
w
\
f
Tanio
, xj
I
ec
H i'i
(9
1978
1 1 1 1 1
ill i
1 cc
a.
i ii i i
cc
a.
ee <9
a 3
3
3 <
C
<
¦t
1979
1980
1981
Figure 5. £ass balance results from the Amoco Cadiz oil spill from Gundlach
et al., (1983). A - Quantitative estimate of oil fates for the first month
afteT the spill; B - Synthesis of data Indicating the relative persistence
of oil in various compartments (Tanio - a new oil spill event.)
24
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The IXTOC _I blowout, in the Gulf of Mexico is hardly typical of an
Alaskan spill. However, subsequent studies of the BIOS program have
indicated that the results from . Boehm and Flest's (1982) IXTOC _I water
column studies are quite relevant. These authors found that, due to the
undersea nature of the IXTOC blowout, petroleum residues persisted in the
water column for up to 40 km from the blowout site. While concentrations
dropped off rapidly, an Important subsurface plume of oil droplets mixed
with an aromatic-enhanced dissolved/colloidal mass was advected up to 20 km
from the blowout site (Figure 6). This was a result of a relatively high
current speed (0.5-1 knot) and a neutrally bouyant oil mass introduced at
the sea floor and rising only slowly to the surface. The ultimate fate of
this oil was not determined, but the investigators hypothesized that a
significant' mass of oil (approximately 30% .of the spill) evaporated, with
another 30% remaining at the water surface as tar balls, and perhaps another
10-30% of the oil sorbed to particulates and possibly distributed over a
wide area in the bottom sediments.
The BIOS program was conceived and executed to examine the comparative
fate and effects of crude oil discharged into nearshore Arctic environments
and either (1) dispersed chemically or (2) left untreated and allowed to
Impact the shoreline. The subsurface nature of the discharge of the
011-chemical di6persant mixture allowed for comparison with the IXTOC
blowout. It was found that the chemically dispersed plume consisted of
neutrally buoyant droplets of relatively unweathered oil similar in
composition and behavior to the IXTOC plume, despite drastically different
temperature regimes. These plumes persisted for considerable periods of
time (hours to several days) in the water column. No significant
sedimentation of chemically dispersed oil occurred, nor was there any
significant beach impact. By contrast, the untreated oil spillage did
result in a massive beaching of the oil. Of the untreated oil initally
Impacting the beach, 0.2 to 2% was eroded and transported offshore after 1
year. This value Increased to 1-10% after two years. Boehm et al., (1985)
hypothesized that this erosion would continue to result in increased
subtldal oil levels with time. The other main conclusions of this study
were (Boehm et al., 1985):
o Chemically dispersing the oil resulted in large initial water column
concentrations (50 ppm).
o Initial water-borne oil concentrations under the untreated surface
oil slick were low and were confined to the top 1 meter of the water
column.
o Longer-term monitoring (one and two years post spill) revealed only
low levels of water column oil in all bays (1 ppb or less).
o Subtldal sediment impact was low (10 ug/g) due to chemically
dispersed oil-sediment interactions. However, sediment oil
concentrations due to the untreated oil Increased over two years
(up to 400 ug/g) as oil eroded from the beach.
25
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(Vertical (uggvitim 5;000*;VV•araa of MnewrMiwM > lOCfca^l; arai of concentration 5— IOmpg/1).
Figure 6. Concentrations of oil along a transect oriented to the northeast of the
IXTOC I blowout, September, 1979.
-------
o Compositional determinations Indicated that biodegradation occurred
in the beach sediments* but did not occur subtidally.
o The dispersed oil discharge resulted in an initially large
bioaccumulatlon of water-borne oil by all bivalves (up to 700 ppm).
Concentrations decreased with time so that two years after the spill
most species had returned to near-background levels. Initial tissue
levels did not directly reflect water column exposures as animals
tended to reduce or cease filtration when exposed to large
quantities of petroleum hydrocarbons.
o Initial subtldal benthlc faunal Impact due to the untreated oil
release was low. Concentrations of petroleum hydrocarbons in
deposit feeders were, however* seen to Increase with time while
concentrations in filter feeders decreased.
o Biodegradation of oil occurred within the gut of all benthic
animals.
The conclusions of the comparative analyses of the fate of chemically
dispersed and untreated oil in the Arctic are summarized in Figure 7. From
this analysis, it may be concluded that the net result of the dispersed oil
release was a 6hort-term extensive water column and related bioaccumulatlon
Impact with little offshore or beach sediment Impact. The benthic animals
exposed to this dispersed oil depurated steadily, but took approximately 2-3
years or more to return to pre-splll chemical conditions.
In contrast, the water column hydrocarbon Impact from the untreated oil
release was minimal. However, the stranding of untreated oil on the beach
resulted in a long-term source of oil to subtldal sediments and increasing
of levels over time of hydrocarbons in benthic animals.
The Tsesis oil spill occurred In the archipelago, south of Stockholm,
Sweden in 1977. An extensive interdisciplinary study was undertaken for
several years. The results are presented in Klneman et al (1980), Linden
et al (1979), and Boehm et al (1982). Several findings regarding the fate of
the distillate oil spilled are of particular Interest. During the weeks
following the spill, many of the zooplankters contained visible oil droplets
in their gut. Subsequently, significant quantities of oil were sedimented
via the fecal transport route to the bottom. Ten to fifteen percent of the
oil spilled during this event was sedimented via incorporation into fecal
pellets or adsorption onto suspended sediment particles, as evidenced from
sediment trap studies. The monitoring strategy of examining filter-feeding
bivalves (Kytilus edulis) as indicative of water column oil exposure, and
deposit-feeding bivalves (Macoma balthica) as indicators of sediment
exposure was extremely effective in this study. This strategy was used
successfully again in the BIOS program and Indicated that lnital water
column exposure is large, but that complete tissue depuration of aromatic
hydrocarbons occurs in one to three years. In contrast, oil residues
persist longer in the bottom sediments where chronic exposure persists for
many years (more than 3 years), the duration depending on the continued
source of hydrocarbons to the benthos.
27
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EavItimibUI
Wiltr Celumn
Binthle Faun*
UatratUf
oue»m
i 1 r
t 1 r
Clialcili WihtnI
OR *01
t 1 r
8ufetid«t 6«dlm«nt«
Btacb Sidlmant*
—I 1 I—
I860 1981 1981 1903
(SpIlD
t 1 r
1 T
I960 1981 1982 I98S
(SpllD
Figure 7. Summary of comparative fates of oil from the BIOS
experimental spill.
28
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ALASKAN PERSPECTIVE
Hydrocarbon assemblages In sediments of the Alaskan coastal and
offshore areas Investigated to date appear to be primarily of natural
biogenic origin. Evidence of pyrogenlc PAH assemblages presumably derived
from deposition of airborne particulate PAH has been reported even in
sediments from fairly remote areas. When oil is spilled in arctic and
subarctic marine environments similar to those in Alaska, It tends to be
more persistent than oil spilled under similar circumstances in temperate
and tropical climates. In coastal areas characterized by high suspended
sediment loads, such as Cook Inlet, the Beaufort Sea and parts of Norton
Sound off the Yukon River, petroleum-derived aromatic hydrocarbons will
rapidly adsorb to suspended particles and be transported to the bottom. In
bottom sediments, they may be quite persistent. Evaporation of the more
toxic low molecular weight aromatic hydrocarbons is Impeded by low water
temperatures and particularly by ice cover.
29
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4. TRANSFORMATION AND DEGRADATIVE PROCESSES
Aromatic hydrocarbons in the marine environment are subject to chemical
transformation and degradation through a variety of processes. The most
Important of these are photo-oxidation, chemical oxidation and biological
transformation by marine bacteria, fungi, plants, and animals. Since a
majority of natural mechanisms of oxidative transformation of PAH require
molecular oxygen, sunlight, or both, PAH are generally quite stable and
persistent once incorporated into oxygen-poor bottom sediments of water
bodies.
PHOTODEGRADATION
Zepp and Schlotzhauer (1979) and Mill et al. (1981) have shown that
when PAH are present in solution in "pure" fresh or sea water, direct
photolysis reactions are quantitatively the most Important mechanisms of
light-induced transformation of PAH. These direct photochemical
transformations can take place In the absence of molecular oxygen. There
is a strong tendency for sensitivity to direct photolysis in the aqueous
phase to increase with Increasing PAH molecular weight (Table 5). Thus
naphthalene is quite insensitive to photolysis, while benzo(a)pyrene is
quite sensitive. Linear PAH like anthracene and naphthacene are much more
sensitive than angular or condensed PAH like phenanthrene and chrysene.
Because light Intensity "decreases rapidly with depth In the water column,
rate of PAH photolysis also decreases with depth (Table 5). Sorption to
bottom sediments decreases photolysis rate further, particularly of higher
molecular weight PAH.
In natural marine and fresh waters, dissolved and particulate
substances In the water column may either inhibit or enhance the rate of PAH
photochemical transformation (Zadells and Simmons, 1983; Zepp et al., 1984).
Inhibition of photolysis may be due to light attenuation or the direct
inhibition of photochemical reactions in various ways.
Other chemicals may serve as sensitizers. Increasing the rate of
photochemical transformation of other chemicals. Dissolved humlc and
fulvic substances, algal cells, detritus, and even inorganic particles may
increase the rate of photolysis of some PAH under the Influence of sunlight,
presumably by acting as sensitizers. Humlc and fulvic adds from different
sources and different species of phytoplankton have Blmllar photosensitizing
abilities.
30
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XABLI 3. OOMPVTD HAU-LIVIS fOt 1U DIUCT PHOTOCHDdCAL TUBSrOUUTlOl
OF 8KVQAL P4B. Tha tlaa raqulradifor oat-half tha PAH aolaculat
to tatet oaar tha mrfaea of fraahvatar (alsulatad latltnd*, 40*1.
mid-day, mld-atamar), at S ¦ dapth, and at 5 ¦ partitioned te
bettea aadlaaat (Proa Sapp and Schlotihauar, 1979. with parmlaalon
•f tan Arbor Sclanca Publlahara)
—nsar-
CompMitf Siritei Jltettr* frenhkmO
Naptahttana
71
350
530
1-Mtthylnaphthatafw
»
110
190
2* Mtthxlnaphthftlm
*
410
MO
Phtnanthrtm
M
99
<9
Amhrtcarw
0.7)
U
O
9>M«thyUnthr»c«nt
041
0.79
U
9, J 0-Dlm«thy (anthracene
0J5
•
•
Pyr«n«
C.tt
U
3.9
FJuer anther*
21.0
1<0
200
ChryMM
M
13
tt
Naphthacene
0.0M
tUO
0.93
R«n»(a)pjrr«ne
0J4
M
19
Bcns(a)vithraccm
0J»
J.7
9.2
31
-------
At high latitudes, the rate of photolysis of organic chemicals is
greatly diminished due primarily to the reduced intensity and dally duration
of solar lrradlance during the winter (Figure 8). At 60°N latitude, there
Is an approximately 10-fold decrease In the rate of photolysis of
benzo(a)pyrene between June and December (Zepp and Baughman, 1978). Low
temperature in the winter, and certainly ice cover, would reduce the rate of
photochemical reactions further.
The situation apparently is quite different In an oil slick on the
surface of the water. Here the primary mechanism of photodegradatlon of PAH
Is photo-oxygenation Involving singlet oxygen (Larson et al., 1976, 1977).
Major photoreactlon products of PAH are peroxides, phenols, acids, and
carbonyls.
There are few direct observations of petroleum photo-oxygenation
following major oil spills. Several alkyl-substltuted dibenzothiophene
sulfoxides were identified In crude oil samples collected from the water
near the Amoco Cadiz spill (Calder et al., 1978; Patel et al., 1979).
Overton et al. (1980) layered crude oil from the IXTOC spill on
seawater in aquaria exposed to direct sunlight. After 4 days, a number of
benzoic acidB, substituted naphthenolc acids, phenanthroic, benzothlophenoic
and dlbenzothiophenoic acids were Identified in the extracts. There also
was evidence of side-chain oxidation of CI- and C2-substituted naphthalenes.
Recent laboratory and field studies have demonstrated that photo-
oxygenatlon reactions played an Important role in the weathering and
transformation of Ekoflsk crude oil as a surface slick on sea water (Tjessem
and Aaberg, 1983; Tjessem et al., 1983). The field studies were conducted
along the coast of Norway at 60°N latitude in June, and so resemble
conditions that might occur along the southern and western coast of Alaska.
However, the Beaufort Sea is at about 70°N and sea temperatures are much
lower and ice cover much more prolonged. The high molecular weight
resin/asphaltene fraction of the oil reacted rapidly with oxygen. Lower
molecular weight aromatics, Buch as phenanthrene, were converted to polar
oxidation products which tended to bind tightly, possibly covalently, to the
resin/asphaltene fraction. In related studies, Barth (1984) Investigated
photo-oxygenation of Ekoflsk crude oil In 1-m x 20-m deep plastic bags
deployed in the ocean and of Statfjord crude oil discharged at sea at 65°N
latitude in August. In the bag experiments, polar hydrocarbon fractions
dominated the hydrocarbons In the water column after the second day, and
their concentration continued to increase for about two weeks. The
composition of this polar fraction was not determined, but had properties
similar to those of aliphatic and aromatic alcohols, phenols, and carboxylic
adds. Similar polar fractions, presumably consisting of photo-oxygenation
products of oil hydrocarbons, were not observed in the water column under
the field spill. This was attributed to the ouch greater dilution of
soluble materials in the field and to overcast weather conditions during the
spill. The major mechanism of loss of low molecular weight aromatic
hydrocarbons (benzenes- naphthalenes) from spilled oil Is through
evaporation (Riley et al., 1980). However, formation of stable oil-water
emulsions, mousses, and pavements tends to slow evaporation of light
32
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l.M
l.»
1.11
1.00
0.87
o.so
o.»
0.11
0.00
MONTM OF TCM
Figure 8« Annual variation of the near-surface half-life
(*50) of benzo(a)pyTene at several northern latitudes.
(Proa Zepp and Baughman 1977).
33
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hydrocarbons (Payne et al., 1981a,b; 1983; Boehm and Fiest, 1982). Hardened
pavements of spilled oil on or burled In coastal sediments may retain
virtually unveathered oil Inside a crust of hardened weathered asphalt.
Such pavements may persist for years, providing a potential source of fresh
oil to the environment, should they be broken up by storms or other
disturbances.
TRANSFORMATION BT BACTERIA, FUNGI, AND ALGAE
Bacteria and fungi show tremendous diversity and adaptability in
utilization of different organic molecules as a sole carbon source. Some
microorganisms are able to oxidize some aromatic hydrocarbons completely to
carbon dioxide and water and use them as a source of energy and as a source
of carbon for biomass accretion.
Others are not able to carry out this complete oxidation of aromatics,
but if an alternative growth substrate is available, they are able to
metabolize them partially to various oxygenated metabolites. The
latter process Is called co-oxidation or co-metabolism (Gibson,
1977). Co-metabolism is thought to reflect a metabolic mistake by the
microorganism, due to incomplete enzyme specificity. In addition, certain
autotrophic bacteria, fungi, and algae can oxidize aromatic hydrocarbons
during photosynthesis, possibly deriving additional energy from the
heterotrophic process (Cerniglla, 1981).
An extremely important feature of the bacterial pathway is that the
cis-dihydrodiol apparently Is produced- through a dloxetane intermediate,
whereas in the mammalian microsomal system, a trans-dihydrodiol is
produced through an arene oxide intermediate. The arene oxides, or their
immediate oxidation products, appear to be responsible for the
carcinogenicity or mutagenicity of PAH shoving these properties
(Oesch, 1982). However, bacteria may be able to metabolize some PAH to
mutagenic or carcinogenic metabolites. Microbial methyl hydroxylation of
7,12-dimethylbenz(a)-anthracene to a carcinogenic product has been reported
(Wu and Wong, 1981).
In most cases, complete degradation of higher molecular weight PAH to
C02 and water proceeds very slowly or does not occur. Instead, various
phenolic or acidic metabolites are produced (Gibson, 1977).
Fungi (yeasts and molds), unlike bacteria, metabolize PAH by a
cytochrome P450- mixed function oxygenase system somewhat similar to that
found In mammalian liver microsomes (Cerniglla, 1981). As In the mammalian
microsomal system, the initial product of fungal attack on PAH is a
trans-dihydrodiol produced through an arene oxide Intermediate. Many
species of fungi from freshwater and marine environments are able to
metabolize naphthalene and other PAH, presumably by the arene oxlde-trans-
dihydrodiol pathway (Cerniglla, 1981). Abundance of fungi was three times
higher near (0.2 m) the Cape Simpson, Alaska oil seep than at locations 50 m
away (Barsdate, 1973).
34
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Several marine micro- and macro-algal species are able to metabolize
naphthalene and to a lesser extent higher molecular weight PAH (Cernlglla,
1961; Klrso et al., 1983; Van Baalen and Gibson, 1984). In some caseB, the
primary metabolic products of BaP degradation In the presence of algae are
qulnones, suggesting that the algal cells are serving as sensitizers for
photo-oxygenation of the PAH. In addition, phenol oxidases, which are
present In high concentration In some algae, could mediate the non-specific
enzymatic oxidation of PAH to qulnones. None of the diatoms from Cook Inlet
or from the ice edge in the Bering Sea, examined by Van Baalen and Gibson
(1982), could metabolize naphthalene completely to CO2. However, they could
metabolize it to 1-naphthol. The investigators suggested that In cold
marine environments, the rate of aromatic hydrocarbon metabolism by marine
phytoplankton might be a significant fraction of the rate of aromatic
hydrocarbon metabolism by marine bacteria.
Many studies have been performed on ability of marine water column and
sediment bacteria to degrade petroleum hydrocarbons, including PAH.
Aliphatic and aromatic hydrocarbons are degraded much more rapidly under
aerobic than anaerobic conditions (Lee et al., 1978; Vard et al., 1980).
Rate of PAH degradation tends to decrease with increasing PAH molecular
weight (Readman et al. , 1982). Bacterial populations from oil-
contaminated areas metabolize PAH more readily than populations from clean
areas (Saltzman, 1982).
Lee and Ryan (1983) reported that naphthalene and phenanthrene were
fairly rapidly blodegraded in water samples from the Rhode Island and
Georgia-South Carolina coasts. Medium molecular weight PAH (anthracene,
fluorene, benz(a)anthracene and chrysene) were not readily blodegraded in
the water column, but were degraded rapidly in sediment-water slurries.
Higher molecular weight PAH (dimethyl-benz(a)anthracene, dibenz(a,h)-
anthracene, and benzo(a)pyrene) -were not readily blodegraded even in
sediment-water slurries. When PAH were added to water or sediment
samples from areas with little or no hydrocarbon pollution, there was a
lag period of at least one day before significant degradation occurred.
The rate.of degradation of PAH was much more rapid in sediments from oil-
contaminated areas and there was no lag before initiation of degradation.
Benz(a)anthracene added to a sediment-water slurry from oil-contaminated
Shipyard Creek, South Carolina, in August (29°C) was degraded at a rate of
about 95 ng/g sediment/day, whereas the rate of degradation was about
14 ng/g sediment/day in a clean sediment-water slurry from the nearby
Skidway River, Georgia, at a temperature of 28°C.
Temperature had a profound effect on the rate of degradation of PAH in
water or sediments (Lee and Ryan, 1983). For example, the calculated
half-life for degradation of phenanthrene at an Initial concentration of 25
ug/liter in sea water was 79 days in June at a temperature of 18°C and
11,000 d in January at a temperature of 2°C. The half-life of
benz(a)anthracene at a concentration of 2.5 mg/kg in sediment Increased from
1,100 days In October at 15°C to 21,000 days In February at 4°C. The
authors also showed that dlmethylbenz(a)anthracene In water was degraded
slowly by a combination of photochemical and microbial processes.
35
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Photo-oxygenation reactions produced mono-hydroxy derivitives and quinones,
which then were degraded further to CO2 by resident bacteria.
Following a spill of crude or refined petroleum, different hydrocarbon
components and classes are degraded simultaneously, but at widely different
rates by indigenous water column and surflclal sediment mlcrobiota (Bartha
and Atlas, 1985). Low molecular weight n-alkanes In the C10-C22 molecular
weight range are metabolized most rapidly, followed by Iso-alkanes,
olefins (if present), mono-aromatics, PAH, and finally highly condensed
cycloalkanes, resins, and asphaltenes. The result is that microbial
degradation, plus evaporation and photo-oxygenation/photolysls markedly
change the composition of the oil over time.
The rate of microbial degradation of petroleum hydrocarbons in the
water column and in bottom sediments is dependent on many factors, including
oxygen and primary nutrients (especially nitrogen and phosphorus)
concentrations, /temperature, chemical composition of oil, and previous
history of hydrocarbon inputs to the area (Atlas, 1981; Bartha and Atlas,
1985). Several attempts have been made to estimate these rates in the
field. Butler et al. (1976) quote degradation rates of 1-10 mg
hydrocarbons/m3/day for open-ocean waters. Following the Amoco Cadiz oil
spill, in which 190,000 metric tons of crude oil was released to the coastal
waters of Brittany, France, Aminot (1981) estimated the rate of microbial
degradation of hydrocarbons in the water column under surface oil slicks to
be about 30 mg/ur/day. The rate of microbial degradation of dissolved and
mlcropartlculate oil:in the water column following the IXTOC _1 oil spill in
the Bay of Campeche, Gulf of Mexico, was in the range of 0.24 to 1,056 mg
aliphatic hydrocarbon respired/m3/d (Buckley and Pfaender, 1980; Pfaender
et al., 1980).
Rates of degradation of crude and refined oils in oxidized 6urficial
sediments are in the range of 0.02 to 0.40 grains/m2/day, with an average of
0.14 grams/m2/day (GESAMP, 1977). Atlas and Bronner (1981) estimated that
petroleum hydrocarbons were being degraded in the upper 5 cm of intertidal
sediments along the Brittany Coast of France Impacted by the Amoco Cadiz oil
spill at a rate of 0.05 grams/m2/day. Extrapolating to the 320 km of
coastline Impacted by the spill, they estimated that it would take more than
20 years to biodegrade the estimated 64,000 metric tons of oil believed to
have impacted the Brittany Coast.
In arctic marine environments, biodegradation of petroleum hydrocarbons
appears to proceed quite slowly, possibly due In part to nutrient limitation
(Bergstein and Vestal, 1978)» low numbers of hydrocarbon-degrading bacteria
(Atlas et al., 1981; Atlas, 1984), and low temperatures (Cundell and
Traxler, 1973; Glbbs et al., 1975). Atlas et al. (1978) simulated spills of
Prudhoe Bay crude oil on water during summer near Point Barrow, Alaska.
There was a 22 percent loss of oil during the first month, due primarily to
evaporation and/or dissolution of low molecular weight hydrocarbons. An
additional 3 percent was lost during the second month and was attributed
primarily to microbial activity. During winter, crude oil spilled on the
ice lost only 5 percent of its weight in nine months, due primarily
to evaporation, with no Indication of any biodegradation. Oil spilled under
36
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ice shoved a small transport of hydrocarbons during three weeks, due to
dissolution. Oil-contaminated sediments placed In 5 meters of vater lost
45 percent of hydrocarbons In 60 days during May and June. These losses
were attributed primarily to biodegradation.
In subsequent experiments performed in Elson Lagoon near Point Barrow,
Haines and Atlas (1982) followed degradation of hydrocarbons over a two-year
period. Total microbial biomass remained relatively constant in the
oil-contaminated sediments. After 1 to 4 months, numbers of
hydrocarbon-utilizing bacteria Increased slowly, but their numbers never
exceeded 4 X 10 ' percent of the total microbial biomass In the surficlal
sediments. Microbial degradation of the sediment hydrocarbons, as Indicated
by a decrease in the n-alkane/isoprenoid hydrocarbon ratio, became evident
only after 1.5 years of exposure. Concentrations of aromatic hydrocarbons in
the naphthalene, phenanthrene, and dlbenzothlophene series in sediments
decreased very slowly. The authors concluded that microbial degradation of
petroleum hydrocarbons in arctic marine environments of Alaska will proceed
very slowly. There also was little evidence of microbial degradation of
petroleum hydrocarbons In subtldal sediments at the site of the Baffin
Island Experimental Oil Spill (BIOS) (Boehm, 1983). It should be kept in
mind that the estimates quoted above for the rate of biodegradation of
petroleum hydrocarbons in the water column and in bottom sediments are for
total hydrocarbons. Since PAH, especially those above phenanthrene, are
among the more refractory components of petroleum, their rates of
biodegradation will be much lower than those for whole petroleum.
Atlas (1982) examined the species composition of water column and
sediment bacteria from coastal and outer continental shelf waters of all
coasts of Alaska. He reported that hydrocarbon-degrading bacteria were
indigenous to all Alaskan coastal waters and sediments. However, they were
present in low numbers, particularly in the water column. Atlas concluded
that the oil-degrading ability of these bacterial populations is not
sufficient to extensively degrade spilt oil before it Impacts sediments and
coastlines. He suggested that the rate of petroleum hdyrocarbon degradation
in Alaskan OCS regions would be slow, and hydrocarbon pollutants resulting
from accidents during OCS oil development and production would persist for
long periods.
TRANSFORMATION BY MARINE ANIMALS
In most Invertebrate and vertebrate animals studied to date, an enzyme
system variously known as the cytochrome P450-dependent mixed function
oxidase, mixed function oxygenase (MFO), aryl hydrocarbon hydroxylase, or
drug metabolizing system is responsible for Initiating the metabolism of
various lipophilic organic compounds, Including xenoblotlcs (foreign organic
compounds such as alkenes, PAH, pesticides, and drugs) and endogenous
compounds such as steroid hormones and bile salts.
The primary function of this system appears to be to render poorly
water-soluble lipophilic materials more water-soluble and therefore more
available for excretion. Although this system effectively detoxifies
certain xenoblotlcs, others, such as certain PAH and alkenes, may be
37
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transformed to intermediates which are highly toxic, mutagenic, or
carcinogenic to the host. Oxidative metabolism of PAH in this system
proceeds via highly electrophillc intermediate arene oxides, some of which
bind covalently to cellular macromolecules such as DNA, RNA, and protein.
It is now generally agreed that metabolic activation by the mixed function
oxygenase (MFO) system is a necessary prerequisite for PAH-lnduced
carcinogenesis and mutagenesis (Oesch, 1982).
The MFO system is located in the liver and usually also in several
other organs of many invertebrates and vertebrates. Within the cell, it Is
localized primarily in association with the smooth endoplasmic reticulum
(smooth microsomal fraction) (Staslecki et al., 1980).
Considerable progress has been made In identifying and quantifying the
MFO system and its components In marine invertebrates and vertebrates (See
reviews of Neff, 1979, 1984; Lee, 1981; Stegeman, 1981). Among the
Invertebrates, mixed function oxygenase activity seems to be highest in some
members of the phyla Arthropoda and Annelida. Of course, many Invertebrate
phyla have not been investigated. Coelenterates and ctenophores apparently
lack MFO activity. Among the Echinodermata, sea urchins Strongylocentrotus
sp. and starfish Asterias sp. have low levels of MFO.
Five species of marine polychaete worms have been examined and all but
one species (Arenicola sp.) have at least a limited ability to transform
PAH. The situation among the mollusca is more complicated. Several
investigators have been 'unable to detect MFO activity in several species of
bivalve molluscs including Mytilus edulls, Mya arenaria, Ostrea edulis, and
Anodonta sp. However, Anderson (1978) measured a low level of MFO activity
in oysters Crassostrea virginlca which have been exposed to BaP for long
periods of time. Stegeman (1981) detected traces of authentic MFO activity
(BaP hydroxylation>- in the marine mussels Mytilus edulis and Modiolus
modiolus. Low levels of MFO activity have been reported in snails Littorina
sp. and squid Illex illecebrosus (Payne and May, 1979).
Mixed function oxygenase activity has been detected in more than 16
species of marine crustaceans (Neff, 1979; Lee, 1981). Levels of In vitro
MFO activity in crustacean hepatopancreatlc preparations usually are low and
variable. In addition, non-hepatic tissues of crustaceans such as green
gland, gonads and stomach may have as high or higher MFO activity than the
hepatopancreas. Metabolite Identification experiments with several marine
crustaceans have shown that the Intact animal can produce PAH metabolites
very rapidly.
The mixed function oxygenase system has been detected in more than
eight species of elasmobranch fish and more than four dozen species of
teleost fish. Hepatic microsomal mixed function oxygenase activity varies
widely among different species and even among individuals of the same
species. These variations do not seem to be related to habitat preference
or feeding habits of the fish. There Is little difference in the range of
MFO activity between freshwater and marine species, or between cold water
and warm water species.
38
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In most cases, the types of PAH metabolites produced by marine
invertebrates and fish are similar to those produced by mammals. They
Include PAH dihydrodlols, phenols, and quinones as veil as PAH conjugates
with sulfate, glucuronic acid, and glutathione. Several species of fish
including starry flounder Platychthys stellatus, English sole Parophrys
vetulus, deep-sea ' rattail fish Coryphaenoldes armatus, and coho salmon
Oncorhynchus kisutch liver nricrosomes metabolize benzo(a)pyrene to a suite
of metabolites similar to those produced by rat liver microsomes (Varanasl
and Gmur, 1980; Stegeman, 1981; Gmur and Varanasl, 1982; Stegeman, 1983).
The major identified metabolites produced "by the fish microsomes are the
7,8- and 9,10-dlhydrodiols which are precursors of the major carcinogenic
BaP metabolite. In fish, as in mammals, most MFO activity, is localized in
the liver. Gills, kidney, and testes also contain significant MFO activity.
Activity in other tissues is low (less than 10% the activity In liver
microsomes). Different fish species may have markedly different basal
levels of MFO activity.
The pattern among marine invertebrates is quite different. In the blue
crab Callinectes sapidus and shore crab Carcinus maenas, highest BaP
hydroxylase activity occurs In the green gland or antennary gland
(analogous to the teleost kidney) and the pyloric stomach (Singer and Lee,
1977; O'Hara et al., 1982). The hepatopancreas (roughly analogous to the
teleost liver) has low activity. Other crab tissues containing significant
MFO activity include gills, testes, eyestalks, nerve ganglia, and heart.
Seasonal changes have been reported in MFO activity in tissues of
several species of fish and Invertebrates (Bend et al., 1978; Walton et al.,
1983). Sexual differences in levels and activities of MFO components have
been observed in fish and crustaceans. BaP hydroxylase activity was nearly
two orders of magnitude higher In green gland of mature female blue crabs
Callinectes sapidus than in green gland of mature males of the same species
(Singer and Lee, 1977).
Malins et al. (1982) determined the activity of the mixed function
oxygenase system in liver or viscera of six species of marine fish and two
species of marine invertebrates. All species were Indigenous to Alaskan
coastal waters. Lowest activity in fish liver was In arrowtooth flounder
(Atharestes stomia's), and highest activity was in Pacific cod (Gadus
mactocephalus). Activity In tanner crab (Chionoecetes sp.) was similar to
that In fish liver, while no activity could be detected in snail (Fusitriton
sp.) tissues. Mixed function oxygenase activity In these Alaskan species
was similar to that in similar species from lower latitudes.
Enzyme Induction Is the process whereby concentration or activity of an
enzyme Is increased in response to exposure of the organism to the normal
substrate of the enzyme. Several investigators have shown that exposure
to PAH or complex mixtures of hydrocarbons, such as petroleum, results in
rapid (within 1-10' days) Induction of elevated levels of MFO activity In
some marine fish (See reviews of Lee, 1981; Stegeman, 1981). Different
components of the PAH-metabolizing system are induced to different degrees
by inducers. Therefore, it is not surprising that the relative proportion
39
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of different types of PAH metabolites produced by uninduced and induced
hepatic microsomes are different.
Much less research has been performed on induction of MFO activity in
marine invertebrates. Several attempts to induce MFO activity in several
species of algae* molluscs, crustaceans, annelids* and echlnoderms by
exposure to PAH or oil have failed. However, Anderson (1978) apparently was
able to Induce a very lov level of BaP hydroxylase activity In oysters
Crassostrea virginlca following exposure for 5-9 months to 1 ug/1
benzo(a)pyrene or methylcholanthrene in the ambient medium. Lee et al.
(1979) Induced MFO activity In the polychaete worm Capitella capitata by
exposing the worm to benz (a)anthracene' or crude oil In its sand substrate
for three to ten weeks. Lee et al. (1981) demonstrated induction of MFO
activity in sand worms Nereis vlrens by .exposure to benzo(a)pyrene or PCBs
In the laboratory or oil-contaminated sediments in the field. Exposure of
copepods Calanus helgolandicus for 7 d to seawater solutions of naphthalene,
2-methylnaphthalene, 3-methylcholanthrene, or benzo(a)pyrene (50-200
ug/1) resulted in significant elevation of BaP hydroxylase activity in whole
animal homogenates (Walters et al., 1979).
Salmonids have a well-developed readily inducible hepatic mixed
function oxygenase system. Gruger et al. (1977a) reported that mixed
function oxygenase was Induced in female coho salmon (Oncorhynchus kisutch)
by exposure for six days to the water-soluble fraction of oil at
concentrations as low as 15 ppb. Males were less sensitive. Subsequently,
Gruger et al. (1977b) fed saltwater-adapted coho salmon (0. kisutch) and
chlnook salmon (0. tshawytscha) test diets containing polychlorlnated
blphenyls, petroleum hydrocarbons, or both. There was substantial induction
of mixed function oxygenase In liver microsomes of coho, but not Chinook
salmon.
Several investigators have demonstrated Induction of MFO activity in
fish as a result of exposure in the field to petroleum. Payne (1976)
reported significantly higher levels of MFO activity among cunners
Tautogolabrus adspersus from oil-polluted stations than from unpolluted
stations along the coast of the Avalon Peninsula, Newfoundland, Canada.
Based on these and other results, he recommended use of fish hepatic
microsomal MFO activity as a monitor for oil pollution.
Davles et al. (1981) reported elevated levels of MFO activity in
hepatic microsomes of several species of fish from the vicinity of offshore
oil production platforms in the North Sea. Spies et al. (1982) studied
activity of MFO enzymes in two species of flounder, Cetharlchthys sordidus
and C. stlgmaeus from coastal southern California. Highest hepatic MFO
activity was observed in flatfish from the vicinity of the Los Angeles
County sewage outfall off Palos Verdes, Intermediate near the 7-mlle
Hyperion sewage outfall In Santa Monica Bay and near an oil seep in the
Santa Barbara Channel, and lowest in unpolluted waters of Monterey Bay.
Payne et al. (1984) reported elevated MFO activity in kidney microsomes of
winter flounder Pseudopleuronectes amerlcanus from the vicinity of an oil
spill in the Bale Verte, Newfoundland.
40
-------
It is apparent that a great many endogenous and exogenous factors can
Influence levels of MFO and related enzyme activity and concentrations
of MFO components in the tissues of aquatic anlmalB. Each species seems to
respond somewhat differently to these factors and there are even significant
intraspeclfic (probably genetic) differences. Endogenous factors may include
age, sex, reproductive status, nutritional status, period of the molt cycle
(in arthropods), etc. Primary exogenous factors influencing rate of PAH
metabolism include temperature, season, possibly salinity (though this has
not been studied), and current and previous history of exposure to
potential Inducers or inhibitors of different components of the microsomal
PAH-metabolizlng system.
Although Induced hepatic MFO activity is a useful and fairly specific
index of previous exposure to certain organic pollutants, its utility for
assessing environmental damage is limited by lack of a complete
understanding of the long-term biological consequences to the fish of
elevated MFO activity. MFO Induction 16 a normal adaptive response of the
animal and is not by, itself an indication of existing or Impending
pathology. Chronic induction may afford a degree of protection from the
inducing pollutant. Alternatively, it may produce pathological side
effects by exceeding the capacity of the conjugation system, by altering
metabolism of endogenous substrates such as steroid hormones, or by
increasing the production of mutagenic and carcinogenic metabolites.
INTERACTION OF DEGRADATIVE PROCESSES IN THE REMOVAL OF PAH FROM THE
MARINE ENVIRONMENT
In the marine environment, the assorted oxidative processes that are
responsible for degrading PAH do not proceed independently of one another.
The nature and extent of interactions between different processes are often
hard to ascertain. Lee et al. (1978) introduced dispersed Prudhoe Bay
crude oil enriched with a number of PAH into a marine controlled ecosystem
enclosure (ca. 2-m diameter and 15-m deep) on the coast of British Columbia,
Canada, and estimated rate of loss of different PAH by evaporation,
photochemical oxidation, microbial degradation, and sedimentation.
Concentrations of added PAH decreased by 50 percent in 3 to 4 days at a
depth of three meters and in 4 to 6 days at seven meters. By Day 17, all
PAH were below the limit of detection at seven meters. Photolysis appeared
to be the main mechanism of BaP degradation in the upper 5 meters. Qulnones
and more polar materials were the major products of BaP photodegradation.
Nearly 40 percent of the BaP introduced into the enclosure was recovered
from bottom sediments. BaP and benz(a)anthracene were not metabolized
significantly by water-column bacteria. However, naphthalene,
methylnaphthalenes and anthracene were metabolized to f a small extent by
water column bacteria. Naphthalene and methylnaphthalenes, and to a much
lesser extent anthracene and fluoranthene, were also lost from the water
column by volatilization. A significant fraction of all added PAH
sedlmented to the bottom of the enclosure adsorbed to colloidal or detrital
material. Accumulation and metabolism of PAH by water column organisms,
mainly phytoplankton and copepods, contributed little to the degradation of
PAH in the system.
41
-------
This study Illustrates the brief residence time of PAH in temperate
marine waters. Lower molecular weight aromatlcs (benzene-phenanthrene) are
removed primarily by evaporation and microbial activity. Higher molecular
weight PAH are removed mainly by sedimentation and photochemical
transformation. Degradation by water column animals Is of minor Importance.
However, as Lee et al. (1979) showed, benthic animals may contribute
significantly to PAH degradation by sediment bloturbation. This keeps
sediments aerobic so that aerobic sediment bacteria can metabolize the PAH.
Hinga et al. (1980) showed in another microcosm study that most of the
microbial degradation of PAH (in this case, benz(a)anthracene) takes place
in the upper 1-2 cm of sediment (the oxidized layer) and not in the water
column or deeper sediment layers. Once buried In anoxic sediments, PAH
appear to be extremely persistent.
ALASKAN PERSPECTIVE
The mechanisms of degradation and loss of polycyclic aromatic
hydrocarbons from Alaskan marine waters and sediments are the same as those
occurring in marine waters and sediments of more temperate and tropical
climates. These processes include evaporation, photochemical and chemical
degradation, and metabolism by marine bacteria, fungi, phytoplankton, and
animals. In Alaskan waters, most of these processes proceed slowly because
of low ambient temperatures and low net incident solar radiation during much
of the year. In particular, the most important mechanisms quantitatively of
PAH removal (I.e., evaporation, photodegradation, and microbial metabolism)
appear to proceed very slowly in arctic and subarctic marine waters. Thus,
PAH Introduced into Alaskan waters by natural mechanisms. Intentional
discharges or accidental spills will tend to persist and may accumulate over
time to high concentrations in Alaskan marine sediments.
42
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5. ACCUMULATION AND RELEASE BY MARINE ANIMALS
It is Important to critically evaluate the published literature on the
rates of accumulation and release of petroleum-derived aromatic hydrocarbons
by marine organisms. The accumulation and release rates are important
factors In the prediction or assessment of the long-term Impacts of oil
spills and chronic discharges on the marine environment and, in particular,
on commercial fisheries resources in the spill or discharge area.
BIOACCUMULATION FROM WATER AND SEDIMENT
There is an extensive body of published literature dealing with the
bioaccumulation of petroleum hydrocarbons, including low to medium molecular
weight aromatics, by marine animals during exposure for different lengths of
time to petroleum hydrocarbons accommodated or dispersed in the water
column. There have been fewer published reports of bioaccumulation of
petroleum hydrocarbons from oil-contaminated sediments. The hydrocarbons
most frequently quantified in exposure media and animal tissues were
naphthalenes and phenanthrenes.
Table 6 summarizes results of several laboratory studies of
bioaccumulation of aromatic hydrocarbons by marine bivalve molluscs,
crustaceans, polychaetes, and fish, many of them indigenous to Alaskan
coastal waters. Although bioconcentratlon factors (concentration in tissues
divided by concentration in exposure medium) cover a wide range, some
general trends emerge. For the few species for which comparative data are
available, bioaccumulation factors for aromatic hydrocarbons from sediments
are much lower than those from water. Bioaccumulation factors tend to
increase with increasing molecular weight, and thus decreasing aqueous
solubility, of polycycllc aromatic hydrocarbons.
Bivalve molluscs, which typically demonstrate the highest
bioconcentratlon factors of the taxa studied, may achieve bioconcentratlon
factors of 100 to 36,000. The higher bioconcentratlon factors usually are
associated with long-tens exposures to low concentrations of hydrocarbons in
water. When bivalves were exposed to relatively high concentrations of
petroleum hydrocarbons in sediments, the bioconcentratlon factors for
specific aromatic hydrocarbons were quite low (0.1-10). Roesljadi et al.
(1978) showed in a carefully designed laboratory study, that
bioconcentratlon factors in the clam Macoma inquinata for four PAH were 0.2
or less from sediment and 10 to 1,349 from the overlying water column.
43
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TABLE 6. SUMMARY OF AROMATIC HYDROCARBON UPTAKE BY MARINE ORGANISMS
Taxon
Species
Exposure Conditions
Oil/Hydrocarbon
Water
Max.
Cone.
(ppm)
Time
idagsj.
Tissue
Concentration (ppm)
Cone.
Factor
Reference
Bivalves
Macoma
balthlca
Macoma
Inqulnata
Hytllus
edulls
Rangl a
cuneata
Rangla
cuneata
Rangla
cuneata
Rangl a
cuneata
Prudhoe Bay Crude (PBC) 0.001
(total aromatlcs)
(OTO)
Prudhoe Bay
Phenanthrene 0.004
Chrysene 0.0004
Dlmethylbenz(a)- 0.0006
anthracene
Benzo(a)pyrene 0.0004
North Sea 0.030
crude oil 0.036
Produced water 1.7
(total naphthalenes)
Naphthalene 0.071
Phananthrene 0.089
Chrysene 0.066
Benzo(a)Pyrene 0.052
Total naphthalenes 3.4
Benzo(a)pyrene 0.030
0.030
Naphthalene 0.84
Methylnaphthalenes 0.4
Dimethyl naphthalenes 0.24
Trlmethylnaphthalenes 0.03
120
15.7
140
28
100
0.04
0.30
0.86
0.04
152.0
78.0
16.0
0.43
2.85
0.54
0.45
60.0
7.2
5.7
1.9
3.5
4.1
0.8
15,700
10
694
1349
861
5,000
2,200
9.4
6.1
32.0
8.2
8.7
17.6
240
190
2.3
8.7
17.1
26.7
Clement et al. 1980
Roesljadl et al. 1978
Nlddows et al. 1982
Fuclk et al. 1977
Neff et al. 1976b
Anderson A Neff 1975
Neff & Anderson 1975
Neff et al. 1976b
44
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TABLE 6. SUMMARY OF AROMATIC HYDROCARBON UPTAKE BY MARINE ORGANISMS (CONTINUED)
Exposure Conditions
Tissue
Taxon
Bivalves
Species
Crassostrea
vi rginica
Crassostrea
virginica
Crassostrea
vi rginica
Oi1/Hydrocarbon
Water (con't.)
Naphthalene
Methylnaphtha1enes
Dimethylnaphtha1enes
Anthracene
Fluoranthene
Benzo(a)thracene
Benzo(a)pyrene
Total hydrocarbons
Max.
Cone.
JasnL
0.005
0.003
0.008
0.003
0.010
0.002
0.013
0.001
0.007
0.004
0.005
0.0001
0.002
0.0001
0.106
Time Cone.
(days) Concentration (ppni) Factor
2 30.0 6,000
8 12.0 4,000
2 56.0 7,000
8 36.0 12,000
2 84.0 8,400
8 72.0 36,000
2 S.6 431
8 2.S 2,500
2 5.0 714
8 4.0 10,000
2 2.8 560
8 1.8 18,000
2 0.36 180
8 0.30 3,000
50 330 3,100
Reference
Lee et al. 1978
Stegeman & Teal 1973
Sediment
Bivalves
Hacewa
inquinata
Hacowa
i nqui nata
Prudhoe Bay
(Total aromatics)
PBC
Phenanthrene
Chrysene
Benzo(a)pyrene
6.0
6.0
2.1-.006
3.3-.01
0.6-.03
54
55
59
58
60
2.0
5.2
0.018
1.3
.15
0.3
0.9
3-6
1-11
1-5
Augenfeld et al. 1980
Roesijadi & Anderson
1979
Augenfeld et al. 1982
45
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TABLE 6. SUMMARY OF AROMATIC HYDROCARBON UPTAKE BY MARINE ORGANISMS (CONTINUED)
Exposure Conditions
Tissue
Taxon
Bivalves
Species
Macowa
inquinata
Maccroa
Inquinata
Macowa
inquinata
Oi 1/Hydrocarbon
Sediment (con't.)
Phenanthrene
Chrysene
Dimethylbenz(a)-
anthracene
Benzo(a)pyrene
PBC
(total aromatics)
PBC + dispersant
(total aromatics)
Max.
Cone.
(ppm)
0.49
8.37
4.63
0.64
<>0
40
Time
Idaysj.
7
7
7
7
30
30
Concentration (ppm)
.10
.31
.30
.06
5.0
5.0
Cone.
Factor
.20
.04
.06
.09
0.13
Reference
RoesiJadi et al. 1978
Anderson et al. 1983
0.13 Anderson et al. 1985
Protothaca
staminea
Protothaca
staminea
Prudhoe Bay
(total aromatics)
PBC ~ dispersant
(total aromatics)
22
40
365
30
0.6
1.9
0.03
0.05 Anderson et at. 1985
Crustaceans
Penaeus
aztecus
Penaeus
aztecus
Water
No. 2 fuel oil,
WSF
(total naphthalenes)
No. 2 fuel oil
(total naphthalenes)
1.9
0.3
0.7
4.0
digestive gland 70
3
2.1 Anderson et al. 1974b
233 Neff et al. 1976b
4.3 Anderson 1975
Penaeus
aztecus
No. 2 fuel oil, WSF
(total naphthalenes)
0.4
46
0.5
18
-------
TABLE 6. SUMMARY OF AROMATIC HYDROCARBON UPTAKE BY MARINE ORGANISMS (CONTINUED)
Taxon
Crustaceans
Species
Exposure Conditions
Tissue
Peneeus
aitecus
(Postlarval)
011/Hydrocarbon
Water
No. 2 fuel oil, NSF
(total naphthalenes)
Max.
Cone.
JjEEEL
0.85
Time
0.2
Concentration (ppw)
55
25
Cone.
Factor
Reference
65 Anderson 1977
29
Palaewonetes
pug to
Penaeus
seti ferus
Hlppolyte
clarhl1
Pandalus
danae
Uca mlnax
No. 2 fuel oil, NSF 0.1
(total naphthalenes)
No. 2 fuel oil, 0.3
Field experiment 0.3
(total napahthalenes)
Prudhoe Bay, NSF 1.0
(naphthalenes and 0.6
phenanthrenes) 0.4
Prudhoe Bay, NSF 0.8
(naphthalenes and 0.5
phenanthrenes) 0.3
No. 2 fuel oil, 0.3
Field experiment
(total naphthalenes)
0.2
1
2
3
0.3
1
3
60
19
18
30
30
20
10
5
22
20 Anderson 1977
200 Cox et al. 1975
63
18
50
75
25
20
17
74
Anderson et al. 1980
Cox et al. 1975
Sesarma
clnereum
Anonyx
1atIcoxae
No. 2 fuel oil,
Field experiment
(total naphthalenes)
Prudhoe Bay, NSF
(total naphthalenes)
0.3
0.120
0.160
0.031
0.022
1
1
4
7-15
32
10
10
2
25
106 Cox et al. 1975
83
62
64
1136
Anderson et al. 1979
47
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TABLE 6. SlFOiftRV OF AROMATIC HYDROCARBON UPTAKE BY MARIKE ORGANISMS (CONTINUED)
Exposure Conditions Tissue
Max.
Cone. Time Cone.
Taxon Species Oil /Hydrocarbon (ppw) (days) Concentration (ptwi) Factor Reference
Sediment
Anonyx 2 4 8 4 Anderson et al. 1979
1at!coxae
Mater
Polychaetes Neanthes
arenaceodentata
No. 2 fuel oil, NSF
(total naphthalenes)
0.6
12
20 Rossi & Anderson 1977a
Sediment
(total naphthalenes)
10
28
<0.1
0 Rossi 1977
Food
Hethylnaphtha1ene)
0.05
16
<0.1
Sediment
Abarenlcola
paclflea
Prudhoe Bay
Phenanthrene
Chrysene
Benzo(a)pyrene
5.1
0.2
0.08
60
58
60
0.28
1.1
0.17
1-4
1-6
1-6
Augenfeld et al. 1982
Nater
Fish
Fundulus
simllus
No. 2 fuel oil, NSF
(total naphthalenes)
1.9
0.2 Liver 4,481 2,358
Gall Bladder 66,704 35,107
Heart 9,714 5,113
Brain 2,387 1,256
Dixit & Anderson 1977
48
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TABLE 6. SUMMARY OF AROMATIC HYDROCARBON UPTAKE BY MARINE ORGANISMS (CONTINUED)
Exposure Conditions
Tissue
Taxon
Fish
Species
Salmo
qalrdnerl
Cllllchthya
ml rebl 11s
011/Hydrocarbon
Water (con't.)
^C-naphtha 1 one
l*»C-2 methyl-
naphthalene
^C-Benzo(a)pyrene
***C-naphtha1ene
Sediment
Max.
Cone.
JjEEEL
0.005
0.005
0.023
0.5
0.005
0.5
0.001
29
Time Cone.
(days) Concentration (ppm) Factor
1 Bile fluid 2.1 415
0.3 Total body 0.3 60
0.3 Total body 2.9 126
1 Bile 36 71
Bile 13 2,566
1 Bile 109 218
4 Liver 0.01 10
Flesh 0.002 2
Heart 0.03 30
Call bladder 7.4 7,400
0.04 Liver 2200 76
Flesh 9100 314
Heart 2500 86
Call Bladder 4,200 145
Reference
Melancon & Lech 1978
Lee et al. 1972
Parophyrs
vetulus
Alaskan North
Slope crude oil
(aromatlcs)
27
51
Liver 4
Liver 0.18
1.3
0.1
McCain et al. 1978
49
-------
Bioconcentration factors for PAH In marine crustaceans exposed via the
water usually fall in the range of 2 to* 200. When the amphipod Anonyx
laticoxae was exposed to a water-soluble fraction of Prudhoe Bay crude oil
for 7 to 15 days, the bioconcentration factor for total naphthalenes reached
1,136 (Anderson et al., 1979). When' exposed to oil-contaminated sediments
for four days, the same species reached a maximum bioconcentration factor of
only 4.
Polychaetes (Neanthes arenaceodentata) accumulated naphthalenes to a
concentration 20 times that in the ambient medium when exposed to a
water-soluble fraction of a light crude oil (Rossi and Anderson, 1977a).
When fed oil-contaminated food or exposed to oil-contaminated sediments, the
same species failed to accumulate significant concentrations of naphthalenes
(Rossi, 1977). On the other hand, the deposit-feeding polychaete,
Abarenicola paclfica, accumulated three PAH from sediments to
bioconcentration factors of 1 to 6 during exposure to the sediments for up
to 60 days (Augenfeld et al., 1982).
A majority of bloaccumulation studies of PAH in marine fish have been
performed during brief exposures to the hydrocarbons in the water. Under
these conditions, fish typically show bioconcentration factors for PAH in
different tissues in the range of about 100 to 1,000. The gall bladder, a
major route of PAH excretion in fish, may exhibit a transitory
bioconcentration factor of up to about 35, 000, reflecting rapid excretion
of accumulated hydrocarbons. If the duration of exposure Is extended beyond
a few days, bioconcentration factors for PAH in fish drop precipitously, due
to Induction of the mixed function oxygenase system, leading to rapid
metabolism and excretion of PAH. However, high concentrations of PAH
metabolites may remain bound to and undetected In the tissues (Mallns
et al., 1982). As a result of this process, marine fish from sites heavily
Impacted by an oil spill may contain low concentrations of unmetabolized PAH
in their tissues (Neff and Haensly, 1982).
Accumulation and release of PAH by marine animals from Alaska exhibit
patterns similar to those .described for marine animals generally (Rice
et al., 1984). Figure 9 summarizes results of experiments in which four
species of Alaskan marine animals were exposed to the water-soluble fraction
of Cook Inlet crude oil for 96 hours. The crustaceans accumulated the
highest concentrations of aromatic hydrocarbons, followed in order by pink
salmon fry and scallops. The pink salmon fry began depurating aromatic
hydrocarbons, probably by metabolic degradation, with 24 hours of the start
of ^exposure. The other species depurated the accumulated hydrocarbons
rapidly when returned to clean seawater. The scallops depurated most
slowly. Three-hour bioconcentration factors In pink salmon fry for aromatic
hydrocarbons Increased with Increasing molecular weight, in the series from
toluene to trlmethylnaphthalene (Table 7).
Juvenile king crabs, Paralithodes camtschatlca, were able to accumulate
aromatic hydrocarbons from both water and sediments contaminated with Cook
inlet crude oil (Rice et al., 1983). They accumulated a maximum of 7 and
3000 ppm total aromatics in muscle and hepatopancreas, respectively, during
exposure for 28 days to a water-soluble fraction of oil containing 0.5 ppm
50
-------
12
Depuration
KING CRAB
Depur
Btion
SHRIMP
—r
/•
PINK SALMON
Depurationv. -.N
Depuration
SCALLOPS
i i
96 144
TIME IN HOURS
240
Figuze 9. Concentrations of aromatic hydrocarbons in pink shrimp, pink
scallops, pink salmon fry, and king crab during exposure to
the water-soluble fraction of Cook Inlet crude oil and during
subsequent depuration. Points on the graph represent the sum
of all aromatic hydrocarbon concentrations found in groups of
five individuals determined at various time intervals (from
Rice et al., 1984).
51
-------
TABLE 7. THE EFFECT OF MOLECULAR WEIGHT ON THE
ACCUMULATION OF AROMATIC HYDROCARBONS BY
PINK SALMON. THE BIOCONCENTRATION FACTOR IS
THE RATIO OF TISSUE CONCENTRATION TO WATER
CONCENTRATION AFTER 3 H EXPOSURE TO THE WSF
OF COOK INLET CRUDE OIL (From Rice et al.,
1984).
Bioconcentratlon
Oil Component Factor
Toluene
3
Naphthalene
64
Methylnaphthalene
139
Dime thylnaphthalene
198
Trimethylnaphthalene
380
52
-------
total aromatlcs. Maximum concentrations were reached in muscle after one
day and in hepatopancreas after four days. During exposure to sediments
containing two percent Cook Inlet crude oil for 90 days, the juvenile crabs
accumulated up to 3 ppm and 370 ppm total aromatlcs in muscle and
hepatopancreas, respectively. Maximum accumulations were reached after 10
and 90 days, respectively.
DISTRIBUTION OF ACCUMULATED PAH IN ANIMAL TISSUES
Studies of PAH bloaccumulation in bivalve molluscs usually have not
Included determinations of the distribution of accumulated hydrocarbons in
the tissues. The visceral mass, including the digestive gland, constitutes
most of the body mass, and most consumers of bivalves, including man, ingest
the entire soft tissues. . However, Neff and Anderson (1975) demonstrated
that when the marsh clam Rangia cuneata was exposed to ^C-labelled
benzo(a)pyrene (BaP) In the water, approximately 50 percent of the
radioactivity accumulated was in the viscera. Mantle tissue contained the
second highest concentration, and the remainder of the BaP was distributed
evenly among the remaining tissues. More recently, Ulddows et al. (1982)
reported that the concentration of accumulated aromatic hydrocarbons in
Mytllus edulls was approximately five times higher in the digestive gland
than in other tissues.
Figure 10A shows the distribution of total naphthalenes in body parts
and organs of the brown shrimp Penaeus aztecus during and following exposure
to the water-soluble fraction of No. 2 fuel oil. The digestive gland
accumulated naphthalenes to the highest concentration during exposure and
retained them longest during depuration. Similar results were obtained n
spot shrimp, Pandalus platyceros, an Alaskan species (Sanborn and Malins,
1980). During exposure to a water-soluble fraction of Prudhoe Bay crude
oil, thorasic segments, including digestive gland accumulated about four
times more aromatic hydrocarbons than abdominal segments. Similarly, in
purple shore crabs, Hemigrapsus nudus, maximum bioconcentratlon factors for
naphthalene ranged from 4.2 In hemolymph to 104 in digestive gland (Rice
et al., 1984).
In fish distribution of accumulated PAH seems to be somewhat more
uniform (Figure 10B). Greatest accumulation is in gall bladder and
digestive tract, major routes of PAH excretion. Brain, probably because it
Is high In lipids, also tends to accumulate high concentrations of PAH.
This may have important Implications relative to the mode of toxic action of
PAH. Research in Malins' laboratory with marine fish from the Pacific
Northwest and Alaska (e.g., Roubal et al., 1971; Collier et al., 1978;
Varanasl et al., 1981) also revealed significant bloaccumulation of PAH in
liver gut and brain. Whereas, much of the PAH In liver was present as
metabolites, most of that in brain tissue was unmetabolized.
RELEASE OF PAH
A majority of the depuration studies have been conducted In very clean
seawater, not environments where levels of contamination slowly decrease.
53
-------
PICW 10. TISSUI OISTKIBUTIOM OF MFBTHAUHIS.
• au» iiiin
• IIIHII
¦ IIU
o fiosmcrea
A IlltltlVt Hilt
Atltll
Me
•i .
• i
MO
WD
TW( ODSUI
Figure 10A. Fenaeus attecus. Accumulation and retention of total naphthalraea
by different body regions of Juvenile blown ahrlnp exposed to J01 dilution of
water-soluble fraction of So. ? fuel ell. Body regions of 3 ehrlop were pooled
fortaeh data point, and background values were subtracted. (Troa KefI et al.
1976b).
e—« mi
«•••« lirll
•—-* mu
»—» Mkl
»-—« Mill
# a
•—4 until
ea
se
• s se is
SIMMM
it n itc
tIMC INOVMI
Figure 10B. Fundulua elttllls. Distribution of total Daphtbalanea In various
tlasues of Culf kllllflah during exposure to water-aoluble fraction of lo. t
fuel oil and at different tlaea following exposure. Water-aoluble fraction
contained approxlnately 2 pp» total naphthalenea. (Ttob Raff et al.( 1976b).
54
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In the natural environment, the reservoir of contaminants in sediments may
continue to provide low levels of bioavallable petroleum components, even
after the sources of Input have been removed.
Data from various studies on the depuration of hydrocarbons In marine
organisms were fit to a hyperbolic kinetic model. To compare experiments
using different exposure concentrations, the data were normalized to percent
of the original hydrocarbon in the animals at time 0 (i.e., before
depuration began). The data were then transformed to linear form and
regression statistics calculated. Thus:
P ¦ t/b + at or 1/P ¦ l> + a
t
where t » time (days)
a ¦ constant
b ¦ coefficient
P ° percent remaining
This hyperbolic model has a demonstrated physical basis in equilibrium
modeling both in adsorption chemistry as the Langmulr equation (Browman and
Chesters, 1977) and in the Michaelis-Menten model (Segjel, 1976). It has
been used to model the bioaccumulation and depuration of organic compounds
from synthetic fuel residuals (Hardy et al., 1985a) and the flux of metals
through water surfaces (Hardy et al., 1985b). The time at which 50% of the
original hydrocarbon remains is estimated by substituting 50 for P in the
above equation and solving for t. Care must be taken in any attempt to
estimate the percent remaining at longer periods since this model does not
fit well the slow rates of release occurring in bivalves after about 10 days
of depuration (Neff and Anderson, 1981).
A major distinction must be made between saturates or n-paraffins and
aromatic hydrocarbons. Saturates have in general been found to produce
substantially less toxicity for marine species than aromatic hydrocarbons
(Neff, 1979). At similar molecular weights, saturates are less soluble in
seawater than aromatlcs and their uptake is frequently associated with the
intake (through filtering) of oil droplets. Other Important factors to
consider are the relative partitioning of different hydrocarbons from lipid
pools in the tissues back to seawater and the rates of metabolic conversion
of aromatic hydrocarbons to degradation products within different marine
species.
Table 8 shows that the half-life (t50) of total hydrocarbons and
naphthalenes in bivalves is on the order of 5 days. The faster depuration
shown for n-parafflns is likely the result of the removal of oil particles
(droplets) associated with the external portions of membranes as compared
to soluble components (aromatlcs) which have entered the tissue. Published
curves for depuration and tables of tissue content consistently show a rapid
decline in bivalve body burdens shortly after transfer to clean water (Neff,
1979; Neff and Anderson, 1981). This rapid 50% loss is followed by a slower
55
-------
TABLE 8. SUMMARY OF DEPURATION KINETICS FOR PETROLEUM HYDROCARBONS
Hydrocarbons
(organisms)
Coefficient
(b)
Constant
(a)
n
t50%
References
n-paraf fIns
(bivalves)
total hydrocarbons
(bivalves)
total naphthalenes
(bivalves)
10141
phenanthrene
(bivalves)
BAP
(bivalves)
BAP II
(bivalves)
naphthalenes
(shrimp, amphipods
and annelids)
naphthalenes
(fish)
0.117
0.030
41.29
(xlO- )
0.83
0.025
0.241
5.177
1
2
3
n =• number of data points,
r = regression coefficient.
t50 ° half-life in days.
¦7264
-0.339
0.048
-7.74
(xlO- )
-3.61
-0.016
0.056
-3.341
12 0.98 0.72 Clark & Finley 1975
Neff et al. 1976b
4 0.96 3.5 Stegeman & Teal 1973
32 0.93 4.8 Anderson 1973
Neff et al. 1976b
Cox et al. 1975
Anderson & Neff 1976
Fucik et al. 1977
Anderson 1977
11 0.98 0.4 Neff et al. 1976a
Neff et al. 1976b
7 0.89 4.5 Neff & Anderson 1975
5 0.95 3.2 Neff et al. 1976a
Lee et al. 1978
29 0.89 0.3 Cox et al. 1975
Anderson & Neff 1976
Anderson 1977
Anderson et al. 1979
Anderson et al. 1974b
Rossi & Anderson 1977a
9 0.99 0.7 Anderson & Neff 1976
Anderson et al. 1974b
-------
release, often requiring up to 50 day6 for complete depuration of petroleum
components. The slow release phase Is generally explained by a gradual
partitioning of the sequestered hydrocarbons from a more stable pool to gill
surfaces and release by seawater flushing. A portion of this slow release
nay be associated with a very low rate of organic hydrocarbon metabolism
(aryl hydrocarbon hydroxalase) reported for bivalves. Regardless of the
mechanisms, clearly bivalves reach some of the highest bioaccumulation
levels for petroleum components and release this material more slowly than
most other marine species.
From Table 8 it appears that there are no major differences In the
rates of release of petroleum hydrocarbons exhibited by worms (annelids),
shrimp, amphlpods and fish (tso " 0.3 - 0.7). Since these species possess
greater capabilities for enzymatic degradation of petroleum components than
bivalves (Section 4 of this report), the final portion of the depuration
curves generally exhibit a steeper slope than in bivalves. It Is possible
to compare the tso values In Table 8 to specific studies conducted in
Alaska. Koto and Rice (1981) measured the uptake and depuration rates of
naphthalene in early life stages of Coho salmon. Figure 11 shows the uptake
by eggs, alevlns and fry reached maximal values at different times but
approximately the same concentration range (50-100 times exposure
concentration). Depuration rates increased with the age of the life stage
from egg to fry. The tso value of 0.7 days shown in Table 8 for Fundulus
slmills is very near the tso value for Coho salmon fry (Figure 11). An
approximate time to complete removal of tissue contamination by adult
polychaetes, shrimp and fish might be 5 to 7 days. Therefore, the time for
complete depuration of PAH from bivalves is about an order of magnitude
greater than the time required by many other marine species.
When petroleum contamination Impacts a coastal zone, hydrocarbon
concentrations do not merely reach a peak and sharply decrease to
nondetectable levels. For this reason the experimental depuration studies
reviewed above may represent the best case for removal of PAH contamination
from tissues. In most spill situations and in regions receiving multiple or
constant oil Inputs, sediments generally become contaminated. Leaching of
hydrocarbons from the substrate produces low-level water contamination
(Including pore water) which may be1' reflected in an apparent slower
depuration rate by infaunal, epifaunal, and demersal species.
57
-------
« I • 10 U
DAYS SINCE START OF EXPOSURE
K.r;-H-.x!TAUV,K
« • 19 13
DAYS SINCE END OF EXPOSURE
Figure 11. Accumulation and depuration of naphthalene In early
life stages of Coho salmon. Fish vere exposed to 100 ng/L
naphthalene then placed In flowing seavater for depuration.
(Tron Korn and Rice 1981).
58
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FACTORS CONTROLLING THE UPTAKE AND RELEASE OF PAH
The parent compound naphthalene (2 aromatic rings) and alkylbenzenes
are generally not retained long on surface sediments due to relatively high
water solubility (Roesijadi et al., 1978). Studies conducted with sediments
or detritus generally have shown that naphthalenes are not accumulated by
organisms exposed to oiled substrates (Rossi, 1977; Roesijadi et al., 1978;
Anderson et al., 1978). With increasing molecular weight of aromatics,
there is also an Increase in the net accumulation within detrltivores such
as the clam, Macoma inqulnata (Table 6). When this species was exposed to
oiled sediments containing relatively high concentrations of those
^C-labeled PAH, bloaccumulatlon was quite low. Time of exposure of benthlc
species to contaminated sediments and the molecular weight of the specific
hydrocarbon measured are probably the most Important factors regulating the
final level of tissue accumulation.
There is increasing evidence indicates that persistence of specific
hydrocarbons in sediments, in the absence of new inputs, is dependent upon
the interaction of chemical, physical' and biological factors. Lower
molecular weight aromatics (naphthalene and smaller) tend to be both
degraded (by sediment-associated microorganisms) and leached (water and
sediment movement) more rapidly than higher molecular weight components. As
these compounds are released from surficlal sediments they may be taken up
by benthlc species, where similar degradatlve pathways and seawater flushing
gradually serve to decrease body burdens. In temperate waters, this cycle
of events may take place rather rapidly, so that animals in the natural
environment exposed to single inputs of lower molecular weight compounds may
not show significant tissue contamination after one to two months of
sediment exposure. Long-term field experiments with contaminated sediments
have been used to verify these assumptions. To our knowledge this approach
has not been applied in Alaskan coastal environments. Just the opposite
pattern in bloaccumulatlon is expected if a constant discharge of petroleum
components is taking place. Naphthalene and alkylnaphthalenes are among the
compounds that are bloconcentrated to the greatest extent in tissues of most
marine species (Table 6) exposed to soluble hydrocarbons.
Examination of data summarized in Table 6 on alkylnaphthalenes,
phenanthrenes, chrysene and benzo(a)pyrene lead to the conclusion that these
PAH behave quite differently, but predictably, when organisms are exposed to
contaminated sediments. For example, when the reservoir (total
concentration) in sediment is sufficiently high (5-10 ppm), benthlc species
may accumulate alkylnaphthalenes and phenanthrenes from sediment in periods
of one to two months (Roesijadi et al., 1978; Augenfeld et al., 1982).
Surficlal sediments release the majority of these components to the water
during this period, but lower layers of contaminated material may continue
releasing significant amounts, which are available for infauna. These
intermediate molecular weight compounds are also available from solution in
constant discharges, but their concentrations are generally quite low. As
In the case of naphthalene, these PAH can be accumulated in organisms to
sufficiently high concentrations to produce mortality and sublethal
effects (Anderson et al., 1980). Because of the slower release rates from
59
-------
sediments (compared to the parent compound), alkylnaphthalenes and
phenanthrenes can also exert a significant sublethal effect from
contaminated sediments.
Chrysene and benzo(a)pyrene and other petroleum components In the same
molecular weight range have very low water solubility and have not been
shown to either reach high concentrations in discharges or to produce
significant mortality for marine species (Rossi and Neff, 1978). Uhile
there Is potential for producing carcinogenic or mutagenic responses In
marine species, clear evidence of this type of Impact still is
circumstantial. Varanasl et al. (1982a, b) and Varanasl and Gmur (1981)
provide evidence of the manner by which English sole metabolize B(a)P and
the occurrence of covalent binding of metabolites to macromolecules. While
such a pathway may well lead to Impacts on populations, Including cancer,
cause and effect relationships In the environment are difficult to document.
Primary concern regarding high molecular weight PAH has been the
contamination of marine organisms exposed to polluted sediments and transfer
to higher organisms in the food web. Since the 4- and 5-ring aromatlce
partition very slowly from sediments to Interstitial water and the overlying
water column, sediments represent a relatively stable reservoir for the
bioaccumulation by Infaunal species. Low solubility makes uptake from
seawater (discharges) unlikely, but organisms in close contact with
sediments may continuously accumulate the low levels of these components
slowly released from particles to the interstitial water. The rate of
uptake by organisms is therefore very slow but continuous, and periods of 30
to 60 days may not be long enough to accurately assess the magnitude of
bioaccumulation (Augenfeld et al., 1982). Foodchaln transfer of
contamination to predators Is possible, but for reasons discussed In
Section 4, the turnover of PAH in predators Is such that biomagnificatlon
has not been shown.
In summary, rapid uptake, particularly from water, characterizes the
movement of naphthalene from the environment to marine species and toxic
effects may be produced. Alkylnaphthalenes and phenanthrenes occupy an
Intermediate molecular weight range, such that uptake and effects occur
over a longer period from sediment exposures. If a marine discharge
produces levels In the receiving water between 0.2 and 1.0 ppm of these
compounds, acute toxicity is possible. A more likely event is the chronic
sublethal effect on infauna from - long-term exposure to contaminated
sediments. Higher molecular weight compounds, including chrysene and
benzo(a)pyrene are less likely to produce acutely toxic effects on the
infauna, but gradual tissue contamination Is possible.
RELATIONSHIPS BETWEEN TISSUE CONTAMINATION BY PAH AND EFFECTS
There are relatively few studies describing the level of PAH
contamination associated with acute or sublethal effects on marine species.
A primary reason for our lack of information in this area is that sensitive
species are generally used in toxicity tests. These often small and fragile
animals may not withstand longer exposure times and if they do, the
small amount of tissue may not lend -itself to detailed chemical analyses.
60
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Bivalves, which are the best species- for demonstrating bloaccumulation, are
generally not the most sensitive organisms for use in assessing toxicity.
During flowing exposure tests with either a seawater extract of
Prudhoe Bay crude (PBC) oil or chemically dispersed PBC, Anderson et al.
(1980; 1984; 1985) have used the shrimp (Pandalus danae) to determine the
relationship between toxicity and body burdens. With different exposure
systems (WSF or dispersed oil) and varying periods of exposure to different
concentrations, they have found that a tissue level of 8 to 13 ppm total
naphthalenes and phenanthrenes correlates with 50Z mortality in Pandalus
danae. Stekoll et al. (1980) described the sublethal effects of chronic oil
exposure on the clam, Macoma balthica and coworkers (Clement et al., 1980)
characterized the extent of bloaccumulation. Clams exposed to 0.03 ppm
total hydrocarbons for 60 to 120 days contained aromatic hydrocarbons
between 12 and 16 ppm. Tissue concentration factors increased with
Increasing substitution of the parent aromatic ring structure for
naphthalenes, dlbenzothlophenes, anthracenes and chrysenes. At the 60-day
exposure to 0.03 ppm, the clams exhibited a negative energy balance, which
the authors conclude would lead to population decreases. Widdows et al.
(1982) reported on an extensive study of the sublethal effects of oil
exposure on the mussel, Mytilus edulis. The study demonstrated the
correlation between tissue content of aromatic hydrocarbons and the
magnitude of suppression In energy balance (scope for growth). At tissue
levels above 7 ppm aromatic hydrocarbons, there was either zero or negative
scope for growth and higher concentrations produced greater reductions in
energy balance.
Roesljadi and Anderson (1979), Augenfeld et al. (1980) and Anderson
et al. (1983) have shown that when the bivalves Protothaca staminea and
Macoma inquinata contain levels of PAH at 1 to 4 ppm, sublethal effects are
demonstrated. Since these studies were all,conducted for one month or more
in contaminated sediments, it is likely that animals contained higher levels
of at least naphthalenes during earlier periods of exposure. The Impacts
observed were probably the result of both short-term high concentrations of
naphthalenes and longer exposures to the higher molecular weight compounds.
ALASKAN PERSPECTIVE
Alaskan marine animals, like those from temperate and tropical
climates, readily accumulate PAH and related hydrocarbons from water.
Bloconcentration factors (concentration In tissue/concentration in medium)
tend to be higher for bivalve molluscs than for other taxa of marine
animals. Accumulation of PAH from sediment is less efficient than from the
water. However, for medium molecular weight aromatics, alkyl naphthalenes,
phenanthrenes and dlbenzothlophenes, sediments may be a major source of
chronic contamination of benthic and demersal specieB.
Low temperatures, characteristic of Alaskan coastal waters, do not seem
greatly to affect the rate of accumulation and release of hydrocarbons by
marine animals. Rate of PAH bloaccumulation may be Increased In some
species such as bivalve molluscs and decreased in others such as
fish by low water temperatures. In addition, low water temperatures tend to
61
-------
decrease the rate of metabolism and excretion of PAH In those species which
depend heavily on the mixed function oxygenase system for PAH metabolism
(Rice et al., 1984). This slower rate of depuration plus the greater
persistence of PAH in low temperature marine environments, may mean that
the potential for chronic impacts of PAH pollution of the Alaskan marine
environment is greater than for more temperate and tropical climates.
62
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6. BIOLOGICAL EFFECTS OF PAH
Over the past 5 years, extensive research has been performed world-wide
to determine the potential effects of petroleum hydrocarbons on species,
populations and communities of marine animals. Recently, a review of Oil in
the Sea was published by the National Academy of Sciences (NAS, 1985). Both
short-term (acute) toxicity tests, designed to estimate the concentration
causing 50 percent mortality in one to a few days (LCfeo)» and short- or
long-term tests designed to detect other responses than mortality, have been
performed with oil or specific PAH and a wide variety of marine organisms.
Collectively, investigations concerning a biological response other than
death have been referred to as sublethal studies and Include the detection
of abnormal biochemical, physiological, histological, developmental,
reproductive or behavioral responses in a single species, at PAH
concentrations not producing significant mortality. This section will
attempt to summarize both acute and sublethal responses, with emphasis on
studies conducted with Alaskan species. It should be noted, however, that
the majority of studies dealing with the acute toxicity or sublethal effects
of oil and specific PAH in marine organisms have been conducted in the lower
48 states. All of these findings relate to Alaskan species since the basic
physiology of organisms and chemical-physical alterations of oil in water
are generically similar throughout the marine environment. It is fortunate
that the Auke Bay National Marine Fisheries Service Laboratory has recently
produced a review of their oil-effects research (Rice et al., 1984). The
majority of this type of research in Alaska has been conducted by these
investigators, and their findings are included in this report.
TOXICITY OF SPECIFIC COMPOUNDS
The fact that oils, and especially their water-soluble fraction (WSF),
vary in the proportion of monoaromatlcs and higher molecular weight
diaromatics and trlaromatlcs, led researchers to investigate the toxicity of
specific petroleum to a variety of marine organisms. Figure 12 illustrates
the concentrations of specific aromatic hydrocarbons required to produce 50%
mortality in 96 h (96 h LC50) or 24 h (24 h LC50) for a range of organisms.
The general trend Is an increase in acute toxicity (lower LC50 value) as
molecular weight increases to the highest molecular weight compound
(fluoranthrene) found to produce acute toxicity. Unfortunately, there is
no complete data set for any one species. The shrimp, Palaemonetes
pugio, was tested with the low to medium range of molecular weight
aromatlcs (benzene- phenanthrene), while the polychaete, Neanthes
arenaceodentata, was tested with PAH of medium to high molecular weight
[naphthalene - dlbenz(a)anthracene]. Data for other species are mostly for
63
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Gbftcar
Solmenfry
Cyprinetfon
Copcpod
Pslotmorwrti
AmpMpod
8trlp«d boit
Panotut oittcut
CfOM thrimp
24 hour tXgo
/ / / // / it /
/
L
Figure 12. Concentrations of specific petroleum compounds producing 50%
mortality for selected marine organisms (Data from Caldwell
et al., 1977; Benville andKorn, 1977; Neff et al., 1976a;
Thomas and Rice, 1979; Young, 1977; Rossi and Neff, 1976;
Ott et al., 1978; and Lee and Nicol, 1978).
64
-------
various Intermediate molecular weight compounds. Therefore, individual
species may not necessarily follow the general trend.
The inverse relationship between molecular weight and acute toxicity of
aromatlcs in the series from benzene to fluoranthene may be attributed in
part to the fact that the lower molecular weight aromatice are volatile and
are lost rapidly from bloassay media, decreasing their apparent toxicity.
In addition, as Hutchinson et al. (19B0) have shown, the degree of membrane
perturbation, and thus the toxicity of an aromatic hydrocarbon, will be
proportional to the concentration of the hydrocarbon associated with the
membrane surface. This* in turn, will be proportional to activity (product
of mole fraction and activity coefficient) of the aromatic hydrocarbon in
the aqueous phase. Since the activity coefficient is inversely proportional
to solubility, less soluble (higher molecular weight) aromatlcs should have
a greater effect at a given concentration than more soluble ones.
TOXICITY OF OIL EXTRACTS
It is important to understand the Interaction of oil and seawater in
producing an extract of soluble petroleum components and to determine the
rates these compounds are dissipated from both oil and oil-and-water phases.
In the laboratory, water-soluble fractions (WSF) of oil have been made by
slowly mixing one part oil over nine parts seawater for 20 hours (Anderson
et al., 1974a). Table 9 shows the relative contribution of different
aromatlcs to several oil extracts (WSF). The percentage of n-paraffins in
the WSF was a very low (1-2%) since they are less soluble than the
monoaromatlcs, which represented about 90% of the WSF. The remaining 4 to
6% is composed of naphthalenes, which are probably the most toxic
components. Total concentrations of aromatlcs in these WSF (10% mixture;
oil over water) were from about 6 to 20 ppm, and similar values were
obtained when a flowing exposure system was used (Anderson et al., 1980).
Weathering of Prudhoe Bay crude oil in the open air over slowly flowing
seawater for 24 days resulted in 100% loss of monoaromatlcs, and
naphthalenes were reduced by about 20 to 60%, based on the molecular weight
and structure of the specific compound (Riley et al., 1980).
Table 10 demonstrates the major difference between the WSF of the very
toxic Mo. 2 Fuel oil and the WSFs of two crude oils. This table (from Rice
et al., 1964) also shows the similarity between treated ballast water and
the WSF from crude oils. Naphthalenes are the major components of No. 2
Fuel oil WSF and are 15 to 2d times higher than naphthalenes in water
equilibrated with crude oil. These authors have also shown that
temperatures between 5 and 15°C have a considerable influence on the rate of
loss of aromatic hydrocarbons from test solutions during 96 h. Low
temperatures substantially decrease the rate of volatilization of light
aromatlcs from the WSF.
Table 11 from the review of Malins (1977) summarizes the data from
numerous toxicity studies. The ranges shown are quite large. Often,
results are based on either no chemical analyses or rather poor analytical
data. The ranges probably present a reasonable estimate of the
concentrations of oil extracts or suspensions of oil required to produce 50%
65
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TABLE 9. THE CONCENTRATIONS OF C,2-CjH N-PARAFFINS ANO AROMATIC HYOROCARBONS IN REFERENCE OILS AND THE 10* MATER-
SOLUBLE FRACTIONS (WSFsj PREPARED FROM THEH. Hydrocarbon concentrations In the oils are given as percent
(g/100 ml) and In the WSFs as parts per billion (ug/L) In Z0°/oo seawater (from Anderson et al., 1974a,b;
and Anderson et al., 1980).
i. Louisiana
Kuwait
12 Fuel Oil
Rufflce* C
Prod hoe Bay
Raae of
Whole
Whole
Whole
Whole
Whole
WSF'
Compound
Oil
VSF
Oil
WSF
Oil
WSF
Oil
WSF
Oil
(X)
(ppb)
(X)
(ppb)
(X)
(ppb)
(X)
(ppb)
(X)
(ppb)
n-paraffins
Cm
0.44
10
0.46
<0.5
0.82
5.0
0.11
0.8
0.36
t
Cib
0.48
10
0.41
<0.5
1.06
7.0
0.11
0.9
0.37
C|(
0.54
12
0.43
0.6
1.20
8.0
0.15
1.2
0.36
Cn
0.41
9
0.42
0.8
0.98
6.0
0.12
1.9
0.34
Cii
0.30
7
0.28
0.5
0.60
4.0
0.10
1.0
0.30
Total C|}*C}|
n-paraffins
3.98
89
4.00
2.9
7.38
47
1.26
12
4.40
2.93
aroawtlcs
benzene
3
6,750
3
3,360
3
S50
a
40
3
36.1
toluene
4,130
3,620
1,040
80
0.93
55.5
alkylbenzenes4
760
730
970
110
3.19
53.8
naphthalene
0.04
120
0.04
20
0.40
840
0.10
210
0.16
1.6
1-ecthylnaphthalene
0.08
60
0.05
20
0.82
340
0.28
190
0.21,
9 M
2-nethylfMphthatene
0.09
50
0.07
0
1.09
400
0.47
200
0.25'
Z. 9
dioelhy(naphthalenes4
0.36
60
0.20
20
3.11
240
1.23
200
0.60
1.4
triacthylnaphthalenes4
0.27
8
0.19
3
1.84
30
0.88
100
0.48
0.6
biphenyls4
<0.01
2
<0.01
1
0.16
28
<0.01
1
3
3
tluorenes4
0.02
3
<0.01
3
0.36
20
0.24
11
3
3
phenanthrenes4
0.0G
4
0.04
3
0.53
20
1.11
23
0.39
0.3
di benzoth iophenes4
0.0?
1
0.01
1
0.07
4
<0.01
1
0.134
3
Total aroMtlcs
0.94
11,948
0.60
7,709
9.18
4,562
4.31
1,166
6.54
151.7
Total hydrocarbons
•easured
4.92
12,037
4.60
7,792
16.56
4,609
5.57
1,178
10.94
154.6
Total hydrocarbons
present (IR analysis)
19,800
10,400
8,700
6,300
141
1 = 30°/Oo Seaweter extract (not 10X) In a flowing system.
f = Individual h-paraffin conccntr.ittons were very low.
3 = Mot measured.
4 = Total of several isomers.
66
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TABLE 10. COMPOSITION OF EXTRACTS PROH CRUDE AND REPINED OILS AND TANKER
BALLAST TREATMENT EFFLUENT (FROM RICE ET. AL., 1984).
Concentration (ppw)
Component
Prudboe
Bay1
WSF
Cook .
Inlet
WSF
Ballast*
effluent
Fuel
011
USF
Benzene
1.8
3.2
3.2
0.11
Toluene
2.0
2.S
2.2
0.17
o-xylene
0.28
0.35
0.32
0.12
m- and £-xylene
0.58
0.78
0.68
0.17
Naphthalene-
0.084
0.15
0.098
0.15
1-Methylnaphthalene
0.032
0.066
0.049
0.13
2-Methylnaphtha! ene
0.048
0.088
0.066
0.25
Ratio of all monoaromatlc
hydrocarbons to
dlaromatlc hydrocarbons
19.3:1
15.7:1
23.6:1
1.1:1
? From Rice et al. 1981.
From Rice et al. 1979.
67
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TABLE 11. ACUTE TOXICITY OF PETROLEUM TO MARINE ANIMALS (FROM MALINS, 1977)
Organism Material Tested Lethal Concentration
PPa
Finfish Soluble hydrocarbons 5-50
Larvae and eggs 0.1-1.0
Pelagic Crustacea 1-10
Benthlc crustacea 1-10
Gastropods 10-100
Bivalves 5-500
Other benthic 1-10
invertebrates
Finfish No. 2 fuel oil/kerosine 550
Larvae and eggs 0.1-4.0
Pelagic crustacea 5-50
Benthlc crustacea 5-50
Bivalves 30,000-40,000
Other benthic 5-50
Invertebrates
Finfish
Larvae and eggs
Pelagic crustacea
Benthic crustacea
Gastropods
Bivalves
Other benthic
invertebrates
Fresh crude oil
88-18,000
0.1-100
100-40,000
56
?
1,000-100,000
100-6,100
Finfish
Gasoline
Diesel fuel
91
240-420
Finfish
Larvae and eggs
Pelagic crustacea
Waste oil
1,700
1—>25
15->50
Finfish
Residual oils
2,000-10,000
68
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mortality after 96 h (4 days) of exposure. Using Alaskan species and Cook
Inlet crude and No. 2 Fuel oil, Rice et al. (1979) found that 96-h LC50
values for fish, shrimp and crab were from about 1 to 4 ppm total aromatic
hydrocarbons. All species were more sensitive to the WSF of No. 2 Fuel oil
than the crude oil WSF (Table 12).
Data from studies in Alaska and elsewhere on comparisons between No. 2
Fuel oil and crude oil WSF combined with several studies reviewed by Neff
(1979) have led to the conclusion that naphthalenes are very significant
contributors to the toxicity of oil extracts. Total naphthalenes include
the parent compound naphthalene and analogues with one to four methyl groups
attached to the diaromatic ring. Some oils, such as Kuwait, only produce a
maximum of 0.07 ppm naphthalenes in the 100% WSF (1:10. oil:water), but
South Louisiana and Prudhoe Bay crudes produce maximum naphthalenes
concentrations of 0.30 ppm in the WSF. No. 2 Fuel and Bunker C oils produce
higher total naphthalenes values (0.90 - 1.93) in the 100% WSF. Table 13
shows the results of toxicity testing with two shrimp species and two
polychaetes in Texas (22°C). Toxicity values (LC5q) were found to be
between 1.7 and 0.3 ppm total naphthalenes, regardless of the type of oil
producing the test solution (Anderson et al., 1974a; Rossi et al., 1976).
The range of toxic concentrations was relatively small (about five times)
for the species tested with four different oils and the values were
equivalent to the 100% WSF. Since these concentrations represent a type of
equilibrium for a relatively high volume of oil over water, these levels
might be present in a spill situation only near the point of oil-seawater
mixing.
Also summarized in Table 13 are values for eight species of marine
animals (mostly crustaceans) tested in Alaska (4-8°C) which exhibited 96 h
LC5q at concentrations of about 0.1 to 0.5 ppm total naphthalenes. There is
an overlap in the LC50 values for all the tests, but somewhat lower values
were consistently derived from tests in Alaska at 4 to 8cC, compared with
those produced at 22°C. There are two possible explanations for these
differences. Organisms living in Alaska may actually be more sensitive.
The more likely explanation is that the petroleum components, particularly
those of low molecular weight, remained in the exposure solutions for longer
periods of time at the lower temperatures. The correct explanation for
these differences can be obtained by exposliig the organisms to constant
concentrations in a flowing system over the same 4-day period.
The findings presented in this section lead to the conclusion that
naphthalenes are, perhaps, the most toxic components in the various oils,
but that a portion of the toxicity Is contributed by some of the
monoaromatlc compounds present in initially high concentrations In extracts
of crude oil. Other Important contributors to the toxicity of some crude
and refined petroleums are the phenanthrenes and dibenzothiophenes.
TOXICITY OF OIL TO DIFFERENT LIFE STAGES
Figure 13 shows the differences in the tolerances of four species of
invertebrates at different life stages to precisely the same extract of
No. 2 Fuel oil under the same exposure conditions. For Palaemonetes puglo,
69
-------
TABLE 12. ACUTE TOXICITY OF WATER-SOLUBLE FRACTIONS (WSF)
OF A CRUDE AND REFINED OIL TO ARCTIC MARINE
SPECIES FROM FOUR PHYLA*
Avg. 96-h TLm (ppm) of WSF of:
Organism Crude Oil No. 2 Fuel Oil
Echinodermata
Sea Cucumber (2 8pp.)
10.8
2.3
Mollusca
10 species
9.8
2.3
Arthropoda
Crabs (4 spp.)
Shrimp (5 epp.)
2.3
2.3
Amphipod, Isopod, Mysid,
and Barnacle (1 sp. ea.)
8.5
4.1
Chordata
Fish (5 spp.)
1.8
1.8
*Data taken from Rice et al. 1977 and 1979.
70
-------
TABLE 13. CONCENTRATIONS OF NAPHTHALENES IN SPECIFIC OILS PRODUCING 50« MORTALITY FOR SEVERAL
MARINE INVERTEBRATES IN 96 HR.
FST
t*s*
Kiwalt
11 %
So. Bunker #2
Louisiana C fuel 011
' 1.7 1.09
4 1.2 1.5
0.32 0.97 0.04
0.31 0.24 1.51
• (0.16)
• 0.17
0.40
(0.41)
Cook
Inltt
JflL
PUfllp
NlMtUt*
HUtW
WganUtM*
IWBKUfcllUte
CffiltjIU*
cumti
fwdilw1
no n I unit
Crawoon*
«lo»kinli
hnHtMn*
cwUcMtlei
NntolM*
hroilnolM
fandalw*
bortilli
Coalus*
M>Helt
Ss!aaaa!s8
nudui
0.28
0.52
(0.14)
(0.20)
0.11
(0.11)
0.05
0.22
(0.20)
(0.12)
0.30
(0.09)
0.11
(0.16)
0.51
(0.09)
(0.20)
(0.13)
(0.13)
(0.10)
(0.1S) (0.17)
22C
22C
I.HC
3.S-M
3.5-OC
).HC
3.5-OC
3.S-8C
3.S-0C
71
-------
ft)
£
N
O
as
e
o
»-
CB
c
o
x>
ll
a
w
o
I-
•o
X.
a
o
4J
X
Pu
P-
0)
3
rH
<0
>
o
m
u
<-}
u
9
0
se
1
vC
a*
is
]
BROWN S*BMP
(PCMUUS
AZTECUS)
WHITE SHRIMP
I PC tutus
KTirCRUS)
CMSS SHRIMP
IfALACMONniS
muoi
POLTCHAET*
INEANTHES
ARCNACEOOCNTATA)
Figure 13. Concentrations of No. 2 Fuel oil WSF producing 502 mortality
in 96-h exposures to various life stages of crustaceans and
polychaetes (from Neff etal., 1976a and Rossi and Anderson,
1976).
72
-------
there is a decrease in sensitivity with increase in size and age. However,
both Penaeus aztectus and Neanthes arenaceodentata adults are considerably
more sensitive to the petroleum extracts than younger stages and sensitivity
increases with increasing size.
Few other publications compare the sensitivity of organisms at various
life stages (including adults) under exactly the same exposure conditions.
Caldwell et al. (1977) and Mecklenburg et al. (1977) discuss the
sensitivities of various larval stages of decapod crustaceans. There are
particularly sensitive stages in the development of many decapods, including
crabs and lobsters (Wells and Sprague, 1976). One should compare the
responses of these larval stages to those of the adult large crustaceans in
looking at the range of sensitivities. In doing so we often find that the
parents are extremely tolerant to oil and are usually resistant to even the
highest exposure concentration tested over 4 days. Apparently the large
size and relatively impermeable carapace of lobsters and crabs have aided
them greatly in their protection from petroleum compounds.
Karinen and Rice (1974) reported Impaired molting by crustaceans
exposed to crude oil. Mecklenburg et al. (1977) compared animals in various
stages of molt. This work and various other observations during toxicity
testing, have shown crustaceans are more sensitive at the molt stage than at
intermolt periods. Mecklenburg et al. (1977) found that coonstrlpe shrimp
larvae (Pandalus hypslnotus) were approximately eight times more sensitive
during molting than during intermolt. There are obviously mechanical and
physiological stresses to overcome during molting; also, there is an uptake
of the surrounding water which would bring high doses of available
contaminants.
In their recent review, Rice et al. (1984) summarized the numerous
studies conducted on eggs and larvae of salmon. They noted that eggs and
alevlns are more tolerant to short-term exposures (96 h) of petroleum
hydrocarbons than fry or juvenile salmon. While eggs of Coho and pink
salmon are quite resistant to benzene in 96-h tests (LCso°340-540 ppm), the
fry of these species exhibited 96-h LC50 values of 10 to IS ppm (Moles
et al., 1979). Longer-term exposures of Coho salmon eggs . to toluene or
naphthalene produced significant reduction in hatching at exposure levels
which were less than 1% of the 96-h LQo values for eggs (Korn and Rice,
1981). It should be noted that eggs and fry of most salmon are found in
freshwater, but pink salmon often deposit their eggs in regions where
temperature and salinity can modify the sensitivity of early life stages of
this species. Korn et al. (1979) have found pink salmon fry more sensitive
at lower temperatures and there is a general trend of Increased sensitivity
of salmon fry and smolt with Increasing salinity (Moles et al., 1979;
Stickle et al., 1982).
Studies at Auke Bay on embryos, larvae and adults of kelp shrimp
(Eualus Buckleyi) and coonstrlpe shrimp (Pandalus hypslnotus) showed that
eggs were more tolerant to petroleum hydrocarbons than the adult females.
Even during long-term exposures (28 d) in a flowing system, the eggs of
shrimp survived if the female carrying them survived. Toxicity values for
eggs, larvae and adults (96 h and 28 d LC50) were between 0.5 and >1.4 ppm
73
-------
total aromatics. These concentrations are very near the 96-h LCjq shown for
white shrimp (Penaeus setiferus) in Figure 13. In other Alaskan studies
(Brodersen et al., 1977), Stage I larvae of king crab (Parallthodes
camtschatlca) exhibited 96-h LC50 values that were about one-half of the
values for juveniles. They also reported an Increase of about one order of
magnitude (0.2 - 2.0 ppm) in the tolerance of coonstrlpe shrimp between
Stage 1 and Stage VI larvae.
ACUTE TOXICITY FROM FLOW-THROUGH TESTS
The above findings Involved static tests primarily with water-soluble
extracts of oil. An extract from a 1:10 (oil:water) mixture produces much
the same component distribution as at least one flowing exposure system
(Anderson et al., 1980). The tremendous advantage of a flowing system is
that component composition does not vary over time of exposure, relative to
the volatility of the specific compounds. Thus, flow-through systems
provide consistency of exposure, so. -that either sublethal or lethal
responses of organisms can be closely 'related to specific components and
concentrations. Vanderhorst et al. (1976) Indicate that flowing exposures
may reduce 96-h LC50 values to about one-half of static LC50 values.
Standard static acute teBts are also difficult to interpret because
time Is rather short (4 days), and time of exposure is not allowed to vary.
At least in mortality tests of relatively short (9 days or less) duration,
Anderson et al. (1980) found that the time of exposure is as important as
concentration. In this study, a 50% mortality of three crustacean species
(Pandalus danae, Hippolyte clarkii and Seomysis awatschensls) was predicted
by extrapolation of a log-log plot of exposure time (up to 9 days) and
concentration of exposure. There was a consistent total hydrocarbon
exposure (time x concentration) in ppm-days which produced 50% mortality for
these crustacean species. This relationship has been confirmed using
Pandalus danae and chemically dispersed oil in constant and diluting
exposures (Anderson et al., 1981). More recent data (Anderson, 1984) on
Pandalus danae also demonstrate a predictable total exposure producing 50%
mortality, regardless of the rate of exposure. Obvious limitations to this
approach are very short exposures, which do not provide sufficient time for
death (a few hours) or very long exposures to low concentrations (over
9 days), which may allow some animals to accommodate to the petroleum
components. Flow-through studies ensure that concentration and composition
of the exposures are consistent so that the tolerance of a species can be
accurately defined. Impacts on species under realistic spill conditions can
be predicted from a knowledge of dilution rates, movement of the organisms
and alterations of the hydrocarbons under specific field conditions.
SUBLETHAL EFFECTS
Volfe (1977) summarized much of the available Information on the fate
and effects of petroleum hydrocarbons In marine ecosystems with considerable
emphasis on Arctic and subarctic environments. In this symposium, Anderson
(1977) reviewed studies on physiological, histological, behavioral and
reproductive effects of different oil components of several crude and
refined oils on marine species (Table 14). While total hydrocarbon values
74
-------
TABU 1». HMWRY Or ETflCTJ Of PCTOOUIH WOROCAIBON} ON DC GROWTH
anp KtPwoucTiow or maim oiiiws inw anoobon. 19771
Cafoture
Cencentretlen(ppa)'
Crowth or ((production
Soeclea
Oil
(day*)
IM
IN
TA
Pereeeter
Reference*
ri$N
«
Fundul m *ienjj«
so.u.r
to
16.6
0.2
9.7
«% hattoof egg*.
Aederaon et al., 1976
Cvorlnodon
i
varleaatu*
no.: ro
7
2.0
0.6
1.7
0% hatch of ogga.
Anderaon et el., 1976
Oncorhvnchu*
Prudtio*
oofbutch*
Cruda
10
0.7
Reduced great* rata of pink
Rice et al., 197S
ealaon fry.
OCCAPOOS
C*ncer eaal*ter
Aletkan
Cruda
60
0.2
Reduced survival of toeae on
Caldwell et al., 1977
longterai eapoaure.
Rlthrosanooeti*
herratll
No.i ro
27
1.0
0.1
0.9
R*dwc*d aurvlval and eatended
Naff et al., 1976a
davalopaant to aegalopa.
PalaaaMfietet
ouolo
No.2 ro
12
0.9
0.}
o.e
Reduced great* rata of larvaa.
Tataa, 1977
J
1.*
0.6
Reduced viability of aggt from
eipcied gravid faa*le*.
ParalIthoda*
Cook
caot*ch*t
-------
found to produce abnormal responses varied from 0.1 to 16 ppm, there was
closer agreement for both total naphthalenes and total aromatlcs. A
relatively small range of total naphthalenes, 0.1 to 0.6 ppm, adversely
affected the growth and reproduction of polychaetes, crustaceans and fish
over exposure periods of 1 to 60 days.
Shaw and co-workers (Clement et al., 1980; Stekoll et al., 1980)
exposed the clam, Macoma balthica. to a dispersion of Prudhoe Bay crude oil
for six months. At the highest concentration (3.0 ppm) of total
hydrocarbons. 50% mortality occurred after about 130 days. At the 120-day
Interval, clams exposed to 0.3 ppm total hydrocarbons exhibited reduced
condition Index. If the exposure concentration is multiplied times the time
In days at the point of significant effect, a total exposure of 36 ppm-days
(total hydrocarbons) or 1.2 ppm-days (total aromatlcs) is derived. The
0.010 ppm measured aromatlcs (at 0.3 ppm total hydrocarbons) In this study
does not Include monoaromatics which are likely to be an order of magnitude
higher in concentration than other aromatlcs.
A long-term study with North Sea crude oil, Involved exposures of a
bivalve, Mytllus edulis (Wlddows et al., 1982). Of the hydrocarbons
in the water-accommodated fraction (WAF), about 78% were monoaromatics;
naphthalenes made up the majority of the remaining aromatlcs. The authors
showed that food absorption efficiency and scope for growth were decreased
and the labilization period for lysosomes decreased after exposure to
0.03 ppm hydrocarbons for 7 days. Energy balance was clearly reduced by
about 35 days. The total exposure producing adverse effects was therefore
between 0.2 and 1.0 ppm-days.
In recent flowing exposures to water extracts and chemical dispersions
of Prudhoe Bay crude oil and distilled fractions of this oil, Anderson
(1984, 1985) has determined the total exposure producing 50% mortality for
fish and shrimp. The total exposure for 50% mortality for shrimp, Pandalus
danae, was 0.05 ppm-days (total aromatlcs) during exposures to WSF extracts
and chemical dispersions of fresh and weathered oil. This value Is very
consistent even with changes in the type of oil, the form of the oil (WSF or
dispersion) and the duration of exposure (up to about 9 days). Sand lance
(Ammodytes hexapterus) was more affected by oil droplets than soluble
aromatlcs, while both contributed to mortality which generally was not
observed until after 4 days. Significant latent mortality, 1 to 4 days
following termination of exposure to 1-2 ppm Indicated histological damage.
Physiological investigations with specific petroleum hydrocarbons or
WSF and Alaskan marine species have recently been summarized by Rice et al.
(1984). Table 15 (from their review) summarizes the concentration of test
material required to produce the abnormal response as a percent of the 96-h
LC50 value for each species. The response of pink salmon growth to WSF of
Cook Inlet crude oil was quite sensitive (10% of LCjq), but most responses
required an exposure of about one-third the 96-h LC50 concentration. The
authors conclude that oil concentrations at 0.1 to 0.2 ppm total aromatlcs
for 30 to 40 days could be harmful to the species listed in Table
15, particularly the invertebrates. This range of concentrations compares
76
-------
Table 15. Sublethal effects of the water-soluble fraction (WSF) and
Individual components of crude oil on Alaskan marine organisms.
Species
Toxicant
Length of
exposure
(days)
Parameters
affected
Percent of
LC50 that
affected
parameters
P1nk salmon
WSF
40
Growth
30
(fry)a'b
Naphthalene
40
Growth
28
WSF
4
Opercular rhythm
20
P1nk salmonc
WSF
10
Growth
10
(alevins)
Coho salmond
Toluene
40
Growth
50
Naphthalene
40
Qrowth
32
Seastar (Evasterlas
troschelil)e
WSF
28
Feeding rate
28
WSF
28
Growth
24
File periwinkle
(Thais lima)*
WSF
28
Scope for growth
18
(now called
Nucella lima)
Blue mussels
WSF
28
Scope for growth
19
(adult)9,h
28
Byssal thread extrusion 15
Blue mussels^
WSF
40
Growth, development
61
J Moles and R1ce 1983.
J! Thomas and Rice 1979.
; R1ce et al. 1975.
® Moles et al. 1981.
J 0*Clair and Rice In press.
I Stickle et al. 1984.
? Stickle et al. 1n press,
j Babcock et al. 1n preparation.
O'Clair and Rice In preparation.
77
-------
closely with the range of 0.1 Co 0.6 ppm total naphthalenes derived from
sublethal studies on temperate species summarized in Table 14.
The use of behavioral measures to assess stress from environmental
pollutants was reviewed by Olla et al. (1980). Many possible behavioral
responses of marine species may be altered by environmental perturbations.
The blue crab, Calllnectes sapidus, was studied by Pearson and Olla (1979;
1980) and Pearson et al. (1981a). They found that this crab could detect
naphthalene at 0.1 ng/1 and that concentrations above 2 ppm (mg/1) produced
defensive displaying and oriented locomotor activity. A VSF of Prudhoe Bay
crude oil at 2 ng/1 was detected by blue crabs and this concentration can be
found in chronically polluted marine areas (Pearson et al., 1981a). As
indicated by antennular behavior, a 24-h exposure to Prudhoe Bay crude oil
(0.27 ppm) reduced the detection of chemical food cues by Dungeness crabs,
Cancer magister (Pearson et al., 1981b). After 1 hour In clean water, the
antennular response recovered indicating no structural damage to sensory
cells.
Attempts to determine the threshold levels of petroleum exposure for
marine species may be most productive when expressed in terms of the total
exposure (days) to specific hydrocarbons (ppm of total aromatics). It
appears that exposures to constant concentrations of oil extracts between
0.05 and 1.0 ppm-days (total aromatics) produce adverse sublethal responses
In marine species.
TOXICITY OF HYDROCARBON-CONTAMINATED SEDIMENTS
Following a major oil spill or near a chronic point discharge of
hydrocarbon-contaminated wastewater, hydrocarbons tend to accumulate In
bottom sediments. The process by which hydrocarbons reach and are
Incorporated into sediments was discussed in detail in Section 3 of this
review. Once incorporated Into marine sediments, particularly in arctic
climates, hydrocarbons may be quite persistent. Therefore It is not
surprising that, following major oil spills, impacts have tended to be most
persistent and recovery slowest in benthic sedimentary environments (see
recent reviews by Teal and Howarthi, 1964; and National Academy of Sciences,
1985). Much of our knowledge of the effects of oil-contaminated sediments
on benthic populations and communities corned from opportunistic studies
following oil spills. In many cases, an incomplete understanding of the
structure and dynamics of the benthic communities prior to the spill has
hampered interpretation of the magnitude and duration of adverse Impacts
from the spill. There have been relatively few carefully designed and
executed laboratory studies of the effects of oil sediments on benthic
marine organisms.
There are two important concerns with respect to Impacts of oil
contamination of marine sediments. The first relates to the toxicity of
petroleum hydrocarbons associated with sediments and their long-term
effects on benthic community structure and function. The second relates to
contamination of benthic fauna and flora by hydrocarbons derived from
contaminated sediments and possible transfer of these hydrocarbons through
the marine food web to commercial fishery species (discussed in Section 5).
78
-------
Effects on Individual Species
Several studies have shown a lack of significant mortality of benthic
animals during exposures to concentrations of oil in sediments in excess of
1000 ppm (Wells and Sprague, 1976; Anderson et al., 1977; Roesljadi et al.,
1977 and 1976). Krebs and Burns (1977) reported the mortality and adverse
responses of the fiddler crab. Pea, living In sediment containing 1000 to
7000 ppm of petroleum hydrocarbons. Shaw et al. (1976) reported mortality
of the clam, Macoma balthlca, from sediment containing 640-3890 ppm (dry
weight).
Several investigations have been performed of the sublethal responses
of benthic marine invertebrates to oil-contaminated sediment or food (Percy,
1976; Shaw et al.f 1976; Taylor and Karinen, 1977; Roesljadi and Anderson,
1979 and Augenfeld, 1980). The Arctic amphipod, Onisimus affinis, avoided
oil-tainted food as well as oil masses, but the response diminished with
preexposure to oil and with the use of weathered oil (Percy, 1976). Taylor
and Karinen (1977) studied the behavior of the detritlvore, Macoma balthlca.
In response to oil-contaminated substrate and oil extracts flowing over
clean substrates. Since the clams burrow to the surface before dying (95%),
they determined that 502 of those exposed three days to 0.'37 ppm naphthalene
equivalents surfaced, and 0.2 ppm Inhibited burrowing within 60 min. Since
a 1Z (v/v) mixture of oil to seawater stirred for 20 h produces a
naphthalene equivalent concentration of about 0.36 ppm, the above values are
relatively high. Surface oiling of 670 mg oll/cm2 produced 50% surfacing by
the clams in 24 h. No death was recorded in these four- to six-day
experiments involving only a single oiling of the sediment. Shaw et al.
(1976) produced significant mortality In the same species by applying fresh
oil at doses of 5 mg of oil/cm2 dally for five days. Roesljadi and Anderson
(1979) found that Macoma lnquinata exhibited reduced survival, reduced
condition index and reduced levels of free amino acids, particularly
glycine, when exposed in the laboratory or field to sediments contaminated
to about 1,200 mg oil per kg. Augenfeld et al. (1980) determined that the
filter-feeding clam, ,;Protothaca staminea, was more resistant to oiled
substrates (about 1200 ppm) than Macoma lnquinata (detritlvore) as
demonstrated by both higher survival and' smaller alterations,in free amino
acid levels and condition index. Similar studies on the polychaete,
Abarenicola paclflca, have shown that concentrations of 500 and 1000 ppm oil
in sediment produced reduced feeding (measured by egestion rate) and
decreased tissue glycogen concentrations (Augenfeld, 1980; Augenfeld et al.,
1983). Clams (Protothaca staminea) less than 30 mm in size showed reduced
growth after one year In sediments contaminated with oil (1,251 - 5,176
ppm). Greater effects were shown when the lower concentration of oil was
mixed to depth (10 cm) in exposure trays (Anderson et al., 1983). Four- and
six-month exposures of the same species to layers of sediment containing
either oil or oil plus chemical dispersant (1:10) at about 2,000 to 3,000
ppm also reduced the rate of Protothaca staminea growth (Anderson et al.,
1985).
In the NAS (1985) review a section was dedicated to the fate
and effects of oil in polar environments. Table 16 contains the summary of
79
-------
TABLE It. SU»W»Y Of CfftCTS Of 01 LP SO I WENT S ON ALASKAN HADItg IUVCTTtajATES
tcocloa
Location
lublathll Iffact
hftfwn
IMpOtfl
Beaufort Sm
Inhibition of growth
*nd aoltlng only ot
high concentrations
Porej (197?)
H. Qfitoapn
N. alblrlca
Beaufort So*
Neutral reaponae to
pretence of oil, oil-
tainted food, end cont-
aalitatad aedlaentt
Percy (1976,
1977)
H. ontcaon
Point Borroa
Readily rocolonlted
contaalnated aedlaent
Atlaa et •).
(1979)
Aaphipod*
Boectaaieua affInia
Point Barron
food March auceoaa
reduced, racovary occur¦;
no offoct on retpiratloni
reduced burroaing In
cantaalnotod aedlaentai
reduced locoeotory
activity
Buadoth and
Atlaa (1977)
B. affinla
C—»aru» ocaanicua
Beaufort Soa
Avoid oil, oil-tainted
food, and contaalnatod
Mdlaont*
Percy (1976)
Coroohlm claronconao Beaufort Sea
Neutral behavior
reaponae to contaminated
aedlaonti
Percy (1977)
Pontogoreia fano
Acaroldea latloe
Hal 1 ta fsraoaa
Cemracanthua
lorlcatu*
Monoculodea ap.
Point Barron
little or no recoloniia-
tlon of coataslnatad
Atlaa et al.
(1978)
Polychoetea
Poctlnarla
hyoorboroa
Nephjfthy.
longoaatoae
Sjio tp.
Acanthoatoohlela
behrenglenala
Point Barron
Point Barron
Attraction to oll-
contaalnatod aedlaenta
No recoloniiatlon of
otl-conuninated
aedlaenta
Atlaa et at.
(1976)
Buadoth and
Atlaa (1977)
Atlaa ot al.
(1978)| Buadoth
and Atlaa (1977)
Moreli veitlHoae
hanaothoa loricate
Alaaka
Sentltlvity
Rice et al.
(1979)
Nolluaka
Wacocu balthlca
Culf of Alaaka
forced to aurfece Aen
oapoted to dlaaelved or
*od1aont»ab*orbod oil
Taylor and
Kartnen (1977)
80
-------
oiled sediment effects studies on Alaskan macroinvertebrates presented in
the review. As pointed out above in the discussion on the toxicity of PAH
in water. It is very difficult to compare the results of different studies
on the basis of the ppm total aromatlcs producing an impact. Concentrations
were not even Included in the NAS review on polar studies. Without data
quantifying concentrations of specific petroleum components in each
exposure system, it is not possible to reach a conclusion on the relative
sensitivities of infaunal species. In a study with the English sole,
Parophrys vetulus conducted in Puget Sound McCain et al. (1978) carefully
examined the change In petroleum hydrocarbon concentrations in the exposure
sediments. They found an effect or the growth of the English sole and liver
pathology resulting from exposure to sediments containing 700 ppm (dry
weight) Prudhoe Bay crude oil. Aromatic hydrocarbons including
trimethylbenzene and naphthalenes were found in the skin, muscle and liver
of the fish at the first time interval (11 d). Later in the exposure (27
and 51 d) hydrocarbons were only found In the liver.
Several behavioral studies have been conducted with oil (Prudhoe Bay
crude) contaminated sediments and bivalves or fish. Pearson et al. (1981b)
showed that the crab, Cancer maglster, consumed more llttleneck clams
(Protothaca staminea) from field enclosures containing oiled sediment (about
1,000 ppm) than from those with clean sand. They found in subsequent tests
that clams in oiled sand exhibited more shallow burial and slower
reburrowing than in control sand. 011a and Bjeda (1983) conducted a similar
series of experiments with the hard clam (Mercenarla mercenarla) and also
found oiling of sediments reduced the depth and rate of burrowing. The
number of clams buried in 96 h was reduced, from control values, at 1,000
ppm with further reduction at 3,000 ppm.
Sand lance, Ammodytes hexapterus, is a prominent forage fish from
California through Alaska and Is known to bury in sand during the night and
for long overwintering periods. Pearson et al. (1984) showed significantly
decreased time spent burled in oiled (306 ppm) sand. Another experiment,
oiled sand at 28 and 256 ppm did not reduce burial, but 3,384 ppm did. The
lack of effect at 256 ppm corresponded to a higher condition index In the
test group than that of fish used earlier (306 ppm). There was an apparent
Interaction between the use of a contaminated refuge and the nutritional
state of the fish. Interaction between oil contamination and sediment grain
size preference for A. hexapterus was studied by Pinto et al. (1984). Sand
lance avoided gravel and silt and preferred to bury in fine or coarse sands.
The fish preferred clean gravel when sand was oiled at about 120 or
1,000 ppm. If only silt and oiled sand are provided as choices, sand lance
prefered to remain in the water column.
BIOLOGICAL EFFECTS OF OIL SPILLS AND CHRONIC LOW-LEVEL DISCHARGES OF
HYDROCARBONS
A vast number of scientific papers has been published and technical
reports produced dealing with the biological impacts of the oil 6pllls
summarized in Table 4. Results of many of these investigations have been
summarized in reviews by Clark and Finley (1977) and Teal and Howarth
(1984), and NAS (1985).
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Massive kills of shallow and intertidal benthic organisms occurred
whenever and wherever large amounts of fresh crude or refined oil came
ashore. Massive immediate kills either did not occur or were impossible to
document in spills and blowouts that occurred well offshore and resulted in
little or no stranding of fresh oil (Argo Merchant, Chevron, Ekofisk,
IXTOC _I). In the Florida diesel fuel spill, massive kills occurred in
heavily oiled sites where sediment hydrocarbon concentrations exceeded about
130 ppm. Following the Amoco Cadiz spill, there were massive reductions of
some species, such as ampeliscid amphipods, where sediment hydrocarbon
concentrations exceeded 50-100 ppm. Ampeliscid amphipod populations also
were severely damaged in the Florida and Tsesls spills. These amphipods
apparently are among the most sensitive dominant benthic species in many
temperate and subarctic marine environments.
Littoral macroalgae appear In most cases to be fairly tolerant of
oiling, though as In the Torrey Canyon spill, they may be smothered by
massive deposits of oil. In some cases, a sensitive species (e.g.
Ascophyllum) was eliminated and was replaced rapidly by a more tolerant
species (e.g. Fucus). In most cases, the macrofauna associated with the
macroalgae was decimated wherever significant amounts of oil came ashore.
In the case of the Torrey Canyon spill, much of the damage to the rocky
Intertidal fauna was attributed to use of large amounts of chemical
dispersents containing aromatic solvents to clean the shore.
Intertidal salt marshes, particularly those containing peaty
high-organic soils, tend to be quite sensitive to oil spills, particularly
if the oil is fresh when it reaches the marsh and the marsh grasses are in
the active growth phase. Coastal marshes of Buzzards Bay, Massachusetts,
Impacted by diesel fuel from the Florida spill were severely damaged and
recovery was Incomplete even eight to twelve years after the spill (Sanders
et al., 1980). Several marshes, especially that at lie Grande, were
severely impacted by the Amoco Cadiz spill, but were subsequently destroyed
during the cleanup operation. Following the Arrow spill, Nova Scotia
marshes damaged by the Bunker C residual oil experienced a die-off of
Spartina alterniflora one year after the spill, with substantial recovery
observed two years later.
Damage to commerlcally important shellfish has been variable. There
were large losses of intertidal clams Mya arenarla following the Florida and
Arrow spills of refined oil products. Following the Amoco Cadiz spill,
oysters, shrimp and lobsters became heavily contaminated with oil, but did
not experience high mortalities. In fact, numbers of commerlcal shrimp
actually Increased along the north coast of Brittany in the two years
following the Amoco Cadiz spill, possibly due to decreased predatlon or
Increased microbial and algal production. Abundance of some mlcrofaunal
and melofaunal species may Increase in moderately oiled sediments,
providing Increased food for benthic deposit-feeders such as the clam
Macoma and the polychaete Arenicola.
Very few Impacts, other than tainting, have been observed in
phytoplankton and zooplankton communities following oil spills. Following
the Tsesls spill in the Baltic Sea, phytoplankton and planktonlc bacterial
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biomass increased, possibly due to a decrease in the abundance or feeding
activity of the larger zooplankton. Some evidence of contamination and
sublethal stress in zooplankton populations was reported following the
Arrow, Argo Merchant and Amoco Cadiz 6pills. These Impacts apparently were
of short duration and areal extent.
Adverse Impacts on commercially important pelagic and demersal fish
were reported following the Argo Merchant, Tsesls and Amoco Cadiz spills. A
year after the Tsesls spill, spawning frequency and hatching success of
herring Clupea harengus were reduced. For up to three years after the
Amoco Cadiz spill, plaice Pleuronectes platessa, collected from oil-
contaminated eBtuarles exhibited a variety of biochemical changes and
histopathological lesions. In female plaice from the contaminated
estuaries, there was evidence of delayed or suppressed ovarian development.
Growth rate and fecundity of the fish were reduced.
The long-term environmental Impacts of chronic low-level point-source
effluents containing PAH are quite different from those of massive oil
spills. There usually Is no massive die-off of marine fauna and flora.
Instead, there may be a gradient of Impact, characterized by altered
community structure, abundance, and diversity extending around the
pollutant source. There have been several Investigations of impacts of
refinery effluents and oil tanker terminals on the coastal marine
environment (see review of Dicks and Hartley, 1962) and a few studies of
impacts of produced water discharges (see review of Neff, 1985).
As discussed above, approximately 150 metric tons per year of petroleum
hydrocarbons, Including about 4 tons of low molecular weight PAH, are
discharged from the ballast water treatment facility to Valdez Harbor. The
mean water depth of Valdez Harbor is 180 meters and the water Is highly
stratified, particularly during the summer. The effluent dlffuser system
discharges treated ballast water below the density discontinuity and there
Is little evidence of vertical mixing of discharged hydrocarbons Into the
surface waters (Lysyj et al., 1981). Maximum hydrocarbon concentrations
occur at about 50 meters water depth In spring and summer. Three to five
meter tides afford fairly rapid exchange of water in the harbor and
probably provide the major mechanism for elimination of hydrocarbons from
the harbor. Because the discharged .hydrocarbons remain deep where
temperatures rarely rise above 5°C, photodegradatlon, evaporation, and
microbial degradation probably play only minor roles In removal of PAH from
the harbor. PAH not flushed from the estuary probably accumulate in bottom
sediments where they may persist indefinitely.
It appears then that the benthos is where biological Impacts of the
discharge are most likely to occur and where they should be sought. Cowell
and Monk (1981) proposed monitoring of intertldal rocky shore populations.
To date, results of several pre-dlscharge baseline studies and some
preliminary post-discharge studies have been published on the benthos of
Valdez Harbor (Feder and Matheke, 1980). Diversity of benthlc fauna is low
in deeper parts of the harbor because of low dissolved oxygen, and
significant impacts of the ballast discharge have not been documented.
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Milford Haven on the vest coast of Great Britain has a physiography
somewhat similar to that of Valdez Harbor and has been the site of a major
oil port and refinery complex since 1960. Extensive environmental studies
at Milford Haven extending over the last two decades have been summarized by
Dicks and Hartley (1982). It has been estimated that approximately 238
metric tons (270*000 liters) of petroleum hydrocarbons enter Milford Haven
each year from all sources. This is nearly twice the amount being
discharged each year from the ballast water treatment plant into Valdez
Harbor. There have been changes in community structure of intertidal and
benthlc fauna and flora In some parts of Milford Haven attributable to the
hydrocarbon Inputs. However, there has been no general decline in the
marine flora and fauna. The lack of more substantial Impacts was
attributed in part to rapid flushing of the estuary.
Reported impacts of effluents from oil refineries have been quite
variable. Impacts of a refinery effluent to Littlewlck Bay in Milford Haven
were restricted to an area of about 200 meters around the outfall. However
a refinery effluent to the Medway estuary on the east coast of Great
Britain caused substantial damage to Intertidal macrofauna extending out to
at least 1.5 km from the outfall (Warfe, 1975). Discharge of refinery
effluent to a salt marsh environment on the south coast of Great Britain
caused substantial damage to the salt marsh vegetation (Dicks and
Hartley, 1982). Subtidal benthlc communities also were damaged. Only
two benthlc polychaete species survived near the outfall. Effluent
quality (hydrocarbon concentration) was improved and volume decreased
between 1972-1974. This resulted in a gradual recovery of the salt marsh
vegetation. After about ten years, the marsh appeared healthy, though
species composition of plants and animals still was different from that in
nearby uncontaminated areas. Subtidally, there was little evidence of
recovery.
The environmental Impacts of the produced water discharges to Cook
Inlet probably are minimal despite their large volume, because of the
dynamic nature and rapid flushing of the water in the inlet. However, no
detailed long-term studies have been performed to date.
The best investigations of the impacts of produced water and related
production discharges from oil platforms are those in Trinity Bay, Texas
and the Forties Field in the North Sea.
Trinity Bay Texas is a shallow, low-salinity estuary, very much
unlike Cook Inlet. However, It does in some ways resemble some nearshore
environments of the Alaskan Beaufort Sea where produced water might be
discharged in the future. During the 20-month timecourse of the Trinity Bay
study, produced water with a mean total hydrocarbon concentration of 15 ppm
was discharged from the separator platform through an outfall one meter
above the bottom at a rate of 650,000 to 1,590,000 liters/day (Armstrong et
al., 1979). Hydrocarbons were diluted nearly 2,500-fold in the water column
within 15 meters of the outfall. Bottom sediments were heavily contaminated
with medium molecular weight alkanes (C10 ~c28 n-paraffins) and aromatlcs
(C3 benzenes - trimethylphenanthrenes).
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There was a gradient of decreasing naphthalenes concentrations from a
mean of about 21 ppm 15 m from the outfall to background 500 to 4800 m from
the outfall, depending on direction. There vas an inverse gradient of
numbers of organisms and numbers of species of benthic infauna with
distance from the outfall. Within 15 m of the outfall, the bottom vas
almost devoid of organisms. Benthic faunal abundance vas significantly
reduced out to approximately 150 m in all directions from the outfall. At
stations located 685-1675 meters from the outfall, there vas an apparent
enhancement of the benthic fauna, vith greater numbers of individuals and
species at these stations than at reference stations 4,000-5,800 m from the
outfall. Thus, a 150-m radius zone of adverse impact was observed, vith an
apparent zone of enhanced faunal abundance and diversity further out from
the discharge, and impacts were correlated to comtaminatlon of sediments
vith petroleum hydrocarbons. Armstrong et al., (1979) estimated that a
nominal concentration greater than about 2 ppm total naphthalenes was
necessary to significantly reduce benthic Infaunal populations of Trinity
Bay. Results of these investigations should be extrapolated to offshore
situations with extreme caution. The shallow turbid nature of the receiving
vater Is unlike the situation encountered offshore, vith the possible
exception of some near shore areas of the Beaufort Sea. Where vater depth
is greater and suspended sediment concentrations are lover than those
encountered in Trinity Bay, a much smaller fraction of the hydrocarbons in
the discharged produced vater vlll be deposited in bottom sediments near the
outfall, and adverse effects on the benthos vlll be much less severe.
Hartley and Ferbrache (1983) recently reported the results of a similar
benthic monitoring study In the Forties Field located in the British sector
of the North Sea in 100-125 meters of vater approximately 177 km northeast
of Aberdeen, Scotland. The 80 veils drilled to date from four platforms
vere drilled vith vater base muds. Current rate of produced vater discharge
vas not stated. It is expected to increase vith the age of the field and
oil-water separators vere built on each platform vith a nominal capacity of
40 million liters per day.
Production from the field began in September, 1975, and three benthic
surveys have been performed to date; in June, 1975, before production began,
and In 1978 and 1981. Concentrations of aliphatic hydrocarbons in sediments
measured by gravimetric techniques have shovn a slight rising trend from a
mean of 5.7 ppm in 1978 to 8.9 ppm in 1981. Grain-size distribution of
sediments did not change significantly at any sampling stations during the
three surveys. No analyses vere performed on metals in sediments.
The benthic fauna vas rich and diverse throughout the study area.
Hovever, directly beneath the Platform C and to about 100 meters to the
vest, the benthic macrofauna vere severely depressed and there vas evidence
of hydrocarbon contamination (305-470 aliphatics), possibly from dlesel
fuel. The abundance of the polychaete Chaetozone setosa vas much higher
near Platforms A and D than at other stations. Abundance of another
opportunistic polychaete, the capltellid Capltomastus minimus, Increased in
muddler sediments to the vest of the rigs betveen 1978 and 1981.
Rate of uptake and mineralization of luC-naphthalene by sediment mlcroblota
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increased with decreasing distance from platforms in the Forties Field
(Saltzmann, 1982). In addition, three species of demersal fish (cod,
whiting and haddock) had slightly elevated activity of hepatic aryl-
hydrocarbon hydroxylase (aromatic hydrocarbon detoxification enzyme system)
compared to the same species collected In a reference area about 30 km west
of the field. Overall, it is apparent that impacts of oil production
activities from the Forties oil field have been localized (primarily within
450 m of the platforms) and of low magnitude.
There is little doubt that a massive oil spill in the near shore
Alaskan marine environment (i.e. from the grounding of a large tanker
leaving Valdez or Cook Inlet) would cause serious damage to important
components of the marine ecosystem, as occurred following the Amoco Cadiz
spill. Low ambient water temperatures are likely to decrease the rate of
evaporation, photooxldatlon and blodegradation of spilled oil Increasing the
residence time and potential environmental damage of the more toxic, low
molecular weight PAH. Available evidence would seem to Indicate that
current discharges of produced water to Cook Inlet and treated ballast water
to Valdez Harbor are not causing significant environmental damage. However,
seasonally low water temperatures and Incident light intensity slow the rate
of degradation of PAH, possibly allowing for their gradual accumulation over
several years in bottom sediments. Efforts should be made to monitor
long-term trends in concentrations of PAH in sediments and associated biota
in the vicinity of these chronic discharges to determine if serious
long-term biological impacts might occur.
RECOVERY OF BENTHIC COMMUNITIES
In the previous section of this chapter we reviewed the biological
effects of oil spills and chronic effluents. Host of the study results
dealt with the Impacts of spills on benthic comunities. Therefore, this
section will summarize findings related to the recovery of benthic
communities from either accidental or experimental contamination with oil.
Following a spill of Mo. 2 Fuel oil in Buzzards Bay off West Falmouth,
Massachusetts in September, 1969, extensive long-term investigations were
performed on the process of recovery In the spill-impacted salt marsh and
shallow benthic environment (Sanders, 1980; Sanders et al., 1980). At the
most heavily impacted benthic station studied in Wild Harbor, hydrocarbon
concentrations In sediments were quite variable, reflecting a patchy
distribution. However, concentrations remained above 1,000 ppm, and as high
as 12,400 ppm between October, 1969 and September, 1971. There was an
immediate near-total eradication of the benthic fauna at this station. This
was followed within three months by a rapid increase In the abundance of the
opportunistic polychaete Capltella capitata, which reached densities of up
to 95,000 per m2 by May, 1970. After about seven months, Capitella
populations crashed and were replaced by a few less opportunistic species.
Number of individuals and species abundances fluctuated widely, and by the
end of two to three years had not risen to levels encountered there at the
time of the spill.
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At stations just offshore from Wild Harbor, petroleum hydrocarbon
contamination of sediments following the spill was much less severe than in
the harbor. At Stations 9 and 10, sediment total petroleum hydrocarbon
concentrations were In the range of 150 to 250 ppm In October to November,
1969, and had dropped to the range of 46 to 88 ppm by mid to late 1971. At
these stations, number of species and faunal densities dropped to low levels
during the seven to eight months after the spill. In July to August, 1970,
densities and numbers of species rose greatly. This resulted from
recruitment of larvae from the water column and resulted in a 20- to 30-fold
Increase In number of species. The benthlc fuana at this time were
dominated by the opportunistic capitellld, Mediomastus ambiseta. This was
followed in late 1970 by a sharp decline In density and species abundances
which continued to fluctuate seasonally at lower levels during 1971 and
1972. Sanders (1978) interpreted the higher species abundance and lower
faunal density and variation during the third year after the spill to be
indicative of some recovery.
Sediments at slightly oiled stations further offshore never contained
petroleum hydrocarbons at concentrations higher than about 50 to 80 ppm.
These stations showed little variation in species composition and much
smaller variations in density than inshore stations during the three-year
post-spill investigation. Thus, it would appear that sediment hydrocarbon
concentrations from No. 2 Fuel oil below about 100 ppm allow early recovery
to proceed, and concentrations in the range of 50 to 80 or less do not
seriously affect benthlc communities.
Investigations have Indicated that melofaunal communities, particularly
of estuarine areas, are quite resilient following natural disturbance or
organic pollution (i.e., Cato et al., 1980). For example, Along! et al.
(1983) studied recolonizatlbn of meiobenthos in trays containing azoic sandy
sediments with or without added crude oil (Prudhoe Bay crude oil). The
trays were deployed In 1 m of water in the lower York River, Virginia
during the summer. Total petroleum hydrocarbon concentrations in the
the sediments in the trays Initially ranged from 47 to 9,100 ppm, with total
aromatic hydrocarbons initially ranging from 0.43 to 136 ppm. The rate of
recovery of the different melofaunal groups was very rapid in surficial
sediments and somewhat slower In deeper sediments, below the redox potential
discontinuity, in the trays. Nematodes were the dominant melofaunal taxon
In all treatments, and also were the slowest to recover in the trays to
abundance and community composition similar to that in study site control
sediments. In fact, recovery of nematodes was not complete in 90 days in
deep sediments in the middle and highest oiled trays. The results of this
study Indicate that recovery of benthlc meiofauna was complete at 28 days at
all depths in sediments containing 3,900 ppm total oil and 29 ppm aromatic
hydrocarbons initially and 1,285 ppm total oil In 15.75 ppm aromatic
hydrocarbons at 28 days.
Similarly, Boucher (1980) reported little change In the abundance of
melofaunal nematodes and harpacticoid copepods in subtidal fine 6and
sediments in 19 m of water in the Bay of Morlaix, France, heavily Impacted
by the Amoco Cadiz oil spill (up to 3,800 ppm saturated petroleum
hydrocarbons in sediments). However, there was evidence of a decrease In
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nematode diversity in the oiled sediment, due to an increase in the
abundance of a few dominant species following the spill.
Two small estuaries on the north coast of Brittany, France, Aber Benolt
and Aber Wrac'h, became heavily contaminated with Amoco Cadiz oil following
the spill. Glemarec and Hussenot (1982a, b) investigated the recovery of the
benthlc macrofauna of the two estuaries following the spill. The estuaries
contain a wide variety of different soft substrates, from fine lntertidal
muds to coarse sand flats, so it was possible to study the course of
recovery of different types of sediment regimes.
During the three months after the spill, oil spread throughout the
estuaries, and sediment petroleum hydrocarbon concentrations in most areas
were in the 1,000- to 10,000-ppm range (measured by Infrared analysis:
Marchand et al., 1981). The exceptions were the coarse sandy areas at the
mouths of the estuaries. There were immediate massive kills of the benthlc
fauna of the estuaries after the spill, and nearly all communities still
were extremely depauperate eight months later. The coarse sand communities
at the mouth of Aber Wrac'h were little affected by the spill and showed
near normal seasonal fluctuations. The highest oil concentration measured
in these coarse sand sedimentB was 1,020 ppm in May, 1976, three months
after the spill. In the subtidal dune-sand sediments off Aber Benoit,
petroleum hydrocarbon concentrations never exceeded 50 ppm, and there was
some evidence of only a temporary decrease in total abundance with no
significant change in benthlc community structure.
Muddy sediments at the heads of the estuaries were the most heavily
oiled and most severely damaged. After eight months, highly opportunistic
species, characteristic of early stages of recovery from organic pollution,
began to appear and reached very high densities. Peak abundance of
opportunists was not reached until 20 to 25 months after the spill. At
most other locations in the estuaries, characterized by fine sands and muddy
sands, dominant benthlc fauna eight months post-spill Included the less
extreme opportunist species, primarily cirratulld and capitellld polychaetes
(e.g., Chaetozone setosa, Heteroclrrus spp., Polydora spp., Cirratulus
clrratulus, etc.). After about 17 months, these opportunists were
joined by a group of sensitive species which typically disappear following a
perturbation and then reappear over a broader ecological range (e.g., Spio
sp., Notomastus laeterlceus, Phyllodoce spp., Nereis diversicolor, etc.).
Winter storms resulted in sediment resuspension and a decrease in some
cases in sediment hydrocarbon concentrations. These were followed,
especially in the second winter after the spill, by a decrease in
opportunists and an increase in later successions! stages. In Aber Benolt,
the decline In opportunistic species, and replacement by later successional
stages began after eight months In well-oxygenated dune, sands, after
13 months In heterogeneous muddy sands, after 17 months In fine sands and
sandy muds and after 25 months In the muddy sand areas at the head of the
estuary. Two to two and a half years after the oil spill, the communities
of the dune sand areas had recovered completely, while earlier successional
stages still were evident at all other substrates. Communities in Aber
Wrac'h recovered more slowly than communities in similar substrates in Aber
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Benolt. Glemarec and Hussenot (1982a, b) estimated that It would require
three to four years for benthic communities of the Abers to return to normal
or to a new equilibrium condition.
Cabioch et al. (1982) and Dauvin (1982) reported similar changes
following the Amoco Cadiz oil spill in benthic communities occupying
different sediment types in deeper coastal waters of the Bay of Morlaix.
Coarse sandy sediments were little affected by the spill and their
communities recovered rapidly. The fauna of fine muddy sand, particularly
several species of astphlpods of the genus Ampelisca, were severely Impacted
by the hydrocarbons and were very slow to recover. Significant
recolonlzatlon of severely impacted sediments did not begin until the second
annual cycle of recruitment (spring, 1980). There was further substantial
recruitment in the third spring following the Bplll. The dominant fauna of
the heavily Impacted offshore fine sand sediments during the next year
after the spill were opportunistic species of capltellid and clrratulid
polychaetes. Approximately three years after the spill, the communities of
the fine sandy sediments in the Bay of Morlaix had returned to the normal
pre-spill range in terms of density, species richness, and blomass.
Abundance of amphipods still was depressed in some areas.
Following the grounding of the Soviet tanker Tsesls on the Swedish
coast of the brackish Baltic Sea south of Stockholm in October, 1977,
approximately 1,000 tons of medium grade (No. 5) Fuel oil were released
(Elmgren et al., 1983). Although about two-thirds of the oil was recovered
mechanically, significant amounts of oil reached the rocky shores and the
benthic environment.
Within 16 days after the spill, an estimated 95 percent of the
dominant amphipods, Pontoporeia affinis and £. femorata and polychaetes,
Harmothoe sarsi were eliminated from the benthos at the most severely
impacted station. . The deposit-feeding clam (Macoma balthica) was resistant
to the pollution but became heavily contaminated with oil (up to 1,000 ppm
dry wt.). All melofaunal groups except nematodes were reduced in abundance
in the affected area.
In the summer following the spill (1973), there was a very heavy
recruitment of Juvenile Macoma balthica in areas where amphipods had been
virtually eliminated. Not until the summer of 1979 did amphipods, Harmothoe
sarsi, and harpactlcoid copepods begin to return to sediments of the most
heavily impacted stations, and oil contamination of Macoma balthica decrease
to 1,000 ppm. By the third year after the spill, abundance of amphipods
still was depressed but Harmothoe sarsi had returned to its pre-spill
abundance. Oil concentrations in Macoma balthica had dropped to about
250 ppm. Because of the long life span of Macoma balthica, the authors
predicted that the disturbed community structure at the heavily oiled
stations would persist for several years. Full recovery could take five to
ten years.
Leppakoski and Lindstrom (1978) studied recovery of benthic macrofauna
in a small low salinity (5.4-6.7°/00 salinity) harbor on the Finish coast
of the Baltic Sea receiving effluents from an oil refinery, following nearly
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complete abatement of the oily waste discharges. The toxicity of sediments
to benthic Insect (Chironomus plumosus) larvae decreased markedly In the
first three years following cessation of oily wastewater discharges. Number
of species and individuals, and species diversity of benthic macrofauna
increased sharply at stations near the former outfalls. However, three
years after abatement, there was a trend toward reduced diversity,
characterized, by a marked increase in the dominance of the bivalve Macoma
balthica, as occurred following the TseslS' oil spill. Thus, substantial but
not complete recovery of the benthic community had occurred three years
after near complete cessation of oily wastewater discharges to the harbor.
Vanderhorst et al. (1978, 1979, 1980 and 1981) utilized trays of
sediment with mesh bottoms placed in the field (lntertldal and subtidal) to
measure recruitment of infaunal species and loss of petroleum contamination.
Using this approach it was possible to predetermine the number of replicate
trays and cores from these trays required to detect a 50% reduction In the
numbers of individuals and species with 95% confidence (Vanderhorst et al.,
1978). Vanderhorst et al. (1981) experimentally evaluated the effect of oil
treatment, site, substrate type, season and tide level on the composition,
density and species richness of organisms colonizing substrates which were
initially free of organisms. Significant differences for some biological
parameters were demonstrated for each of the types of treatment contrast
(site, substrate type, season, tide level, and oil). Significant biological
effects were demonstrated to be due to oil treatments for 70% of 56
biological parameters evaluated in detail.
Predicted full recovery for sand habitats in Puget Sound was 31 months
following an initial oil treatment of 1,800 ppm. Predicted full recovery
for a commercial clam bed habitat was 46 months following an Initial oil
treatment of 2,500 ppm. The studies of Vanderhorst et al. (reviewed in
Vanderhorst, 1984) used sediments devoid of benthos as a starting substrate.
All infauna identified and counted in the trays had to either migrate
from surrounding substrates or settle as larvae. In an actual spill some
organisms may well survive, thus increasing the rate of recovery of the
total benthic community.
SUMMARY
Virtually all the investigations to date of the recovery of benthic
communities from acute or chronic pollutant stress have been performed in
estuarine and nearshore coastal waters. Under these conditions, the
responses to and recovery from petroleum hydrocarbon pollution have been
remarkably similar.
Factors which did seem to affect the rate of recovery Include: water
depth, to the extent that it influenced the amount of storm-induced sediment
resuBpenslon; magnitude of pollutant accumulation in the sediments; sediment
grain size (recovery was more rapid in coarse than in fine sediments); and
availability of planktonic larvae for recruitment.
Fuel oils (No. 2 and No. 5 Fuel oils) seem to produce more severe
Impacts and result in slower recovery than crude oil. This undoubtedly Is
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due to the higher concentration in fuel oils than in most crude oils of
light aromatic hydrocarbons, particularly the toxic naphthalenes and
phenanthrenes (Neff and Anderson, 1981).
Generally, the estuarine and coastal marine benthic ecosystems studied
to date recovered rapidly from petroleum pollution. Recovery times of a
month (meiofauna) to five to ten years have been suggested. This rate of
recovery lends support to the conceptual model of Boesch and Rosenberg
(1981) that estuarine communities are more resistant to stress than strictly
marine communities because they are composed of more eurytolerant species.
In addition, such communities are more'resilient than oceanic communities
because they include more opportunistic species acclimated to frequent
disturbance.
ALASKAN PERSPECTIVE
Alaskan marine animals do not appear to be significantly more sensitive
to aromatic hydrocarbons than similar species from more temperate and
tropical climates, however, because the volatile, toxic light aromatic
hydrocarbons associated with spilled oil tend to persist longer in the
water column at low water temperatures characteristic of Alaskan waters
than at higher water temperatures, biological impacts of oil spills In
Alaskan waters may be more severe than, those In warmer climates. The
greater persistence of spilled oil in arctic marine sediments and
shorelines, compared to more temperature climates, may result in a
substantially slower recovery of damaged marine ecosystems. However, many
Alaskan benthic and coastal environments already are severely stressed by
natural environmental extremes such as seasonal ice cover or scouer,
freshwater runoff, and high suspended sediment loads. The biological
communities inhabiting such environments are dominated by large
populations of rapidly reproducing opportunistic species, the very species
that are the first to return to a site damaged by oil or other organic
pollution. Such marginal, naturally stressed communities tend to recover
rapidly from acute pollution events such as oil spills.
Under such circumstances, the populations most likely to be severely
impacted by and exhibit the slowest recovery from an oil spill are the large
long-lived species such as king crabs, salmon, and marine mammals. Vlth few
exceptions, early life stages of these large animals are much more sensitive
to oil and other toxicants than later life stages. Thus, the most severe
damage to the Alaskan marine environment from an oil spill would result from
destruction of the early life stages of these large species or damage to
their preferred habitats.
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7. INFORMATION NEEDS AND RECOMMENDED MONITORING STRATEGIES
The Alaskan marine environment is so vast and varied that, despite the
tremendous advances made during the last ten years due in large part to the
NOAA-OCSEAP Program, our knowledge of the physical, chemical, and biological
processes by which the Alaskan marine environment functions is still very
fragmentary. The more we know about the natural processes that control a
marine ecosystem, the easier it is to design a meaningful monitoring program
to detect the long-term impacts, or lack thereof, of man's activities.
However, neither time nor financial resources will allow us to learn
everything about the Alaskan marine environment. Therefore, in the present
context, it is necessary to focus on better understanding those processes
which we expect to have the greatest Influence on the fate and effects of
PAH and related hydrocarbons in the Alaskan marine environment.
INFORMATION NEEDS
Based on the Information discussed In this review, we have identified
several data gaps or information needs and have developed some general
strategies for design of long-term monitoring programs.
1. More Information is needed on the sources, both natural and
anthropogenic of PAH and related hydrocarbons In the Alaskan
marine environment. A more complete and quantitative
Inventory should be developed of PAH discharged from major
known point sources: I.e., treated produced water and ballast
water discharges, natural oil seeps, domestic and Industrial
sewage, oil-contaminated drilling fluids, onshore oil handling
and trans-shipment facilities and other Industrial activities.
Non-point sources should be better characterized: I.e.,
vehicular exhaust, including boats; open burning and fores
fires; home heating, particularly with wood and coal; erosion
of peat coal and other PAH-containlng strata; arctic haze;
etc. With such an inventory, it will be possible to predict
better the consequences of potential new Inputs of PAH from
accidental (I.e., oil spills) and intentional (i.e., allowing
discharge of produced water to the Beaufort and Bering Seas)
new sources.
2. Our knowledge of the hydrocarbon composition of produced
waters from different sources is very Incomplete. It is
possible, even likely, that the composition of produced water
will change with time over the life of a well, or be different
from one well to another in a given area. Yet the published
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Information on ehe composition of produced water from offshore
veils comes from a few spot analyses of produced water from
Cook Inlet, the Gulf of Mexico and southern California. Most
analyses have been of the light hydrocarbon fractions only
(low molecular weight alkanes and monoaromatics). There are
very few analyses of medium and high molecular weight PAH In
produced water, and virtually none of the non-hydrocarbon
organic fractions which may constitute more than 90 percent of
the dissolved organlcs in produced waters. A long-term
research program to characterize the organic composition of
produced waters discharged to the Alaskan marine environment
would be very valuable. Samples of produced water could be
taken once or twice a year for several years from all produced
water discharges to Alaskan marine waters. The samples should
be analyzed by the latest methods for individual alkanes and
aromatics, through C3i» and benzo(a)pyrene, respectively, for
sulfur and nitrogen heterocyclics, trlterpenes and steranes,
and for organic acids and phenols.
3. The environmental fate and effects of treated produced water
and ballast water discharges to Alaskan coastal waters should
be studied in more detail. What is the residence time (half-
life) of ballast water hydrocarbons in Valdez Harbor? Are
there any sinks for these materials in the harbor where they
may accumulate to high concentrations over time? What biota
are most likely to come in contact with and be affected by
these chronic discharges? The same questions can be asked
about produced water discharges to Cook Inlet.
4. The focus of much new oil exploration and development offshore
Alaska in the next several years will be in the Beaufort Sea
and northern Bering and Chukchi Seas. Our knowledge of the
biological processes In these arctic seas is very poor. In
many cases, we have very Incomplete knowledge of such basic
processes as the structure of the marine food webs leading to
commercially important or Inherently valued marine species
(i.e., arctic clsco, marine mammals),. In addition, we know
very little about the sensitivity of these truly arctic
species to oil and hydrocarbons. Most work on the toxicity of
petroleum hydrocarbons to Alaskan species has been performed
at the NOAA laboratories in Seattle, Washington, and Auke Bay,
Alaska, with subarctic species. Are arctic species more or
less sensitive than subarctic species to petroleum
hydrocarbons? What Is the timecourse and magnitude of
Induction of the mixed function oxygenase system In arctic
fish? Can arctic invertebrates metabolize PAH? What is the
timecourse and extent of bloaccumulatlon and depuration of PAH
by arctic marine animals?
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5. There have been extensive efforts for the past several years
to model the behavior and fate of spilled oil under arctic
sea-ice conditions. These models should be field-verified and
tested in carefully designed experimental spills in the field.
The models and field data should be used to develop optimal
strategies for treating or counteracting spills under
different environmental and climatic conditions.
MONITORING PROGRAMS
1. A long-term monitoring program should be initiated to assess
the fate and effects of produced water discharges to Cook
Inlet. This monitoring program should Include characteriza-
tion over time of the volume and chemical composition of
produced water discharged from an offshore platform or
shore-based treatment facility; modeling or actual tracking of
the movements and dilution of the produced water plume in the
water column; deposition and accumulation of hydrocarbons and
metals from the produced water discharge In bottom sediments
near the discharge or at possible deposltional sites. If any,
in the vicinity.
Until reliable remote water sampling devices are developed for
taking Integrated samples over weeks, bivalves are likely the
best means of obtaining an integrated sample of bloavallable
(potentially toxic) contaminants. When a sediment trap Is
deployed above the cage of bivalves at various stations
around a discharge, there Is the opportunity to compare
particulate-bound contaminants with those bioaccumulated. In
the benthlc environment, detrltivorous bivalves, such as
Macoma species, have been shown to bioaccumulate contaminants
from sediments. Containers (trays, cages in the bottom, etc.)
with these clams can be placed on the bottom at different
distances from the outfall. In addition to determining
bioaccumulatlon, the condition index, growth, and energy
balance (scope for growth) can be measured on these bivalves.
It is necessary to determine the Impacts on the infaunal
community (balanced indigenous population, BIP). Based on
previous studies, the strongest statistical evidence for
impacts to benthlc populations has come from experimental
Installations of substrate to determine effects on
recruitment. Reciprocal transplants of sediment trays will
provide evidence of impacts frqm the water and sediment
separately.
With proper planning It is possible to design experimental
installations at various distances from the discharge that
will determine:
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a. accumulation of contaminants at the surface microlayer, In
the water column, and on the bottom;
b. relation between the bioaccumulation of toxic components
and the degree of impact on organisms;
c. spaclal extent of Impacts on individual species and the
benthic recruitment of larval stages of many species.
It would appear that these findings would answer all the key
questions posed by the resource manager attempting to protect
the marine environment of Cook Inlet, Alaska.
2. Offshore development and production of oil reservoirs is just
getting underway in the Beaufort and could well start in
several parts of the Bering Sea.in the next decade. A long-
term effects program should be performed during development of
new offshore oil fields in these frontier areas.
At an early stage in the design of this program, a clear
definition should be developed of what constitutes a
significant environmental impact. Particular attention should
be paid to defining Impacts that, if they occurred, would
be of consequence relative to commercially important or
intrinsically valued resources of the area. The monitoring
program should be designed specifically to Identify these
Impacts. Most major environmental monitoring programs in the
past have suffered severely or even failed due to a lack of
clear definition of what constitutes an undesirable Impact.
If the major objective of the monitoring program is to
document long-term trends in the magnitude and areal extent of
environmental change attributable to the development activity,
a fairly modest program may be adequate. A carefully-selected
set of physical, chemical, and biological samples and
measurements could be taken on an annual basis at a relatively
small number of stations along a potential pollution gradient
In the development field. The nearest stations would be
within 200 m of the development platform in order to maximize
the likelihood of detecting a "signal." Observations would be
made and samples taken at the same time each year for at least
the duration of development of the field. Such a design is
particularly appropriate for the Beaufort and northern Bering
Seas, because field sampling is extremely difficult for
all but a brief open water period each year (mid-August to
mid-September). Selection of parameters to measure should be
based upon the definitions of Impact as described above and
might Include characterization of rates and long-term trends
of change in benthic community structure, recruitment and
age-size structure; Induction of mixed function oxygenase
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activity in demersal fish from the area; and measures of
reproductive success, fecundity, and growth in representative
indigenous populations, possibly including sentinal organisms
(i.e., the mussel watch concept). An essential component in
such a monitoring program is to correlate long-term trends in
biological parameters with trends of change In concentrations
of contaminants in sediments and biota. Contaminants of
concern include aromatic hydrocarbons and heavy metals.
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8. SUMMARY
The major point sources of polycyclic aromatic hydrocarbons (PAH) in
the Alaskan marine environment are discharges of treated produced water,
crude oil tanker ballast water and domestic/industrial sewage. These, as
well as new point sources of PAH, can be expected to increase in number and
volume as offshore reserves of oil and gas are developed and Industrial
activity in Alaska increases. Currently, these point sources contribute
only a small portion of the total PAH entering the Alaskan coastal waters
from all sources. Major non-point sources of PAH in Alaskan coastal waters
and sediments are aerial deposition of particle-bound PAH derived from
remote industrial and other combustion sources. Burning of wood for home
heating and in controlled or wild forest fires may be major sources of these
airborne particulate PAH in some parts of Alaska. Additional important
sources of PAH include erosion of peat and coal deposits and submarine oil
seeps.
The composition of hydrocarbon assemblages in marine sediments of
developed and remote areas of Alaska reveal a predominantly biogenic
(natural) and pyrogenic (combustion) origin. Oil spilled in the ocean
under arctic and subarctic conditions similar to those in Alaska tends to
be quite persistent. In coastal areas characterized by high suspended
sediment loads, such as Cook Inlet, the Beaufort Sea and Norton Sound off
the Yukon River, PAH from spilled oil will adsorb rapidly to suspended
sediment and be transported to the bottom, where they will be quite
persistent. Evaporation of low molecular weight aromatic hydrocarbons is
slow at low water temperatures and is nearly completely Impeded by ice
cover.
The main mechanisms of removal of PAH and related hydrocarbons from
Alaskan marine waters are evaporation, photochemical and chemical
degradation, and metabolism by marine bacteria, fungi, phytoplankton, and
animals. In Alaskan waters, these processes tend to proceed more slowly
than at lower latitudes because of low ambient temperatures and low net
incident solar radiation during much of the year. Thus, PAH Introduced
into Alaskan waters by natural mechanisms, intentional discharges, or
accidental spills will tend to persist and may accumulate over time to high
concentrations in Alaskan marine sediments.
Alaskan marine animals readily accumulate PAH and related hydrocarbons
during exposure to these compounds in the water, food, or sediment.
Bloaccumulation is most rapid and efficient from the water. However, since
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PAH reach high concentrations and are more persistent in sediments than in
water* the major source of PAH for benthic and demersal marine animals is
from contaminated sediments.
Bioconcentration factors (concentration in tissues/concentration in
medium) for PAH Increase with Increasing PAH molecular weight and tend to
be higher in marine molluscs than in polychaetes, crustaceans, and fish.
This Is directly related to the relative capability of these taxa to
metabolize and excrete PAH. Because PAH are metabolized by members of
higher trophic levels* there is no evidence of blomagnlflcatlon of PAH in
marine food webs.
Low temperatures* characteristic of Alaskan waters* have only a slight
effect on rate of accumulation of PAH in marine animals* but do seem to slow
metabolic degradation and excretion of accumulated PAH. The slower rate of
depuration plus the greater persistence of, PAH in low temperature marine
environments* may mean that the potential for chronic Impacts of PAH
pollution of the Alaskan marine environment is greater than for more
temperate and tropical climates.
Alaskan marine animals do not appear to be significantly more sensitive
to aromatic hydrocarbons than similar species from more temperate and
tropical climates. However, because of the greater persistence of light
aromatics and PAH in cold Alaskan waters* biological impacts of an Alaskan
oil spill may be more severe and subsequent recovery slower than for a
similar spill in a warmer climate. However* many marine communities in
Alaskan coastal environments are already naturally stressed by the severe
climatic conditions. Such communities recover rapidly following a
disturbance such as an oil spill or cessation of a chronic pollutant
discharge. The Alaskan marine populations most likely to be severely
damaged by oil spills and chronic discharges are the large, long-lived
species such as king crabs, salmon* and marine mammals.
Based on this review* we have made recommendations concerning
additional information needs and the design of long-term monitoring
programs. Specific information needs include: a quantitative inventory of
PAH sources in the Alaskan marine, environment; composition over time of
produced water and ballast water discharges and their long-term fate after
discharge to Alaskan coastal waters; sensitivity to and PAH metabolism by
populations and communities of marine animals from the high arctic (Chukchi
and Beaufort Seas); field validation of arctic oil spill models. The
design of long-term monitoring studies to assess the environmental impacts
of produced water discharges to Cook Inlet and of offshore oil and gas
development in the Beaufort Sea were discussed.
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