GUIDELINES FOR THE BIOREMEDIATION OF MARINE SHORELINES
AND FRESHWATER WETLANDS
Prepared by:
1Xueqing Zhu, 2Albert D. Venosa, 1Makram T. Suidan, and 3Kenneth Lee
1University of Cincinnati
Department of Civil and Environmental Engineering
Cincinnati, OH 45221
U.S. Environmental Protection Agency
National Risk Management Research Laboratory
Cincinnati, OH 45268
department of Fisheries and Oceans-Canada
Bedford Institute of Oceanography
Marine Environmental Sciences Division
Dartmouth, Nova Scotia B2Y 4A2
September, 2001
U.S. Environmental Protection Agency
Office of Research and Development
National Risk Management Research Laboratory
Land Remediation and Pollution Control Division
26 W. Martin Luther King Drive
Cincinnati, OH 45268
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Acknowledgments
The preparation of this guidance document was performed under the direction of
Albert D. Venosa of EPA's Treatment and Destruction Branch, Land Remediation and Pollution
Control Division, National Risk Management Research Laboratory, Cincinnati, OH. The report
was prepared under EPA Contract 68-C7-0057, Task Order 23 by the University of Cincinnati
with the assistance of the Department of Fisheries and Oceans-Canada.
Disclaimer
This report has been reviewed and approved for publication by the Land Remediation and
Pollution Control Division, National Risk Management Research Laboratory of the U.S.
Environmental Protection Agency. Mention of company names, trade names, or commercial
products does not constitute EPA endorsement or recommendations for use.
Further Information
For further information, contact:
Albert D. Venosa, Ph.D.
U.S. EPA
26 W. Martin Luther King Drive
Cincinnati, OH 45268
Tel: 513-569-7668
Fax: 513-569-7105
Email: venosa.albert@epa.gov
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EXECUTIVE SUMMARY
The objective of this document is to present a detailed technical guidance document for use by
spill responders for the bioremediation of marine shorelines and freshwater wetlands
contaminated with oil and oil products. Technical personnel who are responsible for designing
and operating field bioremediation processes as well as consultants and equipment manufacturers
will also find it useful. This manual presents a rational approach for the design of bioremediation
processes pertinent to cleanup of oil-contaminated marine shorelines and freshwater wetlands.
This document evaluates current practices and state-of-the-art research results pertaining to
bioremediation of hydrocarbon contamination relative to types and amounts of amendments
used, frequency of application, assessment of the extent of bioremediation, sampling, and
analysis. The scope of the document is limited to marine shorelines and freshwater wetlands
because of definitive results from recently completed, EPA-sponsored field studies. The final
product is presented in a report form that is understandable by responders, on-scene coordinators,
and remediation specialists. This report includes a thorough review and critique of the literature
and theories pertinent to oil biodegradation, nutrient dynamics in shorelines, and analytical
chemistry of oil and remediation nutrients.
A planning approach to site identification, evaluation, and selection along with information on
field investigations is also presented. The manual includes examples of bioremediation options
and case studies of bioremediation applied to marine shorelines and freshwater wetland
environments.
The contents of this document are arranged in a logical sequence first to provide basic
information for the evaluation of bioremediation as a spill response option followed by
guidelines for application that includes methods to monitor its effectiveness. Thus, Chapter 1
presents an overall introduction and background discussion of bioremediation including
occurrence of oil spills, response methodologies, and a summary of the scope, organization, and
objective of the manual. Chapter 2 covers the basic information about oil, shorelines,
mechanisms of oil biodegradation, and a state-of-the-art review of controlled laboratory
experiments and field trials of oil biodegradation and nutrient dynamics in shoreline
environments. For additional background information, Chapter 3 provides a more thorough
review and critique of current analytical methods used to monitor and verify oil spill
bioremediation success. Chapter 4 summarizes major biostimulatory and bioaugmenting
amendment methods and their application strategies. Chapter 5 is the heart of the document and
presents the actual guidelines for designing, planning, and implementing oil bioremediation in
the field, including site characterization, evaluation of appropriate bioremediation technologies,
and the selection of the most appropriate technology for a specific site. Finally, Chapter 6
provides guidelines for assessment and interpretation of field results and provides help in
assessing endpoints of bioremediation (i.e., when treatment is considered complete).
The overall conclusions reached by the guidance manual are as follows. First, with respect to
marine sandy shorelines, natural attenuation may be appropriate if background nutrient
concentrations were high enough that intrinsic biodegradation would take place at close to the
expected maximum rate. The Delaware study proved this clearly. Certainly in nutrient-limited
places like Prince William Sound, Alaska, nutrient addition should accelerate cleanup rates
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many-fold. However, the decision to use the natural attenuation approach may be tempered by
the need to protect a certain habitat or vital resource from the impact of oil. For example, using
Delaware as the model, every spring season, horseshoe crabs migrate to the shoreline of
Delaware for their annual mating season. Millions of eggs are laid and buried a few mm below
the surface of the sand. Migrating birds making their way from South America to Arctic Canada
fly by this area and feed upon these eggs to provide energy to continue their long flight. If an oil
spill occurred in February or March, it would certainly be appropriate to institute bioremediation
to accelerate the disappearance of the oil prior to the horseshoe crab mating season despite the
expected high natural attenuation rate. So, even in the case where background nutrients are high
enough to support rapid biodegradation, addition of more nutrients would help protect such a
vital resource. If the spill occurred during the summer, and no vital natural resources were
threatened by the spill, then reliance on natural attenuation might be the wisest course of action.
Of course, removal of free product and high concentrations of oil should still be conducted by
conventional means even if a no bioremediation action is warranted by the circumstances.
With respect to freshwater wetlands, the St. Lawrence River study demonstrated that, if
significant penetration of oil takes place into the subsurface, biodegradation would take place
very slowly and ineffectively. This is because of the anaerobic conditions that quickly occur in
these types of saturated environments, and anaerobic biodegradation of petroleum oils is much
slower and less complete than under aerobic conditions. One of the objectives of the St.
Lawrence River experimental design was to determine the amenability of wetlands to
biodegradation when oil has penetrated into the sediment. The oil was artificially raked into the
sediment to mimic such an occurrence. Consequently, no significant treatment effects were
observed because all the nutrients in the world would not stimulate biodegradation if oxygen
were the primary limiting material. If penetration did not take place beyond a few mm, then
bioremediation might be an appropriate cleanup technology, since more oxygen would be
available near the surface. It is clear that whatever oxygen gets transported to the root zone by
the plants is only sufficient to support plant growth and insufficient to support the rhizosphere
microorganisms to degrade contaminating oil.
However, if ecosystem restoration is the primary goal rather than oil cleanup, the St. Lawrence
River study strongly suggested that nutrient addition would accelerate and greatly enhance
restoration of the site. Abundant plant growth took place in the nutrient-treated plots despite the
lack of oil disappearance from the extra nutrients. Furthermore, the stimulation lasted more than
one growing season even though nutrients were never added after the first year. Clearly, the
plants took up and stored the extra nitrogen for use in subsequent growing seasons, so restoration
of the site was abundantly evident in a few short months.
Thus, in conclusion, the decision to bioremediate a site is dependent on cleanup, restoration, and
habitat protection objectives and whatever factors that are present that would have an impact on
success. Responders must take into consideration the oxygen and nutrient balance at the site. If
the circumstances are such that no amount of nutrients will accelerate biodegradation, then the
decision should be made on the need to accelerate oil disappearance to protect a vital living
resource or simply to speed up restoration of the ecosystem. If there is no immediate need to
protect a vital resource or restore the ecosystem, then natural attenuation may be the appropriate
response action. These decisions are clearly influenced by the circumstances of the spill.
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TABLE OF CONTENTS
INTRODUCTION
Occurrence of Oil Spills
Response to Oil Spills to Marine Shorelines and
Freshwater Environments
1.2.1 Natural methods
1.2.2 Physical methods
1.2.3 Chemical methods
Bioremediation as an Oil Spills Cleanup Technology
Scope of This Document
1.4.1 Objectives
1.4.2 Organization of the guidance document
FACTORS AFFECTING NATURAL OIL BIODEGRADATION
AND BIOREMEDIATION SUCCESS
Physical-Chemical Properties of Crude Oil and Oil Products
2.1.1 Chemical composition of crude oils and oil products
2.1.2 Physical properties of oil
Behavior of Oil in the Environment
2.2.1 Weathering processes
2.2.2 Oil and shoreline interactions
2.2.3 Shoreline sensitivities
Biodegradation of Oil
2.3.1 Mechanism of oil biodegradation: a microbiological perspective
2.3.2 Environmental factors affecting oil biodegradation
2.3.3 Evaluation of oil biodegradation: application of biomarkers
Laboratory Studies on Bioremediation of Oil
2.4.1 Bioaugmentation
2.4.2 Biostimulation
Demonstrations of Oil Bioremediation Under Field Conditions
2.5.1 Mesocosm studies
2.5.2 Field demonstration
2.5.3 Kinetics of oil bioremediation
Nutrient Hydrodynamics
2.6.1 Nutrient transport in beaches: a mesocosm study
2.6.2 Nutrient transport in beaches: field trials
METHODS USED IN MONITORING OIL BIOREMEDIATION
Analytical Methods
3.1.1 Microbiological analysis
3.1.1.1 Enumeration of hydrocarbon-degrading microorganisms:
culture based techniques
3.1.1.2 Culture-independent population/community techniques
3.1.2 Chemical analysis of nutrients
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3.1.3 Chemical analysis of oil and oil constituents 51
3.1.3.1 Total petroleum hydrocarbon (TPH) techniques 52
3.1.3.2 Analysis of specific oil constituents 53
3.2 Biomarkers 57
3.2.1 Commonly used biomarkers 57
3.2.2 The effect of contaminant redistribution
on observed remediation rates 61
3.3 Sampling in the Field 63
3.3.1 Sampling strategies 64
3.3.2 Field sampling experiences 65
3.4 Monitoring General Site Background Conditions 66
3.5.1 Oxygen 67
3.5.2 pH 67
3.5.3 Temperature 67
3.5.4 Salinity 68
3.5 Monitoring of Biological Impacts 68
3.5.1 Bioassessment 68
3.5.2 Bioassays 71
3.5.2.1 Benthic invertebrates 72
3.5.2.2 Microtox 72
3.5.2.3 Fish 73
3.5.3 Application of Bioassays to Assess Bioremediation
in Marine Environments 74
Chapter 4 TYPES OF AMENDMENTS AND CONSIDERATIONS
IN THEIR APPLICATIONS 75
4.1 Nutrient Amendment 75
4.1.1 Water-soluble nutrients 75
4.1.2 Granular Nutrients (slow-release) 76
4.1.3 Oleophilic Nutrients 76
4.2 Microbial Amendments 78
4.3 Plant Amendments (Phytoremediation) 80
4.3.1 Mechanisms of phytoremediation 80
4.3.2 Considerations in application of oil phytoremediation 81
4.3.3 Applications in marine shoreline and freshwater wetlands 82
4.4 Oxygen Amendment 83
4.4.1 Tilling 84
4.4.2 Forced aeration 84
4.4.3 Chemical methods 84
Chapter 5 GUIDELINES FOR BIOREMEDIATION OF MARINE SHORELINES
AND FRESHWATER WETLANDS: DECISION-MAKING AND
PLANNING 86
5.1 Pre-treatment Assessment 88
5.1.1 Oil type and concentration 88
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5.1.2 B ackground nutri ent content 89
5.1.3 Type of shorelines 90
5.1.4 Other factors 92
5.1.5 Summary of pretreatment assessment 92
5.2 Selection of Nutrient Products 93
5.2.1 Nutrient selection based on efficacy and toxicity 93
5.2.2 Environmental factors affecting nutrient selection 97
5.3 Determination of the Optimal Nutrient Loading and Application Strategy 97
5.3.1 Concentration of nutrients needed for optimal biostimulation 97
5.3.2 Nutrient application strategies 99
5.4 Sampling and Monitoring Plan 101
5.4.1 Important variables 101
5.4.2 Statistical considerations in sampling plan 103
5.5 Considerations for Freshwater Wetland Bioremediation 104
5.5.1 Characteristics of freshwater wetlands 104
5.5.2 Bioremediation strategies in freshwater wetlands 106
Chapter 6 GUIDELINES FOR ASSESSMENT OF FIELD RESULTS AND
TERMINATION OF TREATMENT 109
6.1 Assessment of Oil Biodegradation Efficacy 109
6.1.1 Verification of oil biodegradation 109
6.1.2 Assessment of physical removal 113
6.1.3 Operational endpoints based on oil biodegradation 115
6.2 Environmental Assessments 115
6.2.1 Operational guidance from environmental assessments
for treatment application 116
6.3 Case Study: Environmental Assessment of Bioremediation Treatments
in a Tidal Freshwater Marsh 117
6.3.1 Alterations in ecosystem structure 117
6.3.2 Alterations in ecosystem function 118
6.3.2.1 Microbial response 118
6.3.2.2 Microtox solid phase test 119
6.3.2.3 Algal solid phase assay 119
6.3.2.4 Cladoceran survival test 121
6.3.2.5 Amphipod survival test 121
6.3.2.6 Gastropod survival/histopathology 122
6.3.2.7 Acute and chronic effects on fish 124
6.4 Ecotoxicological Tests for Risk Assessment 128
6.5 Ecotoxicological Tests to Identify Operational Endpoints 129
REFERENCES 130
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Chapter 1 INTRODUCTION
1.1 Occurrence of Oil Spills
Modern society continues to rely on the use of petroleum hydrocarbons for its energy needs.
Despite recent technological advances, accidental spills of crude oil and its refined products
occur on a frequent basis during routine operations of extraction, transportation, storage, refining
and distribution. It is estimated that between 1.7 and 8.8 million metric tons of oil are released
into the world's water every year (NAS, 1985), of which more than 90% is directly related to
human activities including deliberate waste disposal. Contrary to popular perception, only one
eighth of the oil released into the aquatic environment is from tanker accidents. It is also
estimated that about 30% of the spilled oil enters freshwater systems (Cooney, 1984). These
figures are rather uncertain and can vary greatly from year to year depending on sources of
estimation and spill incidents. Table 1.1 summarizes the number of oil spills and the quantities
of release from 1970 to 1999 based on International Tanker Owners Pollution Federation's oil
spill database (ITOPF, 2001). These data include oil spills of over seven tons from tankers,
combined carriers, and barges. Although the data suggest a reduction in oil spills, the trend may
only represent a temporary downward fluctuation that is part of erratic cycling over the long term
(Etkin and Welch, 1997).
Marine shorelines are important public and ecological resources that serve as a home to a variety
of wildlife and provide public recreation. Marine oil spills, particularly large scale spill
accidents, have posed great threats and cause extensive damage to the marine coastal
environments. For example, the spill of 37,000 metric tons (11 million gallons) of North Slope
crude oil into Prince William Sound, Alaska, from the Exxon Valdez in 1989 led to the mortality
of thousands of seabirds and marine mammals, a significant reduction in population of many
intertidal and subtidal organisms, and many long term environmental impacts (Spies el al.,
1996). In 1996, the Sea Empress released approximately 72,000 tons of Forties crude oil and 360
tons of heavy fuel oil at Milford Haven in South Wales and posed a considerable threat to local
fisheries, wildlife and tourism (Edwards and White, 1999; Harris; 1997).
Compared to marine oil spills, inland oil spills have received much less attention. However,
freshwater spills are very common, with more than 2000 oil spills, on average, taking place each
year in the inland waters of the continental United States (Owens et al., 1993). Although
freshwater spills tend to be of a smaller volume than their marine counterparts (Stalcup, et al.,
1997), they have a greater potential to endanger public health and the environment because they
often occur within populated areas and may directly contaminate surface water and groundwater
supplies. For example, in 1988, an Ashland Oil Company storage tank in Pittsburgh ruptured
and spilled about 2,500 tons (750,000 gallons) of diesel oil in the Monongahela River, which
contaminated drinking water intakes and led to downstream water shortages as far as 200 miles
(Miklaucic and Saseen, 1989).
These catastrophic accidents, especially the Exxon Valdez spill, have increased public awareness
about the risks involved in the storage and transportation of oil and oil products and have
prompted more stringent regulations, such as the enactment of the 1990 Oil Pollution Act by
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that (
counl
Tabk
Spill
Year
1970
1971
1972
1973
1974
1975
1976
1977
1978
1979
1980
1981
1982
1983
1984
1985
1986
1987
1988
1989
1990
1991
1992
1993
1994
1995
1996
1997
1998
1999
wever, because oil is so widely used, despite all the precautions, it is almost certain
and leakage will continue to occur. Thus, it is essential that we have effective
*es to deal with the problem.
spills from tankers and other vessels into world water (adapted from ITOPF's Oil
:,_httg^^^witogfcom/sMs;htm^_
Number of Oil Spills
<700 metric tons
> 700 metric tons
6
29
18
14
49
24
25
32
91
26
97
19
67
25
65
16
54
23
59
34
51
13
49
6
44
3
52
11
25
8
29
8
25
7
27
10
11
10
32
13
50
13
27
8
31
9
30
11
27
7
21
2
20
3
27
10
22
4
19
5
Quantities of Oil Spills
(x 103 metric tons)
301
167
311
166
222
342
369
298
395
608
206
44
11
384
28
88
19
30
198
178
61
435
162
144
105
9
79
67
10
24
2
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1.2 Response to Oil Spills in Marine Shorelines and Freshwater Environments
Strategies for cleaning up an oil spill are greatly affected by a variety of factors, such as the type
of oil, the characteristics of the spill site, and occasionally political considerations. A number of
approaches and technologies have been developed for controlling oil spills in marine shorelines
and freshwater environments. These methods have been reviewed and described extensively in a
number of technical documents, such as: Shoreline Countermeasure Manual (NOAA, 1992),
Options for Minimizing Environmental Impacts of Freshwater Spill Response (NOAA and API,
1994), Understanding Oil Spills and Oil Spill Response (U.S. EPA, 1999), and Oil Spill
Response in the Marine Environment (Doerffer, 1992). The most commonly used shoreline
cleanup options (Table 1.2) are briefly described in the following text.
Table 1.2. Conventional Shoreline Clean-up Options
Category of Response Options
Example Technology
Natural method
Natural attenuation
Physical method
Booming
Skimming
Manual removal (Wiping)
Mechanical removal
Washing
Sediment relocation/Surf-washing
Tilling
In-situ burning
Chemical method
Dispersants
Demulsifiers
Solidifiers
Surface film chemicals
1.2.1 Natural methods
Natural attenuation or natural recovery is basically a no-action option that allows oil to be
removed and degraded by natural means. For some spills, it is probably more cost-effective and
ecologically sound to leave an oil-contaminated site to recover naturally than to attempt to
intervene. Examples of such cases are spills at remote or inaccessible locations when natural
removal rates are fast, or spills at sensitive sites where cleanup actions may cause more harm
than good. It should also be noted that when natural attenuation is used as a clean up method, a
monitoring program is still required to assess the performance of natural attenuation. Major
natural processes that result in the removal of oils include:
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• Evaporation: Evaporation is the most important natural cleansing process during the early
stages of an oil spill, and it results in the removal of lighter-weight components in oil.
Depending on the composition of the oil spilled, up to 50 percent of the more toxic, lighter
weight components of an oil may evaporate within the first 12 hours following a spill (U.S.
EPA, 1999).
• Photooxidation: Photooxidation occurs when oxygen under sunlight reacts with oil
components. Photooxidation leads to the breakdown of more complex compounds into
simpler compounds that tend to be lighter in weight and more soluble in water, allowing
them to be removed further through other processes.
• Biodegradation: Various types of microorganisms that are capable of oxidizing petroleum
hydrocarbons are widespread in nature. Biodegradation is a particularly important
mechanism for removing the non-volatile components of oil from the environment. This is a
relatively slow process and may require months to years for microorganisms to degrade a
significant fraction of an oil stranded within the sediments of marine and/or freshwater
environments.
1.2.2 Physical methods
Physical containment and recovery of bulk or free oil is the primary response option of choice in
the United States for the cleanup of oil spills in marine and freshwater shoreline environments.
Commonly used physical methods include:
• Booming and skimming: Use of booms to contain and control the movement of floating oil
and use of skimmers to recover it. The environmental impact of this method is minimal if
traffic of the cleanup work force is controlled.
• Wiping with absorbent materials: Use of hydrophobic materials to wipe up oil from the
contaminated surface. While the disposal of contaminated waste is an issue, the
environmental effect of this method is also limited if traffic of cleanup crew and waste
generation is controlled.
• Mechanical removal: Collection and removal of oiled surface sediments by using
mechanical equipment. This method should be used only when limited amounts of oiled
materials have to be removed. It should not be considered for cleanup of sensitive habitats or
where beach erosion may result.
• Washing: washing of the oil adhering along the shorelines to the water's edge for collection.
Washing strategies range from low-pressure cold water flushing to high-pressure hot water
flushing. This method, especially using high-pressure or hot water, should be avoided for
wetlands or other sensitive habitats.
• Sediment relocation and tilling: Movement of oiled sediment from one section of the beach
to another or tilling and mixing the contaminated sediment to enhance natural cleansing
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processes by facilitating the dispersion of oil into the water column and promoting the
interaction between oil and mineral fines. Tilling may cause oil penetration deep into the
shoreline sediments. The potential environmental impacts from the release of oil and oiled
sediment into adjacent water bodies should also be considered.
• In-situ burning: Oil on the shoreline is burned usually when it is on a combustible substrate
such as vegetation, logs, and other debris. This method may cause significant air pollution
and destruction of plants and animals.
1.2.3 Chemical methods
Chemical methods, particularly dispersants, have been routinely used in many countries as a
response option. For some countries, such as the United Kingdom, where rough coastal
conditions may make mechanical response problematic, dispersants are the primary choice
(Lessard and Demarco, 2000). However, chemical methods have not been extensively used in
the United States due to the disagreement about their effectiveness and the concerns of their
toxicity and long-term environmental effects (U.S. EPA, 1999). Major existing chemical agents
include:
• Dispersants: dispersing agents, which contain surfactants, are used to remove floating oil
from the water surface to disperse it into the water column before the oil reaches and
contaminates the shoreline. This is done to reduce toxicity effects by dilution to benign
concentrations and accelerate oil biodegradation rates by increasing its effective surface area.
• Demulsifiers: Used to break oil-in-water emulsions and to enhance natural dispersion.
• Solidifiers: Chemicals that enhance the polymerization of oil can be used to stabilize the oil,
to minimize spreading, and to increase the effectiveness of physical recovery operations.
• Surface film chemicals: Film-forming agents can be used to prevent oil from adhering to
shoreline substrates and to enhance the removal of oil adhering to surfaces in pressure
washing operations.
1.3 Bioremediation as an Oil Spill Cleanup Technology
Although conventional methods, such as physical removal, are the first response option, they
rarely achieve complete cleanup of oil spills. According to the Office of Technology
Assessment (OTA, 1990), current mechanical methods typically recover no more than 10-15
percent of the oil after a major spill. Bioremediation has emerged as one of the most promising
secondary treatment options for oil removal since its successful application after the 1989 Exxon
Valdez spill (Bragg et al., 1994; Prince el. al., 1994). Bioremediation has been defined as "the act
of adding materials to contaminated environments to cause an acceleration of the natural
biodegradation processes" (OTA, 1991). This technology is based on the premise that a large
percentage of oil components are readily biodegradable in nature (Atlas, 1984, 1981; Prince,
1993). The success of oil spill bioremediation depends on our ability to establish and maintain
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conditions that favor enhanced oil biodegradation rates in the contaminated environment. There
are two main approaches to oil spill bioremediation:
• bioaugmentation, in which known oil-degrading bacteria are added to supplement the
existing microbial population, and
• biostimulation, in which the growth of indigenous oil degraders is stimulated by the addition
of nutrients or other growth-limiting cosubstrates, and/or by alterations in environmental
conditions (e.g. surf-washing, oxygen addition by plant growth, etc.).
Both laboratory studies and field tests have shown that bioremediation, biostimulation in
particular, can enhance oil biodegradation on contaminated shorelines (Prince, 1993; Swannell,
et al., 1996). Recent field studies have also demonstrated that biostimulation is a more effective
approach because the addition of hydrocarbon degrading microorganisms will not enhance oil
degradation more than simple nutrient addition (Lee et al, 1997a; Venosa et al., 1996; see
Chapter 2 in detail). Bioremediation has several advantages over conventional technologies.
First, the application of bioremediation is relatively inexpensive. For example, during the
cleanup of the Exxon Valdez spill, the cost of bioremediating 120 km of shoreline was less than
one day's costs for physical washing (Atlas, 1995). Bioremediation is also a more
environmentally benign technology since it involves the eventual degradation of oil to mineral
products (such as carbon dioxide and water), while physical and chemical methods typically
transfer the contaminant from one environmental compartment to another. Since it is based on
natural processes and is less intrusive and disruptive to the contaminated site, this "green
technology" may also be more acceptable to the general public.
Bioremediation like other technologies also has its limitations. Bioremediation involves highly
heterogeneous and complex processes. The success of oil bioremediation depends on having the
appropriate microorganisms in place under suitable environmental conditions. Its operational
use can be limited by the composition of the oil spilled. Bioremediation is also a relatively slow
process, requiring weeks to months to take effect, which may not be feasible when immediate
cleanup is demanded. Concerns also arise about potential adverse effects associated with the
application of bioremediation agents. These include the toxicity of bioremediation agents
themselves and metabolic by-products of oil degradation and possible eutrophic effects
associated with nutrient enrichment (Swannell et al., 1996). Bioremediation has been proven to
be a cost-effective treatment tool, if used properly, in cleaning certain oil-contaminated
environments. Few detrimental treatment effects have been observed in actual field operations.
Currently, one of the major challenges in the application of oil bioremediation is the lack of
guidelines regarding when and how to use this technology. Although extensive research has
been conducted on oil bioremediation during the last decade, most existing studies have
concentrated on either evaluating the feasibility of bioremediation for dealing with oil
contamination, or testing favored products and methods (Mearns, 1997). Only limited number of
pilot-scale and field trials, which may provide the most convincing demonstrations of this
technology, have been carried out. To make matters worse, many field tests have not been
properly designed, well controlled or correctly analyzed, leading to skepticism and confusion
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among the user community (Venosa, 1998). There is an urgent need for a detailed and workable
set of guidelines for the application of this technology for oil spill responders that answers
questions such as when to use bioremediation, what bioremediation agents should be used, how
to apply them, and how to monitor and evaluate the results. Scientific data for the support of an
operational guidelines document has recently been provided from laboratory studies and fields
trials carried out by the U.S. Environmental Protection Agency (EPA), University of Cincinnati,
and Fisheries and Oceans Canada.
1.4 Scope of This Document
1.4.1 Objectives
The objective of this manual is to provide detailed and applicable technical guidelines for use by
spill responders for the bioremediation of marine shorelines and freshwater wetlands that are
contaminated with crude oil and its refined products. The document will also provide guidance to
scientists, other technical personnel, and manufacturers with an interest in the design and
implementation of field bioremediation processes. The document evaluates current practices and
state-of-the-art research results pertaining to bioremediation of hydrocarbon contamination and
presents a rational procedure for the design of bioremediation processes pertinent to clean-up of
oil stranded within sediments of shoreline environments.
The scope of this document is limited to marine shorelines and freshwater wetlands where the
effectiveness of bioremediation has been quantified in controlled field and case studies following
spill incidents. To date, there is no conclusive evidence of successful oil spill clean up
operations in the open sea by the application of bioremediation strategies (Atlas, 1995; Swannell
et al., 1996). With respect to freshwater environments, an oil spill is most likely to have the
greatest impact on sensitive wetlands or marshes rather than running rivers. To provide insight
into the feasibility of bioremediation strategies in these habitats, the information presented in this
manual is based on experimental oil spills conducted in the 1990s jointly by EPA, the University
of Cincinnati, and Fisheries and Oceans-Canada. The first study was a field experiment on a
sandy beach in Delaware in 1994. In 1999 and 2000, a similar field investigation was conducted
on a freshwater wetland along the shoreline of the St. Lawrence River, Quebec, Canada. Another
field study was conduced in Dartmouth, Nova Scotia in 2000 on a salt marsh dominated by
Spartina alterniflora. This document will provide oil bioremediation guidelines based on current
practices and research with the emphasis on the findings of these fields studies. Because the
study conducted on the shoreline of Nova Scotia has not been concluded at the time of this
writing, guidelines of oil bioremediation in salt marshes will be available as a supplementary
document upon the conclusion of this investigation.
1.4.2 Organization of the guidance document
For ease of use, the contents of this document are arranged in a logical sequence first to provide
basic information for the evaluation of bioremediation as a spill response option followed by
guidelines for application that includes methods to monitor its effectiveness. Thus, Chapter 2
covers the basic information about oil, shorelines, mechanisms of oil biodegradation, and a state-
7
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of-art review of controlled laboratory experiments and field trials of oil biodegradation and
nutrient dynamics in shoreline environments. For additional background information, Chapter 3
provides a more thorough review and critique of current analytical methods used to monitor and
verify oil spill bioremediation success. Chapter 4 summarizes major biostimulatory and
bioaugmenting amendment methods and their application strategies. Chapter 5 presents
guidelines for designing, planning, and implementing oil bioremediation in the field, including
site characterization, evaluation of appropriate bioremediation technologies, and the selection of
the most appropriate technology for a specific site. Finally, Chapter 6 provides guidelines for
assessment and interpretation of field results.
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Chapter 2 FACTORS AFFECTING NATURAL OIL BIODEGRADATION AND
BIOREMEDIATION SUCCESS
Oil bioremediation is a complex process involving interactions of oil and microorganisms under
the conditions of the prevailing environment. To understand the scope and strategies of oil
bioremediation, it is essential to first understand the properties of oil, the environment of concern
(e.g., marine shorelines and freshwater wetlands), the fate of oil in that environment, the
mechanisms of oil biodegradation and the factors that control its rate.
2.1 Physical-Chemical Properties of Crude Oil and Oil Products
Crude oil and petroleum products are very complex and variable mixtures of thousands of
individual compounds that exhibit a wide range of physical properties. Understanding these
properties is important in determining behavior of spilled oil and the appropriate response option.
The composition and properties of various petroleum hydrocarbons have been described in detail
by Clark and Brown (1977) and the National Academy of Sciences (1985). Large oil property
databases also exist such as the one posted on the Internet by Environment Canada
(www.etcenttre.org/spills). which contains information on over 400 oils (Jokuty, et al., 2000).
2.1.1 Chemical composition of crude oils and oil products
Crude Oil
Crude oil is comprised of both hydrocarbon compounds (accounting for 50-98% of total
composition) and non-hydrocarbon compounds (containing sulfur, nitrogen, oxygen, and various
trace metals) in a wide array of combinations (Clark and Brown, 1977). The chemical
composition and physical characteristics of several crude oils is illustrated in Table 2.1. Some
representative organic compounds found in crude oil are illustrated in Figure 2.1. Petroleum
components may be classified into four major groups based on their differential solubility in
organic solvents (Leahy and Colwell, 1990).
1. Saturated hydrocarbons: Include normal and branched alkanes with structures of CnH2n+2
(aliphatics) and cyclic alkanes with structures of CnH2n (alicyclics), which range in chain
length from one carbon to over 40 carbons. Saturates usually are the most abundant
constituents in crude oils.
2. Aromatic hydrocarbons: Include monocyclic aromatics (e.g., benzene, toluene, and xylenes)
and polycyclic aromatic hydrocarbons (PAHs) (e.g., naphthalene, anthracene, and
phenanthrene), which have two or more fused aromatic rings. PAHs are of particular
environmental concern because they are potential carcinogens or may be transformed into
carcinogens by microbial metabolism.
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3. Resins: Include polar compounds containing nitrogen, sulfur, and oxygen (e.g., pyridines
and thiophenes). They are often referred to as NSO compounds.
4. Asphaltenes: Consist of poorly characterized high molecular weight compounds that include
both high molecular weight and poorly characterized hydrocarbons and NSOs. Metals such
as nickel, vanadium, and iron are also associated with asphaltenes.
Table 2.1 Chemical composition and physical properties of representative crude oils
(adapted from Clark and Brown, 1977)
Characteristic or Component Prudhoe Bay South Louisiana Kuwait
API gravity (20°C) 27.8 34.5 31.4
Sulfur (wt%) 0.94 0.25 2.44
Nitrogen (wt%) 0.23 0.69 0.14
Nickel (ppm) 10 2.2 7.7
Vanadium (ppm) 20 1.9 28
Naphtha fraction (wt%)a 23.2 18.6 28.0
Saturates 19.9 16.5 20.3
Aromatics 3.2 2.1 2.4
Resins & Asphaltenes
High-boiling fraction (wt%)6 76.8 81.4 77.3
Saturates 47.7 56.3 34.0
Aromatics 25.0 16.5 21.9
Resins & Asphaltenes 4.1 8.6 21.4
These analyses represent values for one typical crude oil from three distinct geographical
regions; variations in composition can be expected for oils produced from different formations or
fields within each region.
a Fraction boiling from 20° to 205°C
b Fraction boiling above 205°C
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SATURATES
AROMATICS
h3c—c—c—c—c—ch3
h2 h2 h2 h2
n-hexane
n-heptadecane (n-Q7H36)
PH3
toluene
naphthalene
pristane (C19H40)
17a(H),21 (3(H)- hopane
chrysene
benzo |a|pyrene
RESINS
ASPHATTENES
l\k ^CH3
2-methylpyridine
(J
(C79I i92n2 s2o)3
dibenzo-
thiophene
Figure 2.1 Representative organic compounds found in crude oils
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Refined oil products
Refined petroleum products, such as gasoline, kerosene, jet fuels, fuel oils, and lubricating oils,
are derived from crude oils through processes such as catalytic cracking and fractional
distillation. These products have physical and chemical characteristics that differ according to
the type of crude oil and subsequent refining processes. They contain components of crude oil
covering a narrow range of boiling points. In addition, during catalytic cracking operations,
unsaturated compounds, or olefins (alkenes and cycloalkenes), which are not present in crude
oils, can be formed. The concentrations of olefins are as high as 30% in gasoline and about 1%
in jet fuel (NAS, 1985). A list of chemical compositions of the fractions of crude oils and the
refined products is shown in Table 2.2.
Table 2.2 Chemical compositions of refined petroleum products (adapted from Clark and
Brown, 1977)
Distillation Hydrocarbon Range of Typical Refined
Fraction Types Carbon Atoms Products
Gasoline & naphtha Saturates 4-12 Gasoline
Olefins
Aromatics
Middle distillate
Wide-cut gas oil
Saturates
Olefins
Aromatics
Saturates
Aromatics
10-20
18-45
Kerosene
Jet fuel
Heating oils
Diesel oils
Wax
Lubricating oil
Residum
Resins
Asphaltenes
>40
Residual oils
Asphalt
2.1.2 Physical properties of oil
Important physical properties of oil that affect its behavior in the environment and spill cleanup
responses include:
1. Density: Two types of density expressions for oils are often used: specific gravity and
American Petroleum Institute (API) gravity. Specific gravity is the ratio of the mass of a
substance to the mass of the equivalent volume of water at a specified temperature. The API
gravity arbitrarily assigns a value of 10° to pure water at 10°C (60°F). The API gravity can
be calculated from the specific gravity using the formula:
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API Gravity (°) = 141.5 -131.5
Specific Gavity (16 °C / 60 °F)
(2.1)
Oils with low densities or low specify gravities have high API gravities. Crude oils have
specific gravities in the range of 0.79 to 1.00 (equivalent to API Gravities of 10 to 48) (Clark
and Brown, 1977). Oil density is an important index of oil composition that is frequently
used to predict its fate in water.
2. Viscosity: Viscosity is the property of a fluid that describes how it resists a change in shape
or movement. The lower the viscosity a fluid has, the more easily it flows. The viscosity of
petroleum is related to oil compositions and the ambient temperature. It is an important
index of the spreading rate of a spilled oil.
3. Pour Point: The pour point of an oil is the temperature at which it becomes semi-solid or
stops flowing. The pour point of crude oils varies from -57°C to 32°C. It is another
important characteristic with respect to oil fate and cleanup strategies.
4. Solubility in water: The solubility of oil in water is extremely low and depends on the
chemical composition of the petroleum hydrocarbon in question and temperature. For a
typical crude oil, solubility is around 30 mg/L (NAS, 1985). The most soluble oil
components are the low molecular weight aromatics such as benzene, toluene and xylene.
This property is important with respect to oil fate, oil toxicity and bioremediation processes.
Other important physical properties of oils include flash point, vapor pressure, surface tension,
and adhesion.
2.2 Behavior of Oil in the Environment
2.2.1 Weathering processes
When oil is introduced into the environment, it immediately goes through a variety of physical,
chemical and biological changes (Figure 2.2). These weathering processes will alter oil
composition and properties in ways that may affect spill response strategies. Bioremediation is a
relatively slow process that is often used as a polishing step after conventional cleanup options
have been applied. Thus the residual target oil may be extensively weathered prior to the
deployment of bioremediation strategies.
Weathering processes, including biodegradation, have been reviewed and described extensively
in the literature (Clark and MacLeod, 1977; Jordan, R.E. and Payne, J.R., 1980; National
Academy of Sciences, 1985). Major physical and chemical fates of oil are briefly summarized in
this section and the biological fate will be discussed in section 2.3.
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Spreading
The spreading of oil on water is one of the most important processes during the first hours of a
spill, provided that the oil pour point is lower than the ambient temperature. The principal forces
influencing the spreading of oil include gravity, inertia, friction, viscosity and surface tension.
This process increases the overall surface area of the spill, thus enhancing mass transfer via
evaporation, dissolution, and later biodegradation.
Evaporation
In terms of environmental impacts, evaporation is the most important weathering process during
the early stages of an oil spill in that it can be responsible for the removal of a large fraction of
the oil including the more toxic, lower molecular weight components. For oil on water,
evaporation removes virtually all the normal alkanes smaller than Ci5 within 1 to 10 days.
Volatile aromatic compounds, such as benzene and toluene, can also be rapidly removed from an
oil slick through evaporation. However, these volatile oil components may be more persistent
when oil is stranded in sediments. The volatile components make up 20-50% of most crude oils,
about 75% of No. 2 fuel oil, and about 100% of gasoline and kerosene. As a result, the physical
properties of the remaining slick change significantly (e.g., increased density and viscosity).
Major factors influencing the rate of evaporation include composition and physical properties of
the oil, wave action, wind velocity, and water temperature (Clark and MacLeod, 1977; Jordan,
R.E. and Payne, J.R., 1980).
Dissolution
Although dissolution is less important from the viewpoint of mass loss during an oil spill,
dissolved hydrocarbon concentrations in water are particularly important due to their potential
influence on the success of bioremediation and the effect of toxicity on biological systems. The
extent of dissolution depends on the solubility of the spilled oil, weather conditions, and the
characteristics of the spill site. The low molecular weight aromatics are the most soluble oil
components, and they are also the most toxic components in crude and refined oils. Although
many of them may be removed through evaporation, their impact on the environment is much
greater than simple mass balance considerations would imply (NAS, 1985). Dissolution rates are
also influenced by photochemical and biological processes.
Photooxidation
Photooxidation is another weathering process that may have important biological consequences.
In the presence of oxygen, natural sunlight has sufficient energy to transform many complex
petroleum compounds such as high molecular weight aromatics and polar compounds into
simpler compounds through a series of free-radical chain reactions. This process may increase
the solubility of oil in water, due to the formation of polar compounds such as hydroperoxides,
aldehydes, ketones, phenols, and carboxylic acids. Detrimental effects may be associated with
this increase in the solubility of oil in water (i.e., bioavailability) and the formation of toxic
compounds mediated by photooxidation. On the other hand, the formation of polar compounds
may increase the rate of biodegradation of petroleum, particularly at lower concentrations where
acute toxicity effects are limited (Nicodem et al. 1997).
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Dispersion
Dispersion, or formation of oil-in-water emulsions, involves incorporating small droplets of oil
into the water column, resulting in an increase in surface area of the oil. In general, oil-in-water
emulsions are not stable. However, they can be maintained by continuous agitation, interaction
with suspended particulates, and the addition of chemical dispersants. Dispersion may influence
oil biodegradation rates by increasing the contact between oil and microorganisms and/or by
increasing the dissolution rates of the more soluble oil components.
Emulsification
The process of emulsification of oils involves a change of state from an oil-on-water slick or an
oil-in-water dispersion to a water-in-oil emulsion, with the eventual possible formation of a
thick, sticky mixture that may contain up to 80% water, commonly called "chocolate mousse".
The formation and stability of emulsions are primarily related to the chemical composition of the
oils and are enhanced by wax and asphaltic materials. Surface-active materials generated
through photochemical and biological processes are also involved in formation of the emulsions.
The formation of emulsions makes oil clean-up operations more difficult by decreasing the
effectiveness of physical oil spill recovery procedures and suppressing the natural rates of oil
biodegradation.
Other important physical and chemical weathering processes that influence the rates of oil
degradation include adsorption onto suspended particulate materials, sinking and sedimentation,
and tar ball formation.
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2.2.2 Oil and shoreline interactions
When oil spills occur in marine or freshwater shoreline (e.g., freshwater wetlands) environments,
interactions between the spilled oil and the shore further complicate the weathering processes. It
is very important to understand these interactions in determining the scope and limitations of oil
spill bioremediation.
The behavior of spilled oil in shoreline environments is primarily dependent on the properties of
the shoreline, such as the porosity of the substrate and the energy of the waves acting on a
shoreline. Higher wave exposure enhances both physical removal and weathering processes.
Wave-swept rocky shores tend to recover from oil spills within a matter of months whereas
mangroves and marshes may act as a petroleum sink for many years. However, tidal pumping is
also a factor promoting oil penetration into the sediments. The rate and depth of oil penetration
depend primarily on the porosity of the substrate. On coarse-grained shorelines like cobble and
sandy beaches, oil can penetrate deeper and remain longer (when it is trapped below the limit of
wave action), compared to finer grained sediments such as silts and clay. However, oil is more
easily removed by water flushing from coarse-grained sediments. Interactions of oil with tidal
action, waves, and shoreline substrate may also form asphalt-like oil-sediment mats that are
resistant to further biological and photochemical weathering.
Recent studies have shown that the interactions between oil and fine mineral particles also play
an important role in natural oil cleansing in marine shorelines (Bragg and Owens, 1995; Lee et
al., 1997b). This process of oil and fine-particle interaction reduces the adhesion of oil to
intertidal shoreline substrates through the formation of oil-mineral fine floes that are easily
dispersed by tidal action and currents. More importantly, oil-mineral fine floes enhance the
availability of oil for biodegradation, and thus oil biodegradation rates are accelerated by this
process (Lee et al., 1997c).
2.2.3 Shoreline sensitivities
Marine shorelines and freshwater environments have a wide range of sensitivities to oil and
clean-up activities. The National Oceanic & Atmospheric Administration and the American
Petroleum Institute have developed the Environmental Sensitivity Index (ESI) to classify
shoreline types for spill response (NOAA, 1992; NOAA and API, 1994). This classification
scheme (e.g., Table 2.3) has been used in oil spill contingency planning and spill response
operations (Hayes et al., 1995). Major factors considered in ranking habitat sensitivity include
shoreline type (substrate, grain size, tidal elevation), exposure to wave and tidal energy,
biological productivity and sensitivity, and ease of cleanup. Bioremediation may be effective
and cause the least damage on both the moderately and the most sensitive shoreline types.
The Environmental Sensitivity Index for freshwater shorelines is shown in Table 2.4, based on
NOAA & API (1994) and Hayes et al. (1995 & 1997). Major factors considered in ranking ESI
for these habitats include degree of exposure to natural removal processes, biological
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productivity and sensitivity, human use of the habitat, and ease of oil removal. Bioremediation
may be feasible and cause the least damage on both the moderately sensitive and the most
sensitive shoreline types, although its effectiveness is still uncertain due to the lack of sufficient
research and field efficacy demonstrations.
Table 2.3 Shoreline ESI ranking for habitats in marine shorelines (where 1 is least sensitive
and 10 is most sensitive to oil and clean up actions)
Environmental Sensitivity Index (ESI) Shoreline Type
1 Exposed rocky shores
Sea walls and piers
2 Exposed wave-cut platforms
3 Fine-grained sand beaches
4 Coarse-grained sand beaches
5 Mixed sand and gravel beaches
6 Gravel beaches and riprap
7 Exposed tidal flats
8 Sheltered rocky shores
9 Sheltered tidal flats
1 0 Salt marshes and Mangroves
2.3 Biodegradation of Oil
Biodegradation of oil is one of the most important processes involved in weathering and the
eventual removal of petroleum from the environment, particularly for the nonvolatile
components of petroleum. Numerous scientific review articles have covered various aspects of
this process and the environmental factors that influence the rate of biodegradation (Zobell, 1946
& 1973; Atlas, 1981 & 1984; NAS, 1985; Focht and Westlake, 1987; Leahy and Colwell, 1990).
2.3.1 Mechanism of oil biodegradation: a microbiological perspective
Distribution of hydrocarbon-degrading microorganisms
Microorganisms capable of degrading petroleum hydrocarbons and related compounds are
ubiquitous in marine, freshwater, and soil habitats. Over 200 species of bacteria, yeasts and
fungi have been shown to degrade hydrocarbons ranging from methane to compounds of over 40
carbon atoms (Zobell, 1973). In the marine environment, bacteria are considered to be the
predominant hydrocarbon-degraders with a distribution range that even covers extreme cold
Antarctic and Arctic environments (Floodgate, 1984; Jordan and Payne, 1980). In the freshwater
environment, yeast and fungi may also play a significant role in degrading petroleum
hydrocarbons (Cooney, 1984). Some of the most important hydrocarbon-degrading
microorganisms in both marine and freshwater environments are listed in Table 2.5.
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Table 2.4 Shoreline ESI ranking for habitats in freshwater shorelines (where 1 is least
sensitive and 10 is most sensitive to oil and clean up actions)
ESI
Lacustrine a
Large Rivers a
Small Rivers b
1
Exposed rocky cliffs
Exposed man-made structures
Exposed rocky banks
Vertical, solid revetments
Quiet pools with low-sensitive
banks
2
Shelving bedrock shores
Rocky shoals, bedrock
Ledges
Small, nonnavigable channel
with moderate currents and
low-sensitive banks
3
Eroding scarps in
unconsolidated sediments
Exposed, eroding banks in
Unconsolidated sediments
Navigable channel with
moderate currents and low-
sensitive banks
4
Sand beaches
Sandy bars and gently
Sloping banks
Small, nonnavigable channel
with rapids over bedrock
5
Mixed sand and gravel
beaches
Mixed sand and gravel bars
Navigable channel with rapids
over bedrock
6
Gravel beaches and riprap
Gravel bars and riprap
Channel with associated low-
vulnerable upper bottomland
hardwoods
7
Exposed flats
Not present
Navigable streams with
associated wide swamps on
one side
8
Sheltered rocky shores
Sheltered man-made
structures
Vegetated, steeply sloping
Bluffs
Sheltered man-made
Structures
Navigable streams with
associated wide swamps on
both sides
9
Sheltered vegetated low
banks
Vegetated low banks
Muddy substrates
(unvegetated)
Meandering channel with
abundant leakage points into
associated swamps and ox-
bows
10
Sheltered sand flats
Freshwater marshes and
swamps
Freshwater marshes and
Swamps
Navigable anastomosing
channel with abundant leakage
points into associated swamps
a ESI adapted from Hayes et al. (1995)
b RSI (Reach Sensitivity Index) adapted from Hayes et al. (1997)
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Table 2.5 Representative microorganisms capable of degrading petroleum hydrocarbons (Based
on Atlas, 1984; Focht and Westlake, 1987; Jordan and Payne, 1980; Leahy and Colwell, 1990)
Bacteria
Yeast and Fungi
Achromobacter
Aspergillus
Acinetobacter
Candida
Alcaligenes
Cladosporium
Arthrobacter
Penicillium
Bacillus
Rhodotorula
Brevibacterium
Sporobolomyces
Cornybacterium
Trichoderma
Flavobacterium
Nocardia
Pseudimonas
Vibrio
The distribution of hydrocarbon-utilizing microorganisms is also related to the historical
exposure of the environment to hydrocarbons. Those environments with a recent or chronic oil
contamination will have a higher percentage of hydrocarbon degraders than unpolluted areas. In
"pristine" ecosystems, hydrocarbon utilizers may make up less than 0.1% of the microbial
community; and in oil-polluted environments, they can constitute up to 100% of the viable
microorganisms (Atlas, 1981).
It should be noted that there is no single strain of bacteria with the metabolic capacity to degrade
all the components found within crude oil. In nature, biodegradation of a crude oil typically
involves a succession of species within the consortia of microbes present. Microorganisms
classified as non-hydrocarbon utilizers may also play an important role in the eventual removal
of petroleum from the environment. Degradation of petroleum involves progressive or
sequential reactions, in which certain organisms may carry out the initial attack on the petroleum
constituent; this produces intermediate compounds that are subsequently utilized by a different
group of organisms, in the process that results in further degradation (Karrick, 1977).
Biodegradation of oil components
As described earlier, petroleum components can be classified into four major groups: saturates,
aromatics, resins, and asphaltenes. Major metabolic pathways for many of these compounds
have been well studied and documented (Atlas, 1981 & 1984; Cerniglia, 1992; Watkinson and
Morgon, 1990) to explain their differences in susceptibility to biodegradation.
Saturates In general, the //-alkanes are the most readily degraded components in a
petroleum mixture (Zobell, 1946; Atlas, 1981). Biodegradation of //-alkanes with molecular
weights up to C44 has been demonstrated (Haines and Alexander, 1974). Alkanes in the C10 to
C26 range are considered the most readily and frequently utilized hydrocarbons (Atlas 1995b;
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NAS, 1985). The predominant mechanism of n-alkane degradation involves terminal oxidation
to the corresponding alcohol, aldehydes, or fatty acid functional group. Branched alkanes are less
readily degraded in comparison to //-alkanes. Methyl branching increases the resistance to
microbial attack because fewer alkane degraders can overcome the blockage of beta-oxidation
(NAS, 1985). Highly branched isoprenoid alkanes, such as pristane and phytane, which were
earlier thought to be resistant to biodegradation, have also been shown to be readily
biodegradable. Cycloalkanes, however, are particularly resistant to biodegradation. Complex
alicyclic compounds such as hopanes and steranes are among the most persistent compounds of
petroleum spills in the environment (Atlas, 1981).
Aromatics Although the aromatics are generally more resistant to biodegradation, some low
molecular weight aromatics such as naphthalene may actually be oxidized before many saturates
(Focht and Westlake, 1987). Monoaromatic hydrocarbons are toxic to some microorganisms due
to their solvent action on cell membranes, but in low concentrations they are easily
biodegradable under aerobic conditions. PAHs with 2-4 rings are less toxic and biodegradable at
rates that decrease with the level of complexity. PAHs with five or more rings can only be
degraded through co-metabolism, in which microorganisms fortuitously transform non-growth
substrates while metabolizing simpler hydrocarbons or other primary substrates in the oil.
Alkylated aromatics are degraded less rapidly than their parent compounds; the more highly
alkylated groups are degraded less rapidly than less alkylated ones. The metabolic pathways for
the biodegradation of aromatic compounds have been the subject of extensive study (Atlas, 1981;
Prince, 1993; Cerniglia, 1992). The bacterial degradation of aromatics normally involves the
formation of a diol, followed by ring cleavage and formation of a di-carboxylic acid. Fungi and
other eukaryotes normally oxidize aromatics using mono-oxygenases, forming a trans-diol.
Resins and asphaltenes Compared to saturates and aromatics, very little is known about
biodegradation of resins and asphaltenes; this is due to their complex structures, which are
difficult to analyze. Resins and asphaltenes have previously been considered to be refractory to
degradation. However, there is recent evidence of asphaltene degradation through cometabolism
(Leahy and Colwell, 1990). Some resins, particularly low-molecular-weight resin fractions, can
also be biodegraded at low concentrations (NAS, 1985). Further research is still needed to
understand the biodegradation of these compounds.
In summary, the susceptibility of petroleum hydrocarbons to microbial degradation is generally
in the following order: n-alkanes > branched alkanes > low-molecular-weight aromatics > cyclic
alkanes. However, this pattern is not universal (Perry, 1984). The compositional heterogeneity
among different oils greatly affects the biodegradation rate of their constituents. The degradation
rate for the same oil constituents may vary significantly for different oils. Cometabolism also
plays an important role in oil biodegradation. Many complex branched, cyclic, and aromatic
hydrocarbons, which otherwise would not be biodegraded individually, can be oxidized through
cometabolism in an oil mixture due to the abundance of other substrates that can be metabolized
easily within the oil (Atlas, 1981). The biological fate of oil components in an oil mixture still
requires further research. Particularly, effort should be made to establish a database regarding
the biodegradability of different types of oils and petroleum products.
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2.3.2 Environmental factors affecting oil biodegradation
When oil spills occur in the environment, the rate of oil biodegradation is also greatly influenced
by the characteristics of the contaminated environment. Major environmental factors affecting
oil biodegradation include weathering processes, temperature, availability and concentration of
nutrients, availability and concentration of oxygen, and pH.
Weathering processes
The weathering processes described in section 2.2.1 have profound effects on oil biodegradation.
Evaporation of volatile oil components may benefit microorganisms by removing more toxic
low-molecular-weight components such as benzene and smaller //-alkanes. However, this
process also leads to a lower biodegradable percentage of oil, since these components in general
are readily biodegraded (Atlas, 1981; NAS, 1985). The oil surface area is important because
growth of oil degraders occurs almost exclusively at the oil-water interface (Atlas and Bartha,
1992). Formation of water-in-oil emulsions or mousses reduces the surface area, therefore
decreasing biodegradation. Tarballs, which are large aggregates of weathered and undegraded
oil, also restrict access to microorganisms because of their limited surface area (Leahy and
Colwell, 1990). Dispersion of hydrocarbons in the water column in the form of oil-in-water
emulsions increases the surface area of the oil and thus its availability for microbial attack. The
formation of oil-in-water emulsions through the microbial production and release of
biosurfactants has also been found to be an important process in the uptake of hydrocarbons by
bacteria and fungi (Singer and Finnerty, 1984). In contrast, the application of chemical
dispersants has produced mixed results and has not been shown to be an effective way to
enhance oil biodegradation. Photooxidation leads to the formation of more soluble compounds,
which are often more biodegradable. However, the effect of photooxidation processes on
biodegradation is still not well understood (Nicodem etal., 1997).
Biodegradation rates are also influenced by concentrations of individual oil constituents, which
may be affected by various weathering processes. For example, microbes may attack very low
concentrations of pollutants in the environment inefficiently (Focht and Westlake, 1987).
However, high concentrations of hydrocarbons may cause inhibition of biodegradation by
nutrient or oxygen limitations or toxic effects. There would seem to be, for many hydrocarbons,
an optimum concentration range for metabolism below which degradation is not stimulated and
above which inhibition occurs. Weathering processes will affect the ultimate concentrations of
petroleum hydrocarbons in the environment in different ways. Evaporation may reduce the
concentrations of volatile compounds but concentrate some other constituents. Sorption and
emulsification may concentrate the pollutants, while dispersion and dissolution tend to dilute
them.
Temperature
The ambient temperature of an environment affects both the properties of spilled oil and the
activity or population of microorganisms. At low temperatures, the viscosity of the oil increases,
while the volatility of toxic low-molecular-weight hydrocarbons is reduced, delaying the onset of
21
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biodegradation (Atlas, 1981). Some hydrocarbons are more soluble at lower temperatures (e.g.,
short-chain alkanes), and some low-molecular-weight aromatics are more soluble at the higher
temperature (Focht and Westlake, 1987). Although hydrocarbon biodegradation can occur over
a wide range of temperatures, the rate of biodegradation generally decreases with decreasing
temperature. Highest degradation rates generally occur in the range of 30 to 40°C in soil
environments, 20 to 30°C in some freshwater environments, and 15 to 20°C in marine
environments (Bossert and Bartha, 1984; Cooney, 1984; Jordan and Payne, 1980).
The effect of temperature is also complicated by other factors such as the composition of the
microbial population. In environments where a psychrophilic population has been established,
degradation can occur at significant rates under cold conditions. Hydrocarbon biodegradation
has been observed at temperature as low as 0-2°C in seawater and -1.1°C in a soil. Colwell et
al. (1978) reported greater degradation of Metula crude oil at 3°C than at 22°C with a mixed
culture in beach sand samples. Westlake et al. (1974) also found that bacteria capable of
degradation at 4°C would metabolize oil at 30°C, but those populations that developed at 30°C
had a limited activity at 4°C.
Oxygen
Aerobic conditions are generally considered necessary for extensive degradation of oil
hydrocarbons in the environment since major degradative pathways for both saturates and
aromatics involve oxygenases (Atlas, 1981; NAS, 1984; Cerniglia, 1992). Many studies have
shown that oxygen depletion leads to sharply reduced biodegradation activities in marine
sediments and in soils (Atlas, 1981; Bossert and Bartha, 1984; Hambrick III et al., 1980).
Conditions of oxygen limitation normally do not exist in the upper levels of the water column in
marine and freshwater environments and in the surface layer of most beach environments. It
may become limiting in subsurface sediments, anoxic zones of water columns, and most fine-
grained marine shorelines, freshwater wetlands, mudflats, and salt marshes. Factors affecting the
availability of oxygen also include the action of wave and water flow, the physical state of the
oil, and the amount of available substrates.
Anaerobic oil degradation has been shown in some studies to occur only at negligible rates, as
reviewed by Atlas (1981), leading to the conclusion that the environmental importance of
anaerobic hydrocarbon degradation can be discounted. However, recent studies have shown that
anaerobic hydrocarbon metabolism may be an important process in certain conditions (Head and
Swannell, 1999). The biodegradation of some aromatic hydrocarbons, such as BTEX
compounds, has been clearly demonstrated to occur under a variety of anaerobic conditions
(Krumholz et al., 1996; Leathy and Colwell, 1990). Studies have also demonstrated that in some
marine sediments, PAHs and alkanes can be degraded under sulfate-reducing conditions at
similar rates to those under aerobic conditions (Caldwell et al., 1998; Coates et al., 1997). The
importance of anaerobic biodegradation of oil in the environment still requires further studies.
Nutrients
In theory, approximately 150 mg of nitrogen and 30 mg of phosphorus are utilized in the
conversion of 1 g of hydrocarbon to cell materials (Rosenberg and Ron, 1996). When a major
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oil spill occurs in marine and freshwater environments the supply of carbon is dramatically
increased and the availability of nitrogen and phosphorus generally becomes the limiting factor
for oil degradation (Atlas, 1984; Leahy and Colwell, 1990). In marine environments, nutrient
limitation is generally correlated to the low background levels of nitrogen and phosphorus in
seawater (Floodgate, 1984). Nutrient concentrations are more variable in freshwater systems
where lakes and wetlands range from oligotrophic to eutrophic; rivers can be nutrient-poor at the
source, but generally become nutrient-rich downstream after receiving industrial and domestic
effluents and agricultural runoff (Cooney, 1984). Freshwater wetlands are typically considered to
be nutrient limited, due to heavy demand for nutrients by the plants. They are also viewed as
being nutrient traps, as a substantial amount of nutrients may be bound in biomass (Mitsch and
Gosselink, 1993). Both freshwater lakes and wetlands may also exhibit seasonal variations in
nutrient levels, which will affect the performance of oil biodegradation. Ward and Brock (1976)
found that in an oil-contaminated lake, oil biodegradation was at the highest rate during early
spring when the nutrient content (i.e., N and P) was also high. As N and P levels decreased in the
summer (probably due to algal productivity) oil biodegradation also decreased. Another potential
limiting nutrient is iron, which was found to limit oil degradation in clean offshore seawater, but
is not likely to be limiting in freshwater (Focht and Westlake, 1987).
Other factors
Other important factors affecting biodegradation of petroleum hydrocarbons include pH and
salinity. The pH of seawater is generally stable and slightly alkaline (Bossert and Bartha, 1984).
In contrast, the pH of freshwater and soil environments can vary widely. Organic soils in
wetlands are often acidic, while mineral soils have more neutral and alkaline conditions. Most
heterotrophic bacteria and fungi favor a neutral pH, with fungi being more tolerant of acidic
conditions. Studies have shown that degradation of oil increases with increasing pH, and that
optimum degradation occurs under slightly alkaline conditions (Dibble and Bartha, 1979; Focht
and Westlake, 1987).
Changes in salinity may affect oil biodegradation through alteration of the microbial population.
Dramatic variation in salinity may occur in estuarine environments where marine organisms
mingle with freshwater forms. Many freshwater organisms can survive for long periods in
seawater although few can reproduce. In contrast, most marine species have an optimum salinity
range of 2.5 to 3.5% and grow poorly or not at all at salinity lower than 1.5 to 2% (Zobell, 1973).
In a study of hypersaline salt evaporation ponds, Ward and Brock (1978) showed that rates of
hydrocarbon metabolism decreased with increasing salinity in the range of 3.3 to 28.4%. More
studies are required to understand the effect of salinity on oil biodegradation.
2.3.3 Evaluation of oil biodegradation: application of biomarkers
The evaluation of oil biodegradation is a difficult task, especially in the field, due to the
complication of weathering processes and the heterogeneity of contaminated sites. As described
earlier, physical and chemical weathering can significantly affect the composition and
concentrations of oils. Oil contaminated sites are often highly heterogeneous, where oil
concentrations can vary greatly within a small area. Consequently, variability associated with
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field studies can be so high as to preclude or interfere with one's ability to discern significant
treatment differences. Non-biodegradable or slowly biodegradable components in oil - often
called biomarkers - have been used successfully to mitigate the high variability associated with
field studies (Bragg etal., 1994; Venosa et al., 1996; Lee et al., 1997a). This approach estimates
the extent of biodegradation by evaluating the ratios of target hydrocarbon concentrations
relative to the concentration of these recalcitrant biomarkers. Studies have shown that use of
biomarkers reduces spatial variability of oil data when compared to other mass balance
approaches and allows biodegradation to be monitored effectively by reducing the number of
samples required (Douglas et al., 1994).
Commonly used biomarkers for evaluation of biodegradation of crude oils include the
isoprenoids pristane and phytane, steranes, and the pentacyclic triterpanes such as the hopanoids
(Peters and Moldowan, 1993). While the isoprenoids pristane and phytane are somewhat more
resistant to biodegradation than //-alkanes with similar boiling points (Ci7, Ci8), they should only
be used to monitor the earliest stages of a biodegradation treatment program, as they are known
to be biodegradable under natural conditions (Prince et al., 1994b). Hopanes have become the
biomarker of choice as they are much more resistant to microbial biodegradation (Atlas, 1981;
Peters and Moldowan, 1993). The compound 17a(H), 21(3(H)-hopane was successfully used to
determine the efficacy of bioremediation field trials coordinated with the Exxon Valdez spill
clean up operations (Douglas et al., 1994; Mearns, 1997; Prince, et al., 1994 a&b). However,
caution must be taken with their use as biomarkers since they are also very resistant to physical
and chemical weathering processes that affect many alkanes and aromatics. Therefore, hopane
normalization is more useful in reducing the variability associated with heterogeneous oil
distribution or in cases where the effects of physical and chemical weathering are negligible.
Biodegradation can also be verified as the main removal mechanism by determining the relative
degradation rates for homologous series of alkylated PAHs (Elmendorf et al., 1994; Venosa et
al., 1997a).
2.4 Laboratory Studies on Bioremediation of Oil
Biodegradation as a natural process may proceed slowly, depending on the type of oil (i.e., light
crude oils degrade faster than heavier oils). Bioremediation strategies are based on the
application of various methodologies to increase the rate or extent of the biodegradation process.
The success of oil spill bioremediation depends on our ability to optimize various physical,
chemical, and biological conditions in the contaminated environment. As described in previous
sections, the most important requirement is the presence of microorganisms with the appropriate
metabolic capabilities. If these microorganisms are present, then optimal rates of growth and
hydrocarbon biodegradation can be sustained by ensuring that adequate concentrations of
nutrients and oxygen are present and that the pH is between 6 and 9 (Atlas and Bartha, 1992).
The physical and chemical characteristics of the oil and oil surface area are also important
determinants of bioremediation success. Obviously, some of these factors can be manipulated
more easily than others. For example, on an operational scale, there is nothing that can be done
to alter the chemical composition of the oil.
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There are two main approaches to oil spill bioremediation. 1) bioaugmentation, in which oil-
degrading microorganisms are added to supplement the existing microbial population, and 2)
biostimulation, in which the growth of indigenous oil degraders is stimulated by the addition of
nutrients or other growth-limiting cosubstrates and/or habitat alteration. Both these approaches
have been extensively studied, on a laboratory scale as well as in the bioremediation of oil
contaminated shorelines.
2.4.1 Bioaugmentation
Although hydrocarbon-degrading microorganisms are widespread in nature, bioaugmentation has
been considered as a potential strategy for oil-bioremediation since the 1970s. The rationale for
adding oil-degrading microorganisms is that indigenous microbial populations may not be
capable of degrading the wide range of potential substrates present in complex mixtures such as
petroleum (Leahy and Colwell, 1990). Other conditions under which bioaugmentation may be
considered are when the indigenous hydrocarbon-degrading population is low, the speed of
decontamination is the primary factor, and when seeding may reduce the lag period to start the
bioremediation process (Forsyth et al., 1995).
Many vendors offer microbial agents claiming to enhance oil biodegradation (Prince, (1993).
However, laboratory studies on bioaugmentation have produced mixed results. Aldrett et al.
(1997) tested 12 commercial microbial cultures for bioremediation of Alaska North Slope crude
oil in the lab. After 28 days, four products showed an enhancement of oil biodegradation with
significantly higher degradation rates of alkanes and aromatics when compared to a nutrient
control. In another shaker-flask experiment, Hozumi et al. (2000) investigated the effectiveness
of a microbial product in treating a heavy oil spilled from Nakhodka using the thin layer
chromatography-flame ionization detection (TLC-FID) analysis. They found that approximately
35% of the oil was degraded with addition of the microbial product compared to no oil loss for a
control during a three-week test period. Surprisingly, the asphaltene fraction showed the highest
loss among the four major oil components, which raises the question whether this oil loss was
actually due to biodegradation rather than some quality control problem with the chemical
analysis. Some laboratory studies found that microbial seeding may enhance oil degradation in
seawater but not in freshwater environments (Leahy and Colwell, 1990). To examine whether
microbial products can compete with the indigenous populations, Venosa et al. (1991) tested 10
different commercial microbial products using weathered Alaskan crude oil in shaker flask
microcosms. Although two products showed enhancement compared to a nutrient control, better
degradation was observed in every case when the commercial products were first sterilized,
suggesting that indigenous Alaskan microorganisms were primarily responsible for the oil
biodegradation and seeded microorganisms seemed to compete poorly with the indigenous
population in the closed flask environment. Thus, bioaugmentation may be effective in bench-
scale studies where environmental conditions are well controlled, but this will not guarantee its
effectiveness in the field.
Creation of a "superbug" that combines the genetic information from many organisms and the
ability to degrade a variety of different types of hydrocarbons has also been considered. Friello
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et al. (1976) successfully produced a multiplasmid-containing Pseudomonas strain capable of
oxidizing aliphatic, aromatic, terpenic, and polyaromatic hydrocarbons. Thibault and Elliot
(1980) also developed a multiplasmid P. putida strain that can simultaneously degrade some
lighter alkanes and aromatics. However, the survival of such a strain in the environment could is
questionable. More importantly, the issues of safety, containment, and the potential for
ecological damage must be fully resolved before field testing of these organisms can be
conducted (Leahy and Colwell, 1990). There is also the problem of public perception over the
release of "foreign" and especially "genetically engineered" microorganisms into the
environment.
2.4.2 Biostimulation
Biostimulation involves the addition of rate-limiting nutrients to accelerate the biodegradation
process. In most shoreline ecosystems that have been heavily contaminated with hydrocarbons,
nutrients are likely the limiting factors in oil biodegradation. The main purpose of bench-scale
treatability studies is to determine the type, concentration, and frequency of addition of
amendments needed for maximum stimulation in the field (Venosa, 1998).
Most laboratory experiments have shown that addition of growth limiting nutrients, namely
nitrogen and phosphorus, has enhanced the rate of oil biodegradation. However, the optimal
nutrient types and concentrations vary widely depending on the oil properties and the
environmental conditions. Wrenn et al. (1994) studied the effects of different forms of nitrogen
on biodegradation of light Arabian crude oil in respirometers. They found that in poorly
buffered seawater, nitrate is a better nitrogen source than ammonia because acid production
associated with ammonia metabolism may inhibit oil biodegradation. When the culture pH was
controlled, the performance of oil biodegradation was similar for both amendments with a
shorter lag time for ammonia addition. Ramstad and Sveum (1995) also compared the effect of
nitrate, ammonia, and an organic nitrogen-containing nutrient on biodegradation of topped
Statfjord crude oil in a continuous-flow seawater column system. With no control of pH in this
study, nitrate was found to have the most pronounced effect in stimulating oil degradation when
using pristane as a biomarker. However, in a microcosm study, Jackson and Pardue (1999) found
that addition of ammonia appeared to be more effective than nitrate in stimulating degradation of
crude oil in salt marsh soils. On a weight basis, the amount of ammonia required to achieve the
same increase in biodegradation as nitrate was only about 20%. This was attributed to the fact
that ammonia is less likely to be lost from the system by washout due to its higher adsorptive
capacity to organic matter.
Oil biodegradation largely takes place at the interface between oil and water. Therefore, the
effectiveness of biostimulation depends on the nutrient concentration in the interstitial pore water
of oily sediments (Atlas and Bartha, 1992; Bragg el al., 1994). The nutrient concentration
should be maintained at a level high enough to facilitate bacterial growth. However, caution
must be exercised as excessive concentrations of nutrients, such as ammonia, induce toxic
responses in many marine species (Pritchard et al., 1991). Using nitrate as a biostimulation
agent, Venosa et al. (1994) determined that approximately 1.5 to 2.0 mg N/L supported near
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maximal biodegradation of heptadecane immobilized onto sand particles in a microcosm study.
Du et al. (1999) investigated the optimal nitrogen concentration for the biodegradation of Alaska
North Slope crude oil in continuous flow beach microcosms at a loading of 5g-oil/kg sand. The
results showed that nitrate concentrations below approximately 10 mgN/L limited the rate of oil
biodegradation. The higher nutrient requirement was attributed to the more complex substrate
(crude oil). Ahn (1999) further studied the effect of nitrate concentrations under tidal flow
conditions using the same microcosms, oil, and oil loading as Du et al. (1999). Nitrate
concentrations ranging from 6.25 to 400 mg N/L were supplied semi-diurnally to simulate tide
flow. The results from both oil analysis (hopane as a biomarker) and microbial growth
(phospholipid analysis) showed that the optimal nitrate concentration fed under these conditions
was approximately 25 mgN/L. However, the data showed that at the end of the test (after one
month), an approximately 80% degradation of all target alkanes and PAHs was achieved in all
test cases that covered a range of nitrate concentrations in the test solutions added. This result
suggested that nitrate concentration might only affect the rates and not the extent of oil
degradation. Further research in this regard is required to optimize bioremediation strategies.
One of the main challenges associated with biostimulation in oil-contaminated coastal areas is
maintaining optimal nutrient concentrations in contact with the oil. Oil from offshore spills
usually contaminates the intertidal zone, where the washout rate for water-soluble nutrients can
be very high, and this can adversely impact the effectiveness of biostimulation. Many attempts
have been made in the design of nutrient delivery systems that overcome the washout problems
characteristic of intertidal environments (Prince, 1993). These include oleophilic and slow-
release fertilizer formulations, as well as systems that rely on the subsurface flow of water
through the beach (Wise et al., 1994). Several papers have compared the effectiveness of these
nutrient products to stimulate hydrocarbon biodegradation rates. Croft et al. (1995) tested the
efficiency of an oleophilic organic (Inipol EAP22) and a slow-release inorganic fertilizer (Max
Bac) and found that the oleophilic fertilizer was much more effective at stimulating oil
degradation on sand than the slow-release product. Sveum and Ramstad (1995) also found that
organic products such as fish meal and stick water (a fish meal by-product) were more effective
than a slow-release fertilizer (Mac Bac). The failure for this slow-release fertilizer was attributed
to the nutrient release rate being too slow to affect oil biodegradation. However, in some other
studies, application of organic fertilizers such as Inipol EAP22 also failed to stimulate oil
biodegradation (Sveum and Ladousse, 1989; Safferman, 1991). Safferman (1991) investigated
the rates of nutrient release from several slow-release products and found that Inipol EAP22 was
rapidly washed out from oiled cobble before becoming available to hydrocarbon-degrading
bacteria. However, no attempt was made in most of these studies to estimate the steady-state
nutrient concentrations that would result in an intertidal environment. The variable results from
the laboratory studies indicate that the performance of these products greatly depends on the
nutrient release rates and the prevailing environmental conditions.
Although much research has been carried out on the bioremediation of oil-contaminated marine
shorelines, few studies have been conducted on oil bioremediation in wetland environments.
Purandare et al. (1999) conducted the only reported microcosm study on biostimulation in
freshwater wetland. They investigated different inorganic mineral nutrients for their ability to
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enhance biodegradation of a crude oil. Aquaria of 10-gallon capacity, filled with wetland soil
and planted with species of emergent wetland plants, were used to simulate natural wetlands.
Two levels of water coverage were studied: (1) water level even with soil surface, and (2) water
level 10 cm above the soil surface. Six treatments were evaluated for each water level: unoiled,
no-nutrient control; oiled + no nutrient control; oiled + nitrate addition; oiled + nitrate +
phosphate addition; oiled + ammonia addition; and oiled + ammonia + phosphate addition.
Approximately 14g of weathered light Arabian crude oil was added to each column to form
about a 2mm oil layer. The results showed that for both flooded and unflooded wetland
conditions, the addition of nitrate and phosphate seemed to enhance the degradation of oil above
the natural attenuation rate. The highest biodegradation rates of alkanes and PAHs occurred in
the high water level microcosms receiving nitrate and phosphorus, in which a 90% alkane
reduction and 50% PAH removal was observed, compared to only 50% alkane reduction and
40% PAH removal in the control microcosms. Higher degradation of alkanes and PAHs in the
high water level relative to that in the low water level suggested an increased hydrocarbon
bioavailability. However, a microcosm study conducted by Garcia-Blanco et al. (2001a) in a
simulated tidal salt marsh found that addition of nutrients did not stimulate the biodegradation of
a No. 2 fuel oil. Low oxygen availability was suggested to be the limiting factor for oil
degradation in salt marshes.
In summary, laboratory studies have shown that biostimulation and, in some cases
bioaugmentation, can enhance the rates of oil biodegradation, particularly in marine
environments. Oxygen may become a limiting factor in oil biodegradation under certain
circumstances, such as salt mashes and freshwater wetlands. However, these conclusions still
need to be verified through field evaluations.
2.5 Demonstrations of Oil Bioremediation Under Field Conditions
Field studies can provide the most convincing demonstration of the effectiveness of oil
bioremediation since laboratory studies are not always able to account for numerous real world
conditions such as spatial heterogeneity, biological interactions, and mass transfer limitations.
Compared to laboratory investigations, relatively few tests have been carried out to evaluate the
effectiveness of oil bioremediation in the field because such trials are both difficult and
expensive to conduct. Swannell et al. (1996) conducted the most extensive review available on
field evaluations of oil bioremediation in marine environments. Venosa (1998) presented an in-
depth critical review emphasizing problems in the design and control of the existing field tests. A
review by Lee (2000) addressed the potential significance of enhancing plant growth (i.e.
phytoremediation) and oil mineral aggregate formation as biostimulation treatments. Other
reviews are also available (Prince, 1993; Leahy and Colwell, 1990). This section will summarize
the latest findings from recent field studies on marine shorelines and freshwater wetlands, as well
as major points identified in the previous reviews.
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2.5.1 Mesocosm studies
Mesocosms or pilot-scale systems can help to simulate actual conditions at relatively low cost,
and are frequently used as bridges between microcosms and field systems. Mesocosms have
been used to evaluate the effectiveness of numerous bioremediation strategies.
Bioaugmentation
Unlike results from bench-scale tests, numerous mesocosm studies have demonstrated the
ineffectiveness of bioaugmentation treatments. For example, Tagger et al. (1983) overlaid two
mesocosms with crude oil. One was inoculated with an acclimated culture, while only
indigenous populations were used in the other. Five months after inoculation, no statistically
significant change in oil composition occurred between the two treatments. Neralla el al. (1995)
investigated the effect of seeding in salt marsh conditions. The greenhouse experiment was
conducted with 19-L buckets filled with marsh sediments and actively growing Spartina
alterniflora. There were 10 duplicated treatments in a total of 20 mesocosms. Results showed
that the addition of bioaugmentation products did not enhance the degradation of weathered
Arabian Lube crude oil. The conclusion was not surprising since the soil used in the experiment
came from microcosms used in a similar study, and a large population of hydrocarbon-degrading
microorganisms was already present in the sediment. These results again indicate that oil
biodegradation is unlikely limited by availability of hydrocarbon degraders and that seeded
microorganisms could not compete with indigent populations.
Biostimulation
Contrary to the mesocosm bioaugmentation studies, the addition of nutrients has proven to be an
effective strategy for oil bioremediation. Basseres et al. (1993) conducted a mesocosm trial on
bioremediation of light Arabian crude oil using two 600-liter tanks filled with sandy beach
materials at a site near the Mediterranean. Animal meal containing 60% protein was added at 10
% w/w to one tank. A second was left untreated as a control. Over the 60-day period of the test,
40% of the aliphatic fraction in the treated oil was degraded whereas only 25% was degraded in
the control. The number of hydrocarbon degraders was found to be higher in the meal-treated
mesocosm than in the control. However, no replicate treatments and nutrient measurements
were performed during the study, so it was not possible to determine if the observed differences
were statistically significant.
Mendelssohn et al. (1995) conducted a greenhouse mesocosm study to determine the effect of oil
bioremediation products on salt marsh ecosystems. A randomized block design was used with
three treatments (fertilizer, microbial products and control) at two levels of oil dosage. Each
treatment combination was replicated five times for a total of 30 sods of marsh. The results
demonstrated that the addition of a fertilizer product significantly increased the growth response
of a salt marsh grass {Spartina alterniflora) and the rate of soil respiration, while the microbial
products did not significantly affect either of these processes, suggesting the bioremediation
products had neither toxic nor stimulatory effects on the salt marsh environments.
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Purandare (1999) conducted a mesocosm study following the microcosm study described earlier
(Purandare et al., 1999) to further test the effectiveness of bioremediation of an oil-contaminated
freshwater wetland. Outdoor mesocosms each measuring 6 m x 4.5 m, filled with wetland
sediments, and planted with three species of emergent wetland plants were used in the study to
investigate the effect of different inorganic mineral nutrients on biodegradation of a Louisiana
crude oil. A total of 12 of these mesocosms were used (3 replicates of 4 treatments). The four
treatment strategies included a no-nutrient control, phosphate addition, nitrate + phosphate
addition, and ammonia + phosphate addition. Biodegradation rates were computed from hopane-
normalized analyte data. The results showed that addition of nutrients seemed to enhance oil
biodegradation initially. However, beyond 12 weeks, the untreated control achieved a
comparable degree of biodegradation. No conclusion as to which nutrients were actually limiting
the biodegradation process was reached due to the high variability in the data. The study also
found that addition of nutrients led to better plant and root growth, which suggested that,
although biostimulation may not significantly enhance oil degradation in freshwater wetlands, it
may encourage a faster recovery of the ecosystem. It also suggested that the wetland plants may
have out-competed the oil-degrading microorganisms for nutrients and may have used substrates
other than oil hydrocarbons for their growth (soil organic matter).
2.5.2 Field Demonstrations
Bioremediation field studies are reviewed here, first with respect to bioaugmentation and
biostimulation, then by a more detailed discussion of four case studies that cover a wide range of
shoreline types.
Bioaugmentation
The effectiveness of seeding has been studied in only a few field trials. Venosa et al. (1992)
conducted a field test in Prince William Sound following the Exxon Valdez spill to investigate
the effectiveness of two commercial microbial products vis-a-vis natural attenuation and nutrient
addition alone. These products were selected based on a previous laboratory study (Venosa et
al., 1991). This field trial failed to demonstrate enhanced oil biodegradation by these products.
No biostimulation occurred in the nutrient control plots either. There were no significant
differences between the treatment and control plots during the 27-day trial period. However, the
site where the project took place (Disk Island) was characterized as having highly weathered
(degraded) oil and very calm waters, so dissolved oxygen may have been limiting, thus
precluding effective biodegradation by any means.
Other studies (Lee and Levy, 1987; Tagger et al., 1983) suggested that exogenous microbial
inocula are not able to compete successfully with indigenous populations. One approach in
overcoming this competition has been proposed by Rosenberg et al. (1992). They developed a
product that combined a polymerized ureaformaldehyde (F-l) with a selected oil-degrading
culture capable of using this fertilizer as a nitrogen source. Thus, the bacteria had a selective
advantage over the indigenous population unable to utilize F-l as nutrient source. A field trial
conducted at an Israeli beach showed that this approach seemed to be successful in enhancing oil
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biodegradation. However, conclusions were confounded by the lack of adequate controls in the
study (Swannell etal., 1996; Venosa, 1998).
Studies comparing the performance of bioaugmentation and biostimulation have suggested that
nutrient addition alone had a greater effect on oil biodegradation than did the addition of
microbial products (Lee et al., 1997a; Venosa et al., 1996). ). This is probably because the
microbial population is rarely a limiting factor as compared to the nutrients since the size of the
hydrocarbon-degrading bacterial population usually increases rapidly in response to oil
contamination. Lee et al. (1997a) conducted a 129-day field trial to compare the effect of four
treatments on biodegradation of weathered Venture Condensate in a sandy beach. The four
treatments included: inorganic nutrient and oil, a microbial product and oil, inorganic nutrient, a
microbial product and oil, and oiled control. C2-chrysenes was used as a biomarker due to the
low concentration of hopane in the condensate. The results showed that periodic addition of
inorganic nutrients was the most effective strategy for enhancing oil degradation and reducing
sediment toxicity, and that the full potential of the microbial product was limited by nutrient
availability. A similar study conducted in a wetland by Simon et al. (1999) also show that
addition of bioaugmentation agents did not enhance biodegradation of an Arabian medium crude
oil. However, nutrient addition did not demonstrate any significant effect in their study either,
suggesting other factors, such as oxygen, were limiting oil degradation.
Several other possible reasons for the failure of inocula in degrading contaminants in nature were
summarized by Goldstein et al. (1985), which include: (1) the concentration of the contaminant
may be too low to support the growth of the inoculated species, (2) the natural environment may
contain substances inhibiting growth or activity of the inocula, (3) the growth rate of the
inoculated species may be limited by predation such as protozoa, (4) the added species may use
other substrates in nature rather than the targeted contaminants, and (5) the seeded
microorganisms may be unable to move through the pores of the sediment to the contaminants.
A few field trials did claim success in demonstrating the effectiveness of oil bioaugmentation,
such as using Alpha BioSea™ to treat the Angolan Palanca crude oil spilled from Mega Borg off
Texas coast (Mauro and Wynne, 1990) and using Terra-Zyme™ in enhancing biodegradation of
a heavy oil spilled from Nakhodka in Japan (Tsutsumi et al., 2000). However, the success of
these studies was based on either visual observation (i.e. the Mega Borg study) or digital
photographic image analysis (i.e., the Nakhodka study). No comprehensive monitoring program
was used to verify the oil was indeed removed through enhanced biodegradation. The two
products basically contains the same formula of bacteria cultures and nutrients (Hozumi et al.,
2000). The observed visual effects may have been due to physical or chemical processes such as
surfactant action associated with the products (Swannell et al., 1996).
It seems that in most environments, indigenous oil-degrading microorganisms are more than
sufficient to carry out oil biodegradation if nutrient levels and other adverse environmental
conditions do not limit them. Future research on oil bioaugmentation should focus on
investigating which ecosystems may be deficient in oil degrading microorganisms and what
types of oils or important oil components indigenous bacteria may be incapable of degrading.
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Biostimulation
Although laboratory and pilot-scale studies have shown that biostimulation is a promising
approach in enhancing oil biodegradation, the effectiveness of various types of nutrients and
delivery strategies still require field demonstration.
Sveum and Ladousse (1989) investigated the performance of Inipol EAP 22 in different types of
sediments. The results showed that the oleophilic fertilizer enhanced oil biodegradation in
coarse-grained sediments but not in fine-grained sediments.
Researchers from Fisheries and Oceans-Canada (Lee and Levy, 1987; Lee and Levy, 1989; Lee
and Levy, 1991; Lee and Trembley, 1993; Lee et al., 1995a; and Lee et al., 1997a) conducted a
series of field tests to investigate the effect of different types of fertilizer and different deliver
strategies in a low energy, sandy beach or in a salt marsh. Their studies demonstrated that
biostimulation using periodic addition of inorganic fertilizers (ammonium nitrate and triple super
phosphate) increased the rate of oil removal from beaches as measured by changes in oil
composition relative to conserved biomarkers such as C2-chrysenes and/or the decline in the n-
Ci7/pristane and w-CWphytane ratios. In contrast, the addition of the oleophilic fertilizer Inipol
EAP 22 did not enhance oil degradation in a sandy beach (Lee and Levy, 1987 and 1989).
Another study involved periodic addition of water-soluble fertilizer granules (ammonium nitrate
and triple super phosphate) in an attempt to enhance biodegradation of waxy crude oil in a low-
energy, sandy beach and in a salt marsh (Lee and Levy, 1991). Two concentrations of the
NH4NO3 were tested (0.34 and 1.36 g/L sediment). The oil used was Terra Nova crude at two
different levels (0.3 and 3.0%). Results from the sandy beach showed that at the lower level of
oil contamination, no enhancement by fertilizer was achieved. However, at the higher oil
contamination level, substantial oil degradation occurred in the fertilized plots compared to the
unfertilized ones. Results in the salt marsh were the exact opposite. Enhancement by fertilizer
was significant at the 0.3% contamination level, but no enhancement occurred at the 3% oil
contamination, which was attributed to the penetration of oil into the anaerobic zone where little
degradation is expected. Studies on the utility and efficacy of various slow-release fertilizer
formulations also were evaluated (Lee et al., 1993). They demonstrated that the effects of
environmental factors controlling nutrient delivery from the various formulations under review
(e.g., sulfur-coated urea) were the key to bioremediation success. Another field study conducted
by Lee et al. (1995a) compared the performance of inorganic nutrients with organic fish bone-
meal fertilizer. These results showed that the organic fertilizer had the greatest effect on
microbial growth and activity, while the inorganic nutrients were much more effective in crude
oil degradation.
All these results suggest that the success of bioremediation is case specific, depending on oil
properties, the nature of the bioremediation products and the characteristics of the contaminated
environments. Fortunately, recent studies have shown that the oil biodegradation rate depends
on the nutrient concentrations in the pore water of the sediments, which could provide important
guidance for nutrient applications (Bragg et al., 1994, Venosa, 1996). This finding may also
explain why some earlier trials have failed to demonstrate the effectiveness of nutrient
32
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application since the nutrient concentrations in the interstitial pore water had not been monitored
and controlled in most of these studies. Venosa et al. (1996) found that maintenance of a
threshold nitrogen concentration of 1 - 2 mg N/L in the interstitial pore water would result in
close to maximum hydrocarbon biodegradation in a sandy beach (this will be discussed in detail
in the section on case studies). Future research on biostimulation probably should focus on the
determination of the optimal interstitial nutrient concentration and the best strategies to maintain
this concentration for various environments whenever the degradation is limited by nutrient
availability.
In addition to the demonstration of the efficacy of oil degradation, it is also necessary to
demonstrate that bioremediation does not produce any undesired environmental and ecological
effects. There have been concerns that enhanced microbial degradation of oil might produce
toxic metabolic by-products. To address this concern, Lee et al. (1995b) conducted a field study
using different fertilizers to investigate the effect of bioremediation on the toxicity of Venture
condensate stranded on sandy beach sediments. The toxicity of the sediments was monitored
using the Microtox® Solid-Phase Test. The results indicated that sediment toxicity was not
significantly affected by the addition of an inorganic fertilizer (ammonium nitrate and triple-
superphosphate). However, they did observe a slowing of the decrease in toxicity when the
organic fertilizer (fishbone meal) was applied repeatedly, which was attributed to rapid
biodegradation of the fertilizer and the production of ammonia that exceeded toxicity threshold
limits.
Case studies
Four representative field studies are described in more detail here. Three types of marine
shorelines and one freshwater wetland are covered.
Exxon Valdez
Following the grounding of the supertanker Exxon Valdez on Bligh Reef in Prince William
Sound in 1989, U.S. EPA, in conjunction with the Exxon Corporation and the state of Alaska,
embarked on the largest oil spill bioremediation project ever attempted in the field. Extensive
field trials at various sites were conducted, which have been well documented in the literature
(Bragg et al., 1994; Prince, 1993; Prince et al., 1994; Pritchard and Costa, 1991; Swannell et al.,
1996; Venosa, 1998). Important findings and lessons learned from these studies are summarized
as follows.
• Seeding of bioaugmentation products failed to demonstrate enhanced oil biodegradation. It
was found that oil biodegradation on the shoreline of the Prince Williams Sound was limited
by the concentration of nutrients, particularly nitrogen, and not by the absence of
hydrocarbon-degrading microorganisms (Pritchard and Costa, 1991; Venosa et al., 1992).
• Three types of nutrients or fertilizers were tested in the field: a water-soluble inorganic
fertilizer (23:2 N:P garden fertilizer formulation), a slow-release inorganic fertilizer
(Customblen), and an oleophilic fertilizer (Inipol EAP 22). Each was shown to be variably
33
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effective. Inipol EAP22 and Customblen were chosen as bioremediation agents, and
approximately 50,000 kg of nitrogen and 5,000 kg of phosphorus were applied over 120 km
of the oil contaminated shorelines during 1989 and 1992. Within two weeks after the
fertilizer application, the area of cobble beach treated with Inipol EAP22 and Customblen
appeared to be visibly cleaner than the untreated area (Pritchard and Costa, 1991; Pritchard el
al., 1992). However, it was later found that the oil coating the surfaces of the cobble had
been lifted and re-deposited within the interstices in the beach surface.
• Heterogeneous oil distribution on the contaminated beaches made it difficult to determine the
rates of oil biodegradation using established methods. The changes in the ratios of a
hydrocarbon component to a conserved internal standard or biomarker were used as the basis
for determining the oil degradation rate. They also found that previously traditional
biomarkers such as pristane and phytane degraded rapidly in Alaskan beaches. This
observation rendered use of such biomarkers ineffective in permitting firm conclusions on
bioremediation effectiveness. 17a(H),21(3(H)-hopane, the pentacyclic triterpane containing
30 carbon atoms, was used instead. Using hopane as the biomarker, Bragg et al. (1994)
showed that fertilizer application accelerated the rate of oil removal by a factor of
approximately five-fold compared to natural attenuation. This observation was made,
however, from samples repeatedly collected from only one site, so the statistical basis
supporting this conclusion is tenuous.
• Oil biodegradation rate appeared to be dependent on the nitrogen concentration in the pore
water of the intertidal sediments, suggesting that on-site monitoring of nutrients in the
sediment pore waters could provide practical guidance for nutrient applications (Bragg et al.,
1994).
• According to the EPA/Exxon/State of Alaska joint monitoring program, bioremediation was
an environmentally sound remediation technique. This was based on the results of testing the
toxicity of nearshore water to sensitive marine species such as Mysid shrimp, analyzing
ammonium and nitrate concentrations, evaluating the potential of algal growth, and
monitoring oil release into nearshore water after the application of fertilizers (Prince et al,
1994).
• The results of the fertilizer application following the Exxon Valdez spill generally
demonstrate that bioremediation may enhance oil biodegradation on certain marine
shorelines. However, conclusions on the effectiveness of bioremediation in the Exxon
Valdez study are somewhat questionable, in part because the experimental design was not
entirely based on sound statistical principles (Venosa, 1998). Major flaws included the lack
of replication and the attempt to determine too many factors in a limited number of tests,
resulting in the confounding of different effects. The lessons learned from the Exxon Valdez
project led to the replacement of "post Exxon Valdez excitement" with more scientifically-
valid approaches (Mearns, 1997).
34
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Delaware Field Study
The main purposes of this field study were to obtain credible statistical evidence in determining
if bioremediation with inorganic mineral nutrients and/or microbial inoculation enhances the
removal of crude oil, to compute the rate at which such enhancement takes place, and to establish
engineering guidelines on how to bioremediate an oil-contaminated sandy shoreline.
The trial was conducted on a medium- to coarse-grained sandy beach (environmental sensitivity
index = 4) at Fowler Beach, Delaware (located midway between Dover and Rehoboth Beach). A
randomized complete block design was used in the study. Twenty 4 x 9 m plots were set up in
five replicate blocks. Each block contained four treatments in random order, which included: an
unoiled control plot, a no-nutrient control plot (natural attenuation), a plot receiving water-
soluble nutrients, and a plot receiving water-soluble nutrients supplemented with a natural
microbial inoculum from the site. Weathered Nigerian Bonny Light crude oil was intentionally
released onto 15 plots. Nutrients (NaNC>3 and NasP^Oio) were applied daily through a sprinkler
system at a rate designed to achieve a target of 1.5 mgN/L average interstitial pore water
concentrations. Once a week, 30 L of a suspended mixed population of hydrocarbon-degrading
bacteria was also added to the inoculum plots. Sand samples were collected every 14 days from
the 15 oiled plots for oil analysis, and all analytes were normalized to hopane. Nitrate
concentrations in the interstitial pore water of oiled plots were measured each day.
Figure 2.3a shows the hopane-normalized concentrations of total target alkanes (i.e., the sum of
all alkane analytes from n-Cio to n-C35, plus pristane and phytane), while Figure 2.3b shows the
total target aromatics (i.e., the sum of all groups of PAHs and sulfur heterocyclics analyzable by
GC/MS and their alkyl-substituted homologues) in the nutrient-treated, inoculum-treated, and
control plots, all as a function of time. Although substantial hydrocarbon biodegradation
occurred in the untreated plots, statistically significant differences between treated and untreated
plots were observed in the biodegradation rates of total alkanes and total aromatic hydrocarbons.
The results also show that bioaugmentation, even with indigenous organisms, does not stimulate
further degradation of hydrocarbons beyond simple nutrient addition. The studies further
demonstrated that maintenance of a threshold nitrogen concentration of 1 - 2 mg N/L in the
interstitial pore water would permit close to maximum hydrocarbon biodegradation.
Another important conclusion from this study is that background nutrient concentrations at the
contaminated site should be a determining factor in the decision to apply bioremediation. The
background nitrogen concentration at Fowler Beach was high enough to permit close to
maximum hydrocarbon biodegradation without the need to apply additional fertilizer despite the
enhancement observed from nutrient addition. The enhanced effect, although statistically
significant, was not substantial enough to have warranted a decision to implement
bioremediation on a full-scale basis had there been a real spill at this site. This demonstrates that
assisted bioremediation might not always be necessary if sufficient nutrients are naturally present
at a spill site in high enough concentrations to perform natural cleanup. For coastlines having
low natural input levels of nutrients, bioremediation would be appropriate as an alternative
cleanup option.
35
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200
control
nutrients
n 150
inoculum
100
50
0
16
14
12
10
8
M
6
4
2
0
0
10
20
30
40
50
60
70
80
90
100
time, days
Figure 2.3. Results of the Delaware field study: first-order declines in (a) total target alkanes and
(b) total target aromatics
36
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Bioremediation of a Fine Sediment in England
The previous two cases have demonstrated that bioremediation can be effective for cobble and
sandy shorelines, although the extent of the enhancement in effectiveness is determined by the
natural presence of nutrients. However, much less attention has been given to fine sediment
marine shorelines such as mudflats. A recent study conducted by Swannell et al. (1999a) was
intended to fill this gap. One of the objectives of this study was to investigate the potential of
bioremediation in treating buried oil.
The field site was located within the Stert Flats on the southwest coast of England. Twelve plots
were set up on an 80-m stretch of sand composed of 3.2 % mud and 80% particles in the range of
125-180|im, The experimental area was divided into three replicate blocks. Each block
consisted of four randomly assigned treatments that consisted of an unoiled control plot, an
unoiled plot treated with fertilizer alone, an oiled plot with no amendments added, and an oiled
plot treated with fertilizer. The sediment in each plot was retained in mesh enclosures (0.4 m x
0.4 m x 0.05m) and buried at a depth of 15 cm. Weathered and emulsified Arabian Light crude
oil was applied to the appropriate enclosures at 3.7 kg/m2. Inorganic fertilizer (NaNC>3 and
KH2PO4) was applied using a sprinkling device at rates of 2% of N and 0.2% of P by weight of
the initial oil concentration, and at a frequency of once a week for the first four weeks and then
every two weeks thereafter. Samples were taken on Day 7, 49, and 108 for oil analysis, and
hopane was used as a biomarker.
The results showed that more oil was degraded in 2 of the 3 fertilized plots than in the controls
after 108 days. In blocks 1 and 3, the mean total GC resolvable hydrocarbons (TGCRH)/hopane
ratio decreased by 58.4% and 48.4% respectively, compared to 23.0% and 4.4% in the two oiled
controls respectively. In block 2, oil degradation was slower, with only 14% decrease of
TGCRH/hopane ratios in the treated plot vs. no removal in the control plot. Statistical analysis
demonstrated that differences in the ratio of TGCRH and TPH against hopane between the
fertilized plots and the controls were highly significant (p<0.0001). Microbiological analyses
also showed that nutrient addition increased the numbers of hydrocarbon-degraders on the oiled
plots by ten-fold. This study suggested that bioremediation by nutrient enrichment for the
treatment of buried oil in fine sediments may be feasible after an oil spill incident. However,
because of the third replicate failing to confirm the results of the other two, more definitive
conclusions cannot be made from this study.
One deficiency of this study was that by aerating the subsurface sediments during the sediment
burying and periodic excavation of the site due to nutrient application, the potential problem of
oxygen limitation in this environment could not be evaluated. Oxygen limitation is a major
concern for application of oil bioremediation in subsurface sediments, anoxic zones of a water
column, and most fine-grained marine shorelines (Head and Swannell, 1999). Because this
experimental site was a high-energy beach with a tidal bore, oxygen may not have been a
limiting factor for this specific fine sediment beach. Conclusions from this study, therefore,
should only be confined to the Stert Flats and not extrapolated to similar environments. Further
37
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research is still required to determine the effectiveness of bioremediation in other types of fine
sediments.
Bioremediation of a Freshwater Wetland in Canada
A field study was conducted by the U.S. Environmental Protection Agency, University of
Cincinnati, and Fisheries and Oceans Canada at a freshwater wetland site situated along the St.
Lawrence River, Ste. Croix, Quebec, Canada. The objective of this study was to determine the
effectiveness of biostimulation strategies in accelerating restoration of an oil-contaminated
freshwater wetland site. (Garcia-Blanco et al., 2001b, Venosa el al., 2002). Strategies that were
evaluated were bioremediation by nutrient enrichment in the presence and absence of vegetative
growth of the dominant plant species, Scirpus pungens. Twenty 5m x 4m plots were set up in the
upper intertidal zone of a study site located along the St. Lawrence River, where the water was
far enough away from the Atlantic to be still fresh, yet was tidally influenced. These plots were
divided into four replicate blocks. Each block consisted of five randomly assigned treatments,
which were: (1) oiled, no added nutrients, with intact plants (natural attenuation control), (2)
oiled, NH4NO3 + Ca(H2PC>4)2 H20 added, with all vegetative growth cut back to ground surface
daily to suppress plant growth, (3) oiled, NH4NO3 + Ca(H2P04)2 H20 added, with intact Scirpus
pungens, (4) oiled, NaNC>3 + Ca(H2PC>4)2 H20 added, with intact Scirpus pungens, and (5)
unoiled, NH4NO3 + Ca(H2PC>4)2 H20 added, with intact Scirpus pungens (background control).
Weathered Mesa light crude oil was released onto each plot earmarked for oiling. The amount of
oil released was 12 L per plot. Composite core samples were collected after 0, 1,2, 4, 6, 8, 12,
16, and 21 weeks for quantification of the remaining oil constituents by gas
chromatography/mass spectroscopy (GC/MS) operating in the Selected Ion Monitoring mode or
SIM. To account for differences due to physical washout, all oil constituents were normalized to
hopane.
Figure 2.4 illustrates results from the hopane-normalized concentrations of total target alkanes
and total target PAHs for the four treatments as a function of time. Although the bioremediation
and phytoremediation treatments achieved slightly better degradation of hydrocarbons than
natural attenuation, no statistically significant evidence of stimulation through addition of
nutrients or biodegradation enhancement by vegetation was observed. After 21 weeks, reduction
of target parent and alkyl-substituted polycyclic aromatic hydrocarbons (PAHs) averaged 32% in
all treatments. Reduction of total target alkanes was of a similar magnitude. The pattern of
disappearance of hydrocarbons was characteristic of biodegradation; namely, the lower
molecular weight alkanes declined to a greater extent than the higher carbon-number alkanes, as
did the lower molecular weight PAHs compared to the higher molecular weight PAHs. Since
there was little evidence to support enhancement of biodegradation by nutrient addition with and
without vegetation, it was suggested that oxygen limitation was most likely the dominant cause
of the persistence of oil hydrocarbons on the oil-contaminated plots.
38
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600
500
400
300
200
100
0
125
100
75
50
25
0
1 R
Natural Attenuation
Bio + Ammonium
Phyto + Ammonium
Phyto + Nitrate
—•— Natural Attenuation
~ Bio + Ammonium
Phyto + Ammonium
—~— Phyto + Nitrate
i i i i i i i i i
0 1 2 4 6 8 12 16 21
Time (weeks)
suits of St. Lawrence River field study: concentration declines in (a) total
total target PAHs
39
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While the Ste. Croix results indicate that biostimulation might not be an effective strategy to
mediate the removal of residual oil from the sediments, significant changes in biological
measures of habitat were observed (Lee et al., 2001a). For example, S. pungens, the dominant
plant species, was tolerant to the oil, and its growth was significantly enhanced above that of the
unoiled control by the addition of nutrients. Other biotest organisms (bacteria, Vibrio sp.,
invertebrates, Daphnia, Hyalella, and Viviparus sp) provided additional evidence of both
enhanced recovery and potential detrimental effects. The study indicates that on an operational
scale, natural attenuation may be the most practical treatment option for oxygen-limited
freshwater wetlands.
2.5.3 Kinetics of oil bioremediation
Knowledge of the kinetics of oil biodegradation under different environmental conditions is
important for assessing the potential fate of targeted compounds, evaluating the efficacy of
bioremediation, and determining appropriate strategies to enhance oil biodegradation. Oil
biodegradation rates are difficult to predict due to the complexity of the environment. The rates
of biodegradation vary greatly among the various components of crude oils and petroleum
products. The presence of other substrates may affect the degradation rates of the compounds of
interest. Environmental factors such as temperature, nutrient concentrations, and oxygen tension
also influence the kinetics of oil degradation. The heterogeneity of oil distribution on shorelines
or wetland sediments makes kinetics studies even more difficult.
Very few kinetic studies on oil degradation under field conditions have been conducted. The
Exxon Valdez monitoring program developed a multiple regression model based on field studies
conducted by researchers from Exxon (Prince et al., 1994). The best-fitting model was
expressed as:
where Ch(t) is the time-varying hopane-normalized concentration of an analyte, p is the polar
fraction of the oil, r is the ratio of the average residual nitrogen concentration to oil loading, and
e is the assumed multiplicative error term, while a, 8, y, and co are fitting parameters determined
from the multiple regression analysis. This model matched the experimental results in Alaska
well when the parameters are chosen to fit the data. However, its potential for process
understanding and prediction is limited because the data set used for the regression was limited
to only one small, non-replicated area in the field.
Venosa et al. (1996) developed from field data first-order biodegradation rate constants for
resolvable alkanes and important two- and three-ring PAH groups present in light crude oil. The
first order relationship was expressed as:
Ch(t) = oc[l-p(t)fe
|Ye 8r(t)+fflt £
(2.2)
(2.3)
40
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where (A/H) is the time-varying hopane-normalized concentration of an analyte, (A/H)0 is that
quantity at time zero, and k is the first-order biodegradation rate constant for an analyte.
For the field study conducted in Delaware, first-order biodegradation rate coefficients ranged
from 0.026 to 0.056 day"1 for total resolvable alkanes and from 0.021 to 0.031 day"1 for total
resolvable PAHs (Venosa et al., 1996).
Actual lst-order biodegradation rates are not constant, however. Instead, they are a function of
the residual nutrient concentration:
f
obs
N
\
kK„+N,
(2.4)
where kobs and kmax (T"1) are the observed and maximum first-order hydrocarbon biodegradation
rates, respectively, Kn (MnL"3) is the half-saturation concentration for a specific nutrient, and N
(MnL"3) is the interstitial pore water residual nutrient concentration. Experiments conducted at
the University of Cincinnati showed that the Kn for nitrate is approximately 0.5 mg N/L
(unpublished). The model that incorporates Equation 2.3 and 2.4 will be very useful in the
experimental design and the performance prediction for the bioremediation of oil contaminated
shorelines.
Studies have also been conducted to compare oil biodegradation rates obtained in laboratory tests
to those calculated from the Delaware field study (Venosa et al., 1996, 1997a; Holder et al.,
1999). Venosa et al. (1996, 1997a) found that the degradation rates of all target alkanes and
aromatics in a light crude oil were close to an order of magnitude lower in the field compared to
results from the laboratory. However, when the rate data of PAHs were normalized to the highest
alkyl-substituted homologue in each given PAH series, the first order rate constants in the field
were nearly identical to rate constants from the lab. These relationships were consistent even
with different microbial consortia isolated from eight different marine shorelines of the United
States. Similar results were also observed in a laboratory study using 14 different marine and
freshwater consortia (Holder et al., 1999).
Simon et al. (1999) reported some kinetic data derived from a study conducted in a coastal
wetland contaminated with Arabian light crude oil. First-order biodegradation rate coefficients
ranging from 0.017 to 0.061 day"1 for total target saturates and from 0.009 to 0.027 day"1 for total
target aromatics were reported. These rate coefficients were similar to those of Venosa et al.
(1996). Further research is still required to develop more state-of-art models and to establish a
database of kinetic parameters for different types of oil under various marine shoreline and
freshwater wetland environments.
2.6 Nutrient Hydrodynamics
Since nutrient addition has been found to be the most effective bioremediation strategy in aerobic
environments, particularly for marine shorelines, a full understanding of the fate of water soluble
41
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nutrients on marine beaches and the hydrodynamics controlling their transport and persistence is
necessary. One of the main challenges associated with biostimulation in oil-contaminated
coastal areas is maintaining optimum nutrient concentrations in contact with the oil and the
degrading microorganisms. Oil from offshore spills usually contaminates the upper third of the
intertidal zone, where the washout rate for water-soluble nutrients can be very high. Various
oleophilic and slow-release nutrient formulations have been developed to improve the contact
between oil and nutrients within the environment. However, most slow-release and many
oleophilic fertilizers rely on dissolution of the nutrients into the aqueous phase before they can
be used by hydrocarbon degraders (Safferman, 1991). Thus, design of effective oil
bioremediation strategies and nutrient delivery systems requires an understanding of the
transport of water-soluble fertilizers in a beach ecosystem.
Dissolved nutrients are expected to move with the water in the beach sand. Water flow through
the porous matrix of a beach is driven by a combination of three main factors (Boufadel et al.,
1999b; Wrenn et al., 1997): (1) tides that result in rise and fall of beach groundwater level
(typically, the water level in a beach tracks the level of the rising tide with only a slight lag, but
the beach drains much more slowly when the tide ebbs because of resistance from the porous
matrix), (2) wave action that operates through two main mechanisms (at wave runup, water
enters the beach and percolates vertically through the unsaturated zone until it reaches the water
table; at wave rundown, water moves in a predominantly horizontal direction and exits at the
water line), and (3) flow of fresh groundwater from coastal aquifers, which causes continuous
horizontal advective flow from the beach face at or near the water line (Glover, 1959). This type
of groundwater flow can interact with tidal fluctuations to produce complex variations in the
groundwater level within the beach (See Section 2.6.1).
2.6.1 Nutrient transport in beaches: a mesocosm study
Beach hydraulics and nutrient hydrodynamics were investigated through both theoretical and
experimental approaches by Boufadel et al. (1999 a and b, 1998, 1997). A two-dimensional
finite element model for water flow and salt transport in saturated and unsaturated porous media
was developed. The model also considers the effects of salt concentration on water density and
water viscosity. An experimental wave tank was used to validate the model and to investigate
cases that cannot be simulated by the numerical model. Tracer tests were carried out to
investigate the separate and combined effects of tide, waves, and buoyancy on the transport of
soluble inorganic nutrients in sand beaches. The major findings from these studies were:
• In the absence of regional seaward groundwater flow, the tide generated a predominantly
downward and seaward hydraulic gradient that caused the washout to the sea of nutrients
applied to the top section of beaches. The presence of waves under these conditions
accelerated the washout, the rate of which was found to increase by about 30% when waves
were superimposed on the tide.
• Beach geometry plays a major role in beach hydraulics and hydrodynamics because the flow
lines are perpendicular to the beach surface. Under tidal action, seawater enters the beach
42
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from above, flowing vertically downward, and may cause the entrapment of less saline water
in the beach. This finding is very important because systems that rely on continuous or
intermittent injection of nutrient solutions through trenches or horizontal wells installed in
the supratidal zone rely on subsurface flow to carry the nutrients through the contaminated
area (Wise et al., 1994). The approach that was proposed by Wise et al. (1994) assumes that
nutrients dissolved in the freshwater plume will be brought into contact with the oiled beach
material periodically by the rising tide, because the freshwater plume should float on top of
the saltwater. The finding of freshwater trapped between two saltwater wedges indicates that
subsurface injection of nutrients may not be an effective method for providing nutrients to
the bioremediation zone when a freshwater plum exists because the impact zone will never
be exposed to the nutrients.
• Different nutrient application strategies at low tide were investigated in the mesocosm beach
since application of nutrients at the beach surface at low tide (Venosa et al., 1996) appears to
be the best application strategy for biostimulation. The results showed that applying nutrients
at the beach surface pre-dissolved in water resulted in generally longer residence times and
larger spreading of the nutrient plume in the top section of the beach compared to applying
nutrients in granular form at the beach surface and hosing them in with a water spray.
However, it should be noted that in practice, addition of granular fertilizer may be prudent in
cases where it is not possible to sprinkle pre-dissolved nutrients as the application method.
2.6.2 Nutrient transport in beaches: field trials
To verify the findings from the mesocosm studies and to further investigate nutrient transport
under field conditions, tracer studies were conducted in the intertidal zone of three different
marine beaches in Delaware and Maine (Suidan and Wrenn; 2001; Wrenn et al., 1997a & b).
The Delaware Tracer Study
The purpose of this study was to characterize the transport of water-soluble nutrients in the
intertidal region and to estimate their washout rates from the bioremediation zone (i.e., the oil-
contaminated area). The study was conducted on a moderate-energy, sandy beach (Slaughter
Beach) on Delaware Bay. The typical wave heights at this beach were between 15 to 30 cm. A
conservative tracer (LiNCh) was applied to eight replicate 5 m x 10 m plots in the upper
intertidal zone at low tide during full moon spring tide and the last-quarter moon neap tide. The
tops of the plots were placed approximately at the spring high tide line. Sand samples were
collected for tracer analysis.
This study showed that the rate of tracer washout from the bioremediation zone was more rapid
when the tracer was applied at spring tide (when the tidal amplitude is largest) than at neap tide.
When the conservative tracer (LiNCh) was applied to the beach surface in the upper intertidal
zone at the full moon spring tide, it was completely removed within one day. When it was
applied at neap tide, however, the tracer persisted in the bioremediation zone for several days.
The results indicated that the amount of nutrient remaining in the bioremediation zone was
highly correlated with the maximum extent to which the treated area had previously been
43
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submerged by water at high tide; submergence resulted in nearly complete removal of dissolved
compounds from the bioremediation zone. Therefore, the fertilizer release rates should be
designed to achieve optimal nutrient concentrations while the tide is out. The Delaware tracer
study indicates that nutrient transport in sandy beaches is driven by tidally influenced hydraulic
gradients and wave activity. However, it is impossible to clearly separate the influence of tide
and wave action in this study.
The Maine Tracer Study
The main purpose of this field study was to separately evaluate the effect of tide and wave action
on nutrient transport since the results of the Slaughter Beach field studies could not distinguish
between these two processes. This was accomplished by comparing the transport rate and
characteristics of a dissolved conservative tracer (lithium nitrate) on a high-energy beach to those
on a low-energy beach to determine how waves affect solute transport (Suidan and Wrenn, 2001;
Wrenn et al., 1997b). The two beaches are located in the town of Scarborough, in southern
Maine. Scarborough Beach is a high-energy beach that faces the Atlantic Ocean with average
heights of breaking waves between 0.3 to 1 m during the course of this study, whereas Ferry
Beach faces a protected harbor with the typical wave height of less than 3 cm. The tidal range at
both beaches was essentially the same. Dissolved tracer (LiN03) was applied to four replicate
plots on each beach at low tide during full moon spring tide and the last-quarter moon neap tide.
Both water and sand samples were collected for tracer analysis.
Washout of lithium from the upper intertidal zone during the spring-tide experiment is shown in
Figure 2.5. The differences between the two beaches are very clear. Whereas lithium was
completely removed from the entire experimental domain within two days on the high-energy
beach, more than two weeks were required to achieve the same degree of washout from the low-
energy beach. Washout during the neap-tide experiment was much slower than that observed
during the spring-tide experiment. Lithium was completely removed from the experimental
domain on Scarborough Beach within one week, whereas the total mass of lithium was
essentially unchanged for more than two weeks at Ferry Beach. Slower washout was expected
during the neap-tide experiment, because only the bottoms of the plots were covered by water at
high tide during most of the first five or six days. Since plot coverage by the tide was essentially
the same on the two beaches during the first week of this experiment, the differences in washout
rate must have been due primarily to wave activity.
The Maine field study clearly shows that the washout rate of nutrients from the bioremediation
zone is strongly affected by the wave activity of the contaminated beach. Wave action in the
upper intertidal zone may cause nutrients from the surface layers of the beach to be diluted
directly into the water column, resulting in their immediate loss from the bioremediation zone.
On the other hand, washout due to tidal activity alone is relatively slow, and nutrients will
probably remain in contact with oiled beach material long enough to effectively stimulate oil
biodegradation on low-energy beaches.
44
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Salinity distribution was also examined in the Maine study. In both beaches, the "sandwiching
phenomenon" (a layer of lower salinity water located between two higher salinity water layers)
was observed, which was confirmed by Boufadel et al. (1999b) in the mesocosm study. Both
these mesocosm and field studies are very helpful in providing guidelines for the optimal
applications of nutrients on marine beaches, which will be discussed later in this document.
Although many questions remain unanswered, we have made tremendous progress in
understanding various aspects of oil bioremediation in the last decade. The development of an
operational guideline for bioremediation of oil contaminated marine shorelines and freshwater
wetlands is not only possible but also necessary in ensuring that full-scale cleanup in the future
proceeds rapidly and efficiently.
46
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Chapter 3 Methods Used in Monitoring Oil Bioremediation
In order to demonstrate that biodegradation is taking place in the field, the chemistry or
microbial population must be shown to change in ways that would be predicted if bioremediation
were occurring (NRC, 1993). Environmental conditions, particularly nutrient concentrations,
should also be monitored for evaluating the effects of bioremediation. Some of the most
important methods used in monitoring oil bioremediation will be overviewed in this chapter.
References will be given for detailed descriptions of methods.
3.1 Analytical Methods
3.1.1 Microbiological analysis
Existing methods for microbial analysis can be classified into cultured-based techniques and
culture-independent techniques. Commonly used microbial analysis methods in monitoring oil
bioremediation are summarized in this section and table 3.1.
3.1.1.1 Enumeration of hydrocarbon-degrading microorganisms: culture based techniques
Microbial counts are often used to monitor the bioremediation process. In general, the more
microbes, the more quickly the contaminants will be degraded. Correlating an increase in the
number of contaminant-degrading bacteria above normal field conditions is one indicator that
bioremediation is taking place. Analysis of the microbial communities that take part in in-situ
hydrocarbon biodegradation activities has been a challenge to microbiologists (Macnaughton et
al., 1999). The reason for this is that most (-90 to 99%) of the species making up competent
degrading communities do not form colonies when current laboratory-based culture techniques
are used (Rollins and Colwell, 1986; Rozsak and Colwell, 1987; Wilkinson, 1988). The
techniques are briefly described below.
Plate count
Plate count is a traditional technique, which quantifies the number of bacteria capable of growing
on a prescribed set of nutrients and substrates in a solid medium, by counting the colonies
formed (National Research Council, 1993). The general procedure involves (1) making the solid
medium or gel from a liquid solution with appropriate nutrients and substrates, using a
solidifying agent like agar, (2) Spreading a sample containing the bacteria of interest thinly over
the surface of the gel in plates, (3) Incubating the plates, (4) counting the bacterial colonies
formed. Each colony is assumed to have arisen from a single bacterial cell.
A number of studies have used hydrocarbon incorporated into either agar-based or silica-based
media to enumerate of hydrocarbon-degrading microorganisms (Horowitz and Atlas, 1978;
Sexstone and Atlas, 1977; Walker and Colwell, 1976). However, other researchers reported that
plate counts are unsuitable for enumerating hydrocarbon-utilizing microorganisms because many
marine bacteria can grow and produce micro-colonies on small amounts of organic matter
existing in the solid media, resulting in the counting of non-hydrocarbon utilizers (Atlas, 1981;
Higashihara, et al., 1978). Plate counts also underestimate the number and diversity of bacteria
because of the difficulty in enriching viable colonies from environmental samples. Culturable
47
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techniques have been found to be inferior to techniques that do not rely on viable culturing for
enumeration (Macnaughton et al., 1999).
Most-probable-number (MPN) procedures
MPN procedures have been viewed as a more reliable method for enumerating hydrocarbon-
utilizing microorganisms because such procedures eliminate the need for a solidifying agent and
permit direct assessment of the ability to actually utilize hydrocarbons (Atlas, 1981; Wrenn and
Venosa, 1996). MPN procedures use liquid nutrient media in test tubes or microtiter plates and
hydrocarbons as the sole carbon source. The enumeration is carried out through a statistical
analysis based on the numbers of a series of diluted liquid samples that show evidence of
bacterial growth. This evidence of bacterial growth can be established based on turbidity, release
of 14C02 from radiolabeled hydrocarbons, disruption of oil sheen, and reduction of dyes (Rice
and Hemmingsen, 1997). Either statistical tables (Eaton et al., 1995) or a computer program
(Klee, 1993) can be used to determine the MPN.
Most existing MPN procedures use crude oil or a refined petroleum product as the selected
hydrocarbons, which can not distinguish different groups of hydrocarbon degraders. For
example, the Sheen Screen Method uses dispersion or emulsification of the crude oil substrate to
identify positive wells (Brown and Braddock, 1990). But these effects are associated primarily
with growth on aliphatic hydrocarbons (Hommel, 1990). Aliphatic hydrocarbons are of less
environmental concern than PAHs, however, because they are less toxic and are biodegraded
more rapidly. Wrenn and Venosa (1996) recently developed an MPN procedure that can
separately enumerate aliphatic and aromatic hydrocarbon-degrading bacteria. The size of the two
populations is estimated using separate 96-well microtiter plates. The alkane-degrader MPN
method uses hexadecane as the selective growth substrate, and positive wells are detected by
reduction of iodonitrotetrazolium violet, which is added after incubation for 2 weeks at 20°C.
PAH degraders are grown on a mixture of PAHs in separate plates. Positive wells turn yellow to
greenish brown from accumulation of the partial oxidation products of the aromatic substrates
after 3 weeks incubation. This method is simple enough for use in the field and provides reliable
estimates for the density and composition of specific hydrocarbon-degrading populations.
For a detailed description of existing MPN procedures, readers can refer to Rice and
Hemmingsen (1997), and Wrenn and Venosa (1996).
3.1.1.2 Culture-independent population/community techniques
The main challenge for accurate analysis of hydrocarbon-degraders using existing culture-based
techniques is that most these species are not able to be cultured (Atlas & Bartha, 1987,
Macnaughton et al., 1999). The emergent culture-independent molecular techniques have made it
possible to identify the diversity and composition of uncultivated microbial communities and to
enumerate bacteria in more precise ways.
Phospholipidfatty acid (PLFA) analysis
Phospholipid fatty acid (PLFA) analysis is based on the characteristic "signature" of fatty acids
present in the membranes of all cells (National Research Council, 1993). The distribution of fatty
acids is unique and stable. Therefore, it can be used as an identifying index. Determination of
48
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biomass through analysis of the extractable lipids avoids culture bias. This technique also
provides a quantitative means to measure viable biomass, community composition, and
nutritional stature (White et al., 1998).
Phospholipids can be extracted from the sample and the phosphate can be measured by
colorimetric techniques (Findlay et al., 1989). The results can represent the amounts of viable
cells and biological activities in the sample. A more powerful PLFA method involves extraction
and separation of lipid classes into neutral-, glyco-, and polar-lipid fractions, followed by
quantitative analysis using gas chromatography/mass spectrometry (GC/MS) (Macnaughton et
al., 1999; White et al., 1998). This procedure can quantitatively determine the characteristics of
microbial communities. However, PLFA analysis cannot identify species composition.
Nucleic acid-based molecular techniques
Nucleic acid-based molecular techniques can identify bacterial species by the unique sequence of
molecular codes in their genes. One of most useful methods for determining the diversity of
bacterial communities is denaturing gradient gel electrophoresis (DGGE) (Muyzer et al., 1993).
The method provides a means of separating the PCR (polymerase chain reaction) products from
mixed cultures based on the melting properties of the DNA. Usually the 16s rDNA portion of
the bacterial genome is targeted for PCR amplification, since this region is commonly used for
bacterial identification. PCR primers can be designed to detect a broad range of bacteria
(universal) or can be designed just for a specific group of interest.
The DGGE gel is made of acrylamide, and contains a gradient of formamide and urea, which
both act to denature, or pull the strands of the DNA apart. The PCR products are loaded onto the
gel, and a voltage is established across the gel for several hours. As DNA, which carries a net
negative charge, is carried through the gel, it encounters an increasing gradient of denaturant,
which causes the DNA chains to separate and the effective size of the molecules increases,
causing movement through the gel to cease. The end result is a DNA banding pattern, where
DNA requiring more chemical potential to denature travels farther, and DNA requiring less
chemical potential to denature stays near the top of the gel. Each DNA band approximately
corresponds to the presence of one kind of organism in the mixed culture. This banding pattern
is sometimes referred to as a "community fingerprint," and allows for a quick approximation of
number of bacterial species (diversity) present. Bands can be excised for sequencing analysis,
and sequences can be compared to the Ribosomal Database Project (Maidak et al., 2000)
containing the 16S rDNA sequences of currently known organisms.
The use of DGGE for quantitative purposes is still not well established. However, it can be used
in conjunction with other quantitative methods such as PLFA analysis to provide insight into
microbial species distribution. The PLFA-DGGE techniques were successfully used in
determination of microbial population changes during the Delaware field study on oil
bioremediation (Macnaughton et al., 1999).
49
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Table 3.1 Commonly used microbial analysis methods in monitoring oil bioremediation
Method
Advantages
Disadvantages
Plate counts
Well-established and easy to
perform
Counts only organisms viable
on solid-media and may allow
growth of non-hydrocarbon
degraders
Most probable number (MPN)
techniques
More reliable method for
enumerating hydrocarbon-
utilizing microorganisms since
hydrocarbons are used as the
sole carbon source. Some of
the procedures can separately
enumerate aliphatic and
aromatic hydrocarbon-
degrading bacteria
Relatively labor-intensive
(large number of incubations)
and time consuming, and may
still be subject to culture bias
Phospholipid fatty acid
(PLFA) analysis
Eliminate culture bias and
quantitatively determine
viable biomass, community
composition, and nutritional
stature
Require specified knowledge
and equipment, and can not
identify species composition
Denaturing gradient gel
electrophoresis (DGGE)
Identify species distribution
without culture bias
Also require more specified
knowledge and equipment,
and quantitative analysis is
still not well-established
3.1.2. Chemical analysis of nutrients
Since oil biodegradation is limited by availability of nutrients in most marine shorelines,
monitoring nutrients, particularly the nutrient concentrations in pore water, is critical in
developing proper bioremediation strategies and assessing the effect of oil bioremediation
(Bragg et al., 1994; Venosa el al., 1996). Important nutrient analyses include measurements of
ammonium, nitrate, nitrite, and phosphorus. Commonly used methods for these nutrient analyses
are summarized as follows.
Sample preparation
Analysis of nutrients in sediments can either be conducted on site or be frozen and shipped to a
lab for measurements. Before the analysis, available nitrogen and phosphorus species can be
extracted from sediments using a 2M solution of KC1 or an acidified 0.1% NaCl solution (Page
et al., 1986; Tan, 1996). Total nitrogen and phosphorus can be liberated from sediment samples
by persulfate digestion at 121°C.
50
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Ammonia analysis
The most commonly used methods for ammonia analysis are automated colorimetric methods
due to their high sensitivity and ease of use. Two major automated colorimetric methods are
available (1) Automated phenate method (4500-NH3 H, Eaton el al., 1995) (2) Automated
salicylate-hypochlorite method. The latter was recently developed due to the environmental
concerns associated with the phenol used by the phenate method (Tan, 1996). The extracted
ammonia nitrogen can also be measured in the field using an ammonia-selective electrode (4500-
NH3 F, Eaton et al., 1995) or some commercial kits, such as a Chemetrics kit (Chemetrics Inc.,
Calverton, VA) although these methods are less sensitive and more susceptible to interference
than colorimetric methods.
Nitrate and nitrite analysis
Commonly used techniques for nitrate analysis include the ultraviolet spectrophotometric
method (4500-NO;?" B, Eaton et al., 1995), automated cadmium reduction method (45OO-NO3" F,
Eaton et al., 1995) and nitrate electrode method (45OO-NO3" D, Eaton et al., 1995). The
ultraviolet spectrophotometric method is suitable for rapid measurement or screening samples
that have low organic matter contents. The automated cadmium reduction method is more
sensitive and suitable for nitrate analysis in various types of water and wastewater. When using
this method, nitrate is reduced to nitrite by passing through a Cu-Cd reduction column. The
nitrite is then determined using a colorimetric procedure. Therefore, without the reduction step,
this method can be used for nitrite analysis. Total nitrogen can also be measured as nitrate by the
automated cadmium reduction method following oxidation of urea, ammonia, and nitrite by
potassium persulfate at 121 C (4500-Norg D, Eaton etal., 1995). The nitrate electrode method can
be used in the field for rapid nitrate analysis although it is less reliable than cadmium reduction
method.
Phosphorus analysis
Phosphorus analysis involves two general procedures: (1) conversion of the phosphorus form of
interest to dissolved orthophosphate and (2) colorimetric determination of dissolved
orthophosphate (Eaton et al., 1995). Total phosphorus analysis usually uses persulfate digestion
procedures (method (4500-P B, Eaton et al., 1995). Various digestion procedures have been
developed for analysis of Available Phosphorus (Tan, 1996), a variable concept that reflects the
amount of phosphorus available to plants or microorganisms. The ascorbic acid method (4500-P
E, Eaton et al., 1995) is recommended for analysis of phosphate in the concentration range of
0.01 to 6 mg P/L. Commercial kits, such as Hach® phosphate analysis procedures (Hach
Company, Loveland, CO), are also available for use in the field.
3.1.3 Chemical analysis of oil and oil constituents
One of the primary measures of the success of bioremediation treatments is reduction in the
concentrations of spilled oils and target oil constituents in particular. Various techniques have
been developed and used in petroleum hydrocarbon analysis, which include gravimetric
methods, infrared spectroscopy (IR), gas chromatography/flame ionization detection (GC/FID),
gas chromatography-mass spectrometry (GC/MS), and thin-layer chromatography-flame
ionization detection (TLC-FID). Oil analysis methods can be generally classified into two
51
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categories: nonspecific methods to measure total petroleum hydrocarbons (TPHs), and specific
methods using various chromatographic techniques to quantify target oil constituents.
Commonly used oil analysis methods in monitoring oil bioremediation are summarized in this
section and table 3.2.
3.1.3.1 Total petroleum hydrocarbon (TPH) techniques
TPH techniques mainly include gravimetric and infrared spectroscopic methods. These
techniques are widely accepted methods to rapidly quantify the oil due to their simplicity and
low costs. However, these methods provide little information about oil components, exhibit high
detection limits and are susceptible to various interferences (Douglas et al 1991, Xie el al.,
1999). Furthermore, they are unable to distinguish between abiotic and biotic losses.
Gravimetric analysis involves solvent (e.g., dichloromethane) extraction, evaporation, and
gravimetric measurement (EPA 413.1 and EPA 9071). This method does not distinguish between
petroleum hydrocarbons and naturally occurring biogenic compounds (such as plant lipids) and
may result in overestimating TPH. Infrared spectroscopy (IR) involves solvent extraction
normally using trichlorotrifluoroethane (Freon 113). TPH is subsequently measured by
comparing the infrared absorption of the extraction liquid against that of a defined hydrocarbon
mixture (e.g., EPA 418.1 (U.S. EPA, 1992)). Although this technique is a more sensitive
measure of hydrocarbons than gravimetric methods, it may also overestimate or underestimate
TPH for a variety of reasons (Xie et al., 1999). Environmental concerns regarding the use of
Freon as a solvent is also a potential problem for its application (Romero and Ferrer, 1999).
Fully halogenated solvents such as tetrachloroethene or tetrachloromethane can be substituted for
Freon.
Commercial kits, such as Petroflag® test kit (Dexsil, Hamden, CT) and Hach® DR/2000 test kit
(Hach Company, Loveland, CO), are also available for use in the field with limited reliability
(Lambert et al., 1999a&b). All TPH techniques are severely affected by the spatial
heterogeneity. A larger quantity of oil at one spot could give a misleading high TPH value. They
do not distinguish between abiotic and biotic losses that are important in correctly interpreting
the data for the fate of the petroleum in the environment. So, one has to be careful in carrying out
these tests as well as in interpreting the data. A sufficiently large number of samples may help to
overcome some of the variability.
3.1.3.2 Analysis of specific oil constituents
To assess the effect of oil bioremediation, identification and quantification of individual oil
components and compounds, particularly those constituents that are of significant environmental
concern , is required. Various chromatographic techniques, particularly GC/FID and GC/MS, are
widely used to provide specific and sensitive analysis of oil constituents (Douglas et al., 1994;
Wang et al. 1997).
Gas chromatography/flame ionization detection (GC/FID)
GC/FID combines chromatographic separation of hydrocarbon fractions on a capillary GC
column with quantification by FID (e.g., EPA 8100(U.S. EPA, 1992)). Pretreatment of oiled
52
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sediment samples involves drying of samples, addition of surrogates, and solvent extraction (e.g.,
EPA 3540c (U.S. EPA, 1992)). This method has been mainly used for detection of aliphatic
hydrocarbons such as individual Cioto C35 n-alkanes, and isoprenoid hydrocarbons (Table 3.3).
GC/FID can also be used to determine TPHs by the method of internal standards (Douglas et al,
1994). Total GC-detectable hydrocarbons make up more than 50% of the oil components (Prince
et al., 1994). However, GC/FID may not be used to identify and quantify PAHs and biomarker
compounds because it can not clearly separate many of these compounds, especially the
alkylated PAHs. Since PAHs are of most environmental concerns and biomarker compounds are
critical in distinguishing between biodegradation from physical weathering processes and
reducing spatial variability of oil data, measurement of these compounds is very important in
monitoring oil bioremediation.
Gas chromatography-mass spectrometry (GC/MS)
GC/MS, which combines chemical separation by GC and spectral resolution by MS, is the choice
of methods for specific compound determination, especially for identification and quantification
of PAHs and biomarkers (Wang et al., 1997). The mass spectrometer is often operated in the
selected ion monitoring mode (SIM) to further increase sensitivity and selectivity relative to
conventional full-scan GC/MS. For a detailed description of GC/MS analytical procedures,
readers can refer to Douglas et al. (1994), EPA Method 8270 (US EPA, 1989), and Venosa et al.
(1996). Typical target alkanes and PAHs analyzed using GC/MS in recent USEPA-sponsored
projects are listed in Table 3.3 and 3.4 (Purandare, 1999; Venosa et al., 1996). Because the
distribution of oil on shorelines contaminated by offshore spills can be highly heterogeneous
(Bragg et al., 1994), the concentrations of all target analytes are often reported relative to a
conservative biomarker such as 17a(H),21(3(H)-hopane (Douglas et al., 1994). Detailed
discussions on biomarkers will be presented in Section 3.2.
Thin-layer chromatography flame ionization detection (TLC-FID)
Thin layer chromatography-flame ionization detection (TLC-FID), which uses a special
instrument called the Iatroscan, separates hydrocarbons on a Chromarod thin layer based on
characteristic chemical types or fractions such as aliphatic, aromatic, polar, and asphaltene
compounds (Goto et al., 1994). This method has advantages in measuring high-boiling-point
hydrocarbons such as higher molecular weight saturates, aromatics, resins, and asphaltenes,
some of which may not be detectable by GC or HPLC. Unlike the other analytical techniques,
which are either too gross (e.g. TPH techniques) or very specific (e.g. GC based methods), the
TLC-FID can measure the relative percentages of the four major fractions of petroleum in a short
period of time. This method has been successfully used for monitoring oil bioremediation in a
wetland environment (Stephens et al., 1999). However, TLC-FID can not identify specific
compounds and may only be used as a screening tool. Controversy also exists concerning the
reliability of the methods due to some confusing results, and modifications of the technique have
been suggested (Cebolla et al., 1998).
53
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Table 3.2 Commonly used methods for petroleum hydrocarbon analysis
Method
Advantages
Disadvantages
References
TPH techniques
Inexpensive and easy to
perform. Used as quick
screen tools
Low sensitivity and
selectivity; can not be
used for identification of
oil components. Not
recommended.
USEPA, 1992
GC/FID
A specific method used
for detection of aliphatic
and a limited number of
aromatic hydrocarbons.
Unable to identify and
quantify alkylated PAHs
and biomarkers
Douglas et al.,
1994; Wang et
al., 1997
GC/MS
Highly sensitive and
selective method for
identification and
quantification of a wide
range of hydrocarbons,
including PAHs and
biomarkers
Expensive equipment and
complicated procedures
Douglas et al.,
1994; USEPA,
1992; Venosa et
al., 1996; Wang et
al., 1997
TLC/FID
Quick detection of a wide
range of oil components,
including high molecular
weight saturates,
aromatics, resin, and
asphaltenes; can be used
as an effective screening
tool.
Unable to identify
specific compounds;
quantitative analysis is
still not well-established
Stephens et al.,
1999; Cebolla et
al., 1998
54
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Table 3.3 Target alkanes list (
3urandare, 1999)
Compound Name
Misc.
Info.
Qlon
Response Factors
Reference
Compounds
Internal
Standards
D26 n-dodecane
surrogate
66
D26 n-dodecane
D22 n-
decane
Qlon: 66
wCIO
57
wCIO
nCll
57
wCll
nC12
57
wC12
nC\3
57
wC13
nC\4
57
wC14
nC\5
57
wC15
D36 n-heptadecane
surrogate
66
D36 n-heptadecane
D34
n-hexadecane
Qlon: 66
nCll
57
wC17
Pristane
57
Pristane
«C18
57
wC18
Phytane
57
Phytane
nC\9
57
wC19
D50 n-tetracosane
surrogate
66
D50 n-tetracosane
D42 n-
eicosane
Qlon: 66
nC20
57
nC20
nC2\
57
nC2\
nC22
57
nC22
nC23
57
nC23
nC24
57
nC24
nC25
57
nC25
nC26
57
nC26
nC21
57
nC21
D66 n-dotriacontane
surrogate
66
D66-dotriacontane
D62
n-tnacontane
Qion: 66
nC2%
57
«C28
nC29
57
nC29
nC30
57
nC30
nC3\
57
nC3\
nC32
57
nC32
nC33
57
nC33
nC34
57
nC34
nC35
57
nC35
5 -cholestane
surrogate
217
5 -cholestane
5 -antrostane
Qlon:245
Hopane
alkane
191
HoDane
55
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Table 3.4 Target PAHs analyzed by GC/MS
(Purandare, 1999)
Compound Name
Misc.
Info.
Qlon
Response Factors
Reference
Compounds
Internal
Standards
D10 1-methyl naphthalene
surrogate
152
D10 1-methyl
naphthalene
D8-
naphthalene
Qlon: 136
Naphthalene
2-ring PAH
128
Naphthalene
CI naphthalene
2-ring
alkyl PAHs
142
Naphthalene
C2 naphthalene
156
Naphthalene
C3 naphthalene
170
Naphthalene
C4 naphthalene
184
Naphthalene
Fluorene
3 ring PAH
166
Fluorene
D10-
anthracene
Qlon:188
CI fluorenes
3 ring
alkyl PAHs
180
Fluorene
C2 fluorene s
194
Fluorene
C3 fluorene
208
Fluorene
Dibenzothiophene
3 ring PAH
184
Dib enzothi ophene
CI dibenzothiophene
3 ring
alkyl PAHs
198
Dib enzothi ophene
C2 dibenzothiophene
212
Dib enzothi ophene
C3 dibenzothiophene
226
Dib enzothi ophene
D10 phenanthrene
surrogate
188
D10 phenanthrene
Phenanthrene
3-ring PAHs
178
Phenanthrene
Anthracene
178
Anthracene
CI phenanthrenes
3 ring
alkyl PAHs
192
Phenanthrene
C2 phenanthrene s
206
Phenanthrene
C3 phenanthrenes
220
Phenanthrene
C4 phenanthrenes
234
Phenanthrene
N aphthob enzothi ophene
4 ring PAH
234
Dib enzothi ophene
CI naphthobenzothiophene
4 ring
alkyl PAHs
248
Dib enzothi ophene
C2 naphthobenzothiophene
262
Dib enzothi ophene
C3 naphthobenzothiophene
276
Dib enzothi ophene
Fluoranthene
4-ring PAH
202
Fluoranthene
D12-
chrysene
Qlon: 240
D10 pyrene
surrogate
212
D10 pyrene
Pyrene
4-ring PAH
202
Pyrene
CI pyrenes
4 ring
alky PAHs
216
Pyrene
C2 pyrenes
230
Pyrene
56
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Table 3.4. Target PAHs (Contd.)
Compound Name
Misc.
Info.
Qlon
Response Factors
Reference
Compounds
Internal
Standards
Chrysene
4 ring PAH
228
Chrysene
D12-
chrysene
Qlon: 240
CI chrysenes
4 ring
alkyl PAHs
242
Chrysene
C2 chrysenes
256
Chrysene
C3 chrysenes
270
Chrysene
C4 chrysenes
284
Chrysene
benzo(b)fluranthene
5-ring
PAHs
252
Benzo(b)fluoranthene
D12-
perylene
Qlon: 264
benzo(k)fluoranthene
252
B enzo(k)fluoranthene
benzo(e)pyrene
252
Benzo(e)pyrene
benzo(a)pyrene
252
Benzo(a)pyrene
indeno(l, 2, 3-cd)pyrene
6-ring
PAHs
276
Indeno (1, 2, 3-cd)
pyrene
dibenzo(a,h)anthracene
5-ring
PAHs
278
Dibenzo
(a,h)anthracene
benzo(g,h,i)perylene
6-ring
PAHs
276
Benzo(g,h,i)perylene
3.2 Biomarkers
As mentioned in Chapter 2, internal biomarkers have been widely used to distinguish between
biodegradation and the physical or chemical loss of oil from treated plots in bioremediation field
studies (Bragg et al., 1994; Venosa et al., 1996; Lee et al., 1997b). An ideal biomarker should
(1) provide source specific information, (2) not be formed during physical, chemical weathering
and biological processes, (3) be non-biodegradable or relatively resistant to biodegradation on
time scales relevant to the study or cleanup, (4) be extracted from the sample with similar
efficiency to the other associated compounds (Douglas et al, 1994; Prince et al., 1994a,b). The
ideal biomarker would also be subject to the same physical and chemical removal mechanisms as
the target analytes so that any differences could be attributed to biodegradation. Unfortunately,
existing biomarkers can rarely meet all these criteria. An overview of commonly used
biomarkers in monitoring oil bioremediation is given here. Issues to be aware of when using
biomarkers will be discussed.
3.2.1 Commonly used biomarkers
Pristane andphytane
Hydrocarbon degrading microorganisms usually degrade branched alkanes or isoprenoid
compounds such as pristane and phytane at lower rates than straight-chain alkanes. Pristane and
phytane are also subject to the same physical and chemical removal mechanisms as their
57
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corresponding straight-chain alkanes (Wang et al., 1998). Therefore, n-C17:pristane and n-
C18:phytane ratios have been traditionally used to interpret the extent of biodegradation
(Gaudlach E.R et al., 1983; Lee and Levy, 1987,1989,1991). However, it was later found in
Prince William Sound that these compounds were resistant to degradation in the initial time
period, but they degraded rather rapidly over longer periods of time (Prince 1993, Bragg, et al.,
1994). This phenomenon was also observed in a study on bioremediation of Light Arabian crude
oil in a freshwater wetland environment as shown in Figure 3.1 (Purandare, 1999). The study
examined the dry weight normalized pristane, phytane, and hopane concentrations over the 32-
week experimental period. It can be seen that the pristane and phytane showed significant
degradation, but hopane remained constant throughout the course of the study. These results
show that the n-alkane:isoprenoid ratio may be useful indications of biodegradation in very short
term or the earliest part of a study. However, they may substantially underestimate the extent of
biodegradation in a long run.
20
s
16
12
lib
I1
M)
M)
a
o
Figure
Hopanes
Hopanes, derived from the molecular fossils of prokaryotic and eukaryotic membranes (Peters
and Moldowan, 1993), are very resistant to biodegradation as shown in Figure 3.1.
17a(H),21(3(H)-hopane (Figure 3.2) has been found to be neither generated nor biodegraded
during the biodegradation of crude oil on time scales relevant to estimating the cleaning of oil
spills and therefore has appropriate characteristics to serve as an internal standard for monitoring
~ prystane
~ phytane
~ hopane
0 weeks 2 weeks 4 weeks 8 weeks 16 weeks 32 weeks
Weeks of Operation
3.1 Change of pristane, phytane and hopane in a freshwater wetland
58
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biodegradation of both specific petroleum compounds and total oil in crude oil in the
environment (Prince et al., 1994). Hopanes have been viewed as the biomarkers of choice
(Mearns, 1997) since the successful application of 17a(H),21(3(H)-hopane for evaluating oil
bioremediation after the Exxon Valdez spill (Douglas et al., 1994; Prince, et al., 1994a&b). Many
recent studies have also chosen this hopane as the biomarker (Garcia-Blanco et al., 2001b;
Purandare, 1999; Swannell etal., 1999a; Venosa etal., 1996).
Figure 3.2 shows alkane analysis results in soil core samples during the first 21 weeks of the St.
Lawrence River field study using dry weight soil normalization and hopane normalization
(Garcia-Blanco et al., 2001b). Due to the substantial heterogeneity of oil distribution in the
freshwater wetland sediments, large standard deviations (17-72% of the mean concentrations)
were obtained when using dry weight soil normalization. No convincing conclusions were able
to be reached based on these data. However, the standard deviation of oil concentrations in the
sediment samples were much lower (2-15% of the mean oil concentrations) when using hopane
normalization, which enabled evaluation of oil biodegradation with high levels of statistical
confidence.
However, it should be noted that hopanes are also very resistant to those physical and chemical
weathering processes that affect most target alkanes and aromatics (e.g., dissolution,
volatilization, and photooxidation). Therefore, although hopane normalization is very useful in
reducing the variability associated with heterogeneous oil distribution, changes in the hopane-
normalized analyte concentrations alone may not be used to verify that biodegradation is the
primary removal mechanism. Use of hopane normalization to distinguish biodegradation from
physical loss of oil is valid mostly when effects of dissolution and volatilization are negligible
(Prince, 1993; Venosa et al., 1996), which is not always true. Other means of verifying
biodegradation are needed such as the use of alkylated PAH isomers (see next section).
Alkylated PAH isomers
Biodegradation can be verified as a removal mechanism by determining the relative degradation
rates for homologous series of alkylated PAHs. Preferential biodegradation of aliphatic and
aromatic hydrocarbons based on molecular structure has long been recognized (Jobson et al.,
1972; Walker et al., 1976; Roubal and Atlas, 1978; Fedorak and Westlake, 1981; Elmendorf et
al., 1994, Wang et al, 1998). This is particularly true for alkylated PAHs, which are biologically
transformed more slowly as the extent of alkyl substitution increases (Elmendorf et al., 1994;
Venosa et al., 1997). Biodegradation results in unique characteristic changes in the distributions
of homologous series of alkylated PAHs. On the other hand, physical weathering does not cause
the same types of alterations in their relative distributions (Wang et al., 1998). In other words,
physical weathering causes equal losses in all homologues irrespective of the extent of alkyl
substitutions. Recent research has also shown that the relative biodegradation rates for alkylated
homologs of the 2- and 3-ring PAHs were remarkably similar for mixed cultures of hydrocarbon-
degrading bacteria isolated from a wide variety of sources (Venosa et al., 1997), and they were
also very similar in the field, despite the much higher absolute rates that were observed in the
laboratory (Venosa et al., 1996). Therefore, the distribution patterns of alkylated PAHs, when
used in conjunction with other oil analysis data, can be very useful in accurate assessment of the
extent and progress of oil biodegradation.
59
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a) Dry soil weight normalized
1000
800 -
-
-o
W)
*
600 -
I 400
6J3
E
200
-•— Natural Attenuation
Bio + Ammonium
Phyto + Ammonium
-4— Phyto + Nitrate
-A— Unoiled + Ammonium
A A
0 1 2
8 12
Time (weeks)
16
21
b) Hopane normalized
600
500
ca
a
o
.=
6J3
400 -
s 300
e
« 200
6J3
100
0
t o 1
p
1
^—-1
1 I
—•— Natural Attenuation
—~— Bio + Ammonium
Phyto + Ammonium
—+— Phyto + Nitrate
0 1 2
16
21
4 6 8 12
Time (weeks)
Figure 3.2 Comparison of alkanes analysis results in soil core samples from St. Lawrence
field study using (a) dry soil weight normalization and (b) hopane normalization
60
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Other Biomarkers
For refined petroleum products such as diesel fuel and fuel oil #2 that do not contain the hopane,
C4-phenanthrenes/anthracenes may be substituted (Douglas el al., 1994). These compounds are
degraded, but very slowly. Lee et al. (1997b) used C2-chrysenes as biomarkers in their studies
on bioremediation of weathered Venture Condensate because hopane concentration in the oil was
at or near the detection limits of their instruments and C2-chrysenes are also considered highly
degradation resistant. C4-chrysenes can also be used if their concentrations are high enough to be
detectable.
Table 3.5 Biomarkers and their characteristics
Biomarkers
Advantages
Limitations
Applicable to
References
Pristane
Biodegradable
Monitoring
Lee et al.
& Phytane
early stages of
1987&1989
biodegradation
Hopanes
Resistant to
Can not
Reducing
Prince et al.,
biodegradation
distinguish
spatial
1994; Douglas
biodegradation
variability
etal., 1994
from abiotic
weathering
processes
Alkylated
Loss pattern
Verification of
Venosa et al.,
PAHs isomers
characteristic of
biodegradation
1997a; Wang et
biodegradation
al., 1997&1998
Phenanthrenes,
Refined
Douglas et al.,
Anthracenes,
petroleum
1994
and Chrysenes
products
3.2.2 The effect of contaminant redistribution on observed remediation rates
Loss of oil due to physical washout and sand redistribution can be significant in an oiled beach.
Hopane normalization is an effective way to distinguish biodegradation from the effects of the
physical washout and sediment exchange between the inside and outside of experimental plots
when all of the oil is initially present inside of the plots and most of the beach is clean, such as in
a study involving intentional oiling (Venosa et al., 1996). However, it will not work well in a
study involving small plots set up on a beach contaminated by a "spill of opportunity" or a real
oil spill. The reason is that when a study is carried out on a beach completely contaminated by an
oil spill, oil or oiled sand will transport between relatively small treated areas (i.e., experimental
plots) and large untreated but oiled areas. Since hopane is a conservative biomarker, its
concentration will be the same inside and outside of the plots (assuming that the beaches were
61
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uniformly oiled initially). Therefore, oiled sand coming into the plots will have the same amount
of hopane as the sand leaving the plots (assuming that treatment does not result in physical
removal of bulk oil from sand inside the plots), but the concentrations of target analytes will be
higher, because the biodegradation rates will be lower in untreated areas. Since the hopane
concentration inside the plots will not be affected by sand transport, it does not allow us to
quantify the rate of sand exchange between the inside and outside of the plots.
A theoretical analysis conducted by Wrenn et al. (1999) illustrates how physical exchange of
sand between treated and untreated areas of the beach affects the observed biodegradation rate of
target analytes when small plots are set up on a large beach. For the nutrient-treated plots, the
rate of change in the analyte concentrations, Atreat, is given by:
dAtreat _Areat dT
dt T dt treat treat [h dt T dt J treat
For the untreated control plots and in the untreated areas of the beach, the rate of analyte,
Acon, disappearance is:
(32)
Where ktreat and kcon are the first-order biodegradation rate coefficients in the presence and
absence of nutrients; H is the concentration of hopane; and T is the concentration of a
hypothetical nonbiodegradable tracer that is present in the oil inside of the plots but not outside
of the plots (e.g., a hydrophobic fluorescent dye that is added to the oiled sand inside the plots at
the start of treatment). Equation (3.1) describes the rate of change of the analyte concentration
inside the treated plots due to transport of treated oiled sand out of the plots, biodegradation, and
transport of untreated oiled sand into the plots. Assuming the rates of physical loss of treated
oiled sand from inside the plots and loss of oiled sand from the beach to be first-order processes,
the hopane-normalized analyte concentration inside of the plots at any time can be solved as:
f A \ ( A \ ,, , , n
1 ^(k kT ktreat)t
v H
yHoJ
(ktreat ^con ) (^T ^ H
(kj- + k[reat kcon ks)t
kT ktreat
(3.3)
Where kT and kH are the first-order loss coefficients for the nonbiodegradable tracer and hopane,
respectively.
Representative results from this model are shown in Figure 3.3 using parameters obtained from
Delaware field trial (Venosa et al., 1996). It can be seen that the effect of exchange of oiled sand
between the inside of treated plots and untreated beach is to reduce the observed degradation rate
relative to the true rate. Sand exchange has no effect on the observed biodegradation rate in the
control plots, because biodegradation is assumed to occur at the same rates inside and outside of
those plots. The apparent rate of remediation in the treated plots, however, will decrease while
relatively large amounts of oil remain even when bioremediation would be capable of achieving
a complete cleanup if the entire contaminated shoreline were treated. This could lead to the
62
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incorrect conclusion that, whereas bioremediation can stimulate the initial cleanup rate, it cannot
restore the contaminated shorelines to acceptable conditions. The analysis shows that, if
behavior of this type is observed, it is probably an artifact of treating a small fraction of the total
contaminated area, and more complete remediation would be expected when a larger area is
treated. Nevertheless, there are other possible explanations for incomplete remediation. These
include the inability to maintain sufficient nutrients in the bioremediation zone and, especially on
high-energy beaches, a high loss rate of bacteria from the oiled surfaces (e.g., due to scouring by
waves). Therefore, a thorough monitoring program is a very important component of this
research, because data on nutrient concentrations and microbial activity is required to properly
interpret the results if the cleanup goals are not achieved during the field study.
50
control degradation rate
(true and observed)
Ifl
X
•<
-20
"s
o
H
true degradation rate
(treated)
observed degradation rate
treated)
10
0
0
100
Time (days)
Figure 3.3 Reduction in the observed oil biodegradation rate due to exchange of oiled sand
between the inside of treated plots and untreated surrounding beach
3.3 Sampling in the Field
Because oil contaminated sites are highly heterogeneous, representative sampling is difficult but
also extremely important for proper evaluation of bioremediation. Field sampling procedures
must be designed to achieve statistically valid sampling and to minimize contamination or
changes in the samples. Variables that affect the representativeness of samples and their methods
of collection include characteristics of media, concentration distribution of analytes, and bias
introduced during collection, preparation, and transportation (Lee, 2000). General principles to
achieving statistically valid sampling in soil environments and solid waste have been well-
documented (Tan, 1996; USEPA, 1992). However, little information is available regarding
sampling protocols for monitoring oil bioremediation in marine shorelines and freshwater
63
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wetlands. Major considerations to achieve representative sampling are summarized as follows
based on the soil-science literature and field experiences from recent oil bioremediation projects.
3.3.1 Sampling strategies
Types of sampling
Sampling methods can be classified into simple random sampling, systematic sampling, stratified
sampling, and compositing (Tan, 1996). The Simple random sampling or the grab method
involves collecting samples at random in a sampling area. This method depends completely on
the luck of the draw without considering the variation of analytes in the sampling field.
Purandare (1999) found that results of oil analysis in a wetland sediment were highly variable
and unreliable using this sampling approach during the early stage of his study. Therefore, this
approach may only be suitable for use in relatively homogeneous systems. Systematic sampling
involves taking samples based on certain patterns, such as collecting samples in a grid pattern.
This method will ensure that the entire sampling area is represented in the sample. Stratified
sampling involves dividing the sampling field into a number of sectors or quadrants and taking
independent samples in each sector according to the rule of proportionality (e.g., taking more
samples in more heavily oiled sites). These approaches can often provide more accurate results
than simple random sampling, because with this method the samples are distributed more evenly
over the population. Compositing involves the mixing of sampling units to form a single sample,
which has the advantage of increased accuracy through the use of large numbers of sampling
units per sample. This approach in combination of stratified sampling has been frequently used
in recent field studies on oil bioremediation (Venosa et al., 1996; Garcia-Blanco et al., 2001b).
Depth of sampling
Sampling depth in oil spill sites mainly depends on the distribution of the analytes of interest,
especially the depth of oil penetration. Crude oil rarely penetrates coastal sediments to depths of
greater than one foot (Gundlach, 1987). Penetration of oil in wetland environments will be even
less deep than in most marine sediments. Purandare (1999) found that oil only penetrated to 2.5
cm in a wetland sediment in 16 weeks. The top 2 cm layers of the sample cores were then used
for oil analysis. Therefore, a survey of oil penetration in the contaminated site is critical in
determination of sampling procedures for a bioremediation application.
Size of sampling
The size of the sample required depends on the available resources, the required degree of
confidence, and the objectives of the analysis (Rupp and Jones, 1993). Generally, the more
heterogeneous the system, the more intense must be the sampling efforts to reach a given
accuracy. However, economic considerations often restrict both the quantity and the number of
samples taken, and a balance should be obtained between the size of samples to be taken for
required confidence and economic factors. Following expressions are some examples of
statistical approaches, which can be used to calculate the required number of samples with
respect to an acceptable error (Peterson and Calvin, 1986; Tan, 1996):
n = 4o/E2 (3.4)
or n = t2S2/E2 (3.5)
64
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where n is number of samples, t is t-test value, E is acceptable error, S2 is sum of squares of
sample deviation, and o is standard deviation.
Sample handling and storage
Sample handling procedures must be designed to minimize cross-contamination or chemical and
biological changes in the samples. For example, different sets of sampling tools should be used
for different treatments to avoid cross contamination. Clean and proper containers, such as PVC
bags and tin cans, should be used for sample storage. Samples, if not being analyzed in the field,
should be frozen using blue ice or dry ice until they reach their destination. And once they arrive
the destination, samples should be stored in a freezer at -18 to - 20 °C until needed.
3.3.2 Field sampling experiences
Two examples of well-designed sampling protocols used in definitive bioremediation field
studies are shown below (Venosa et al., 1996; Garcia-Blanco et al., 2001b).
Sampling protocols used in the Delaware field bioremediation study (Sandy Beach)
Each plot was divided into 4 quadrants for sampling purposes, and the sand samples were
collected at the nodes of a 0.5 m by 0.5 m grid in each quadrant (Figure 3.4). As Figure 3.4
shows, all sample nodes were at least 0.5 m from the plot boundary on all sides. This buffer
zone was designed to minimize the impact of edge effects on the observed extent of
biodegradation. Since the plots were at least 9.5 m long, this provided a minimum of 28 sample
nodes for each quadrant.
Samples were composited from 2 randomly selected nodes in each quadrant at 1 or 2 week
intervals for a period of 14 weeks, but they were never collected from the same node twice. The
sampling frequency was higher near the beginning of treatment and followed the order of Weeks
0 (i.e., just before treatment began), 1, 2, 3, 4, 6, 8, 10, 12, and 14. With 10 sample events in this
study, 20 sample nodes were required per quadrant. Sand samples were collected with hand
augers to a depth of 15 cm, which was determined based on preliminary oil penetration study.
Sampling protocols used in St. Lawrence River field study (Freshwater Wetland)
A randomized sampling plan was designed to eliminate sampling bias. Each plot was divided
into six sectors, each measuring 1.5 m x 1.0 m, with the 1.5 m dimension parallel to the
shoreline. Each sector was subdivided into 10 subsampling zones, corresponding to the
predetermined sampling events. Each subsampling zone had dimensions 50 cm x 30 cm.
At each sampling event, a 9-cm core sample was collected using a tulip bulb planter from pre-
assigned random subsampling zones from each of the six sectors within each plot. These samples
were combined into 2 composites (3 predetermined samples per composite). Both composite
samples were placed into quart size paint cans, frozen, and shipped to the University of
Cincinnati (OH) for oil analysis. Samples were kept in the freezer (-18 to -20 °C) until they were
extracted. Composite 1 served as the sample to be analyzed by GC/MS. Composite 2 was
analyzed for three of the sampling events to check within-plot variability. The remaining
composite 2 samples were frozen and archived. The first sampling event (week 0) was carried
65
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out at low tide the day after the application of oil and nutrients. Subsequent sampling events
were at weeks 1, 2, 4, 6, 8, 12, 16, 21, 48, and 65.
The sampling designs mentioned above are important because they incorporate replicate plots
with random placement on the experimental plane. These are essential for permitting proper
statistical analysis of treatment effects.
4 m
9.5 m
V
T
a
Figure 3.4 Example of an experimental plot in the Delaware field study showing sampling
quadrants and nodes
3.4 Monitoring General Site Background Conditions
Monitoring general site background conditions is very important for properly evaluating effects
of oil bioremediation. Major background conditions include dissolved oxygen, pore water pH,
temperature, and salinity.
66
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3.4.1 Oxygen
Oxygen availability is crucial for rapid bioremediation because hydrocarbon biodegradation is
primarily an aerobic process. Therefore, the dissolved oxygen (DO) of pore water should be
monitored on a regular basis. Water samples from the oil-contaminated region of the subsurface
can be collected through the multi-port sample wells and sealed in DO bottles. Conventional
methods for DO measurement include iodometric procedures and membrane electrode method
(Eaton et al., 1995).
The iodometric technique is the most precise and reliable titrimetric procedure for DO analysis.
Detailed procedure of this method is described in Standard Methods No. 4500-0 B (Eaton et al.,
1995). Various iodometric modifications have been developed. One procedure that is suitable for
use in the field involves using Hach® high range dissolved oxygen ampoules (Hach Company,
Loveland, CO). Once the Hach® ampoules are filled and capped and the reaction of the reagents
with DO in the water sample is complete, the color is stable indefinitely. Therefore, the capped
ampoules can be transported back to the laboratory trailer where the DO concentration in the
water samples can be determined with a Hach® colorimeter. (Note: the caps often leak gaseous
02 into the samples; so, the samples aren't really stable indefinitely, even though the color is
stable as long as additional O2 can be excluded.)
The membrane electrode method is more suitable for regular field monitoring and in situ DO
determination. It is also recommended for DO analysis in highly polluted waters and colored
waters. The general procedure for this method is described in Standard Methods No. 4500-OG
(Eaton et al., 1995). Detailed analytical procedures may vary depending on the manufacturers of
the DO probes.
3.4.2 pH
Biodegradation of oil is affected by the background pH (Atlas and Bartha, 1992). Oil
biodegradation can also be severely inhibited by dramatic reductions in pH when ammonia is
provided as the nitrogen source (Wrenn et al., 1994). The latter is true for closed environments
such a laboratory flasks where no dilution is possible from continuously changing aqueous
conditions such as tides. Nonetheless, monitoring pH in the field is of particular importance in
evaluating the effect of oil biodegradation.
The pH values of pore-water samples are normally measured using a portable pH meter with a
combination electrode. The pH can be measured in the field either immediately after the samples
are collected or by putting pH electrodes directly into water in sampling wells. Sediment pH can
also be measured in the field by mixing the soil samples with reagent water according to EPA
method 9045c (US EPA, 1992).
3.4.3 Temperature
As discussed in Chapter 2, temperature affects both the properties of spilled oil and the
biodegradation processes. All the other measurements are also temperature-dependent.
Temperature profiles in air, water, and sediment should be monitored regularly using appropriate
67
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thermometers. Many instruments for other analyses, such as DO or pH, have built-in
thermometers, and the temperature should be recorded along with the other measurements. Any
temperature measurement devices should be calibrated with a National Institute of Standards and
Technology (NIST)-certified thermometer before field use (Eaton el al., 1995).
3.4.4 Salinity
Salinity of the environment may be an important factor in oil bioremediation, particularly in
estuarine environments or in marine shorelines where regional seaward groundwater flow exists
(Boufadel, et al., 1999b; Zobell, 1973).
Salinity of pore water can be measured by either a conductivity method or density methods (e.g.,
Standard Methods No. 2520B or 2520 C; Eaton et al., 1995). The conductivity method is the
most commonly used method for salinity analysis due to its high sensitivity and ease of
measurement. Conductivity meters can be installed in the field to monitor salinity profiles in
marine shorelines. The density method involves using a precise vibrating flow densitometer.
Water salinity can also be determined rapidly in the field using this method.
3.5 Monitoring of Biological Impacts
The public has responded favourably to bioremediation since its implicit goal is that of reducing
toxic effects by converting organic molecules to cell biomass and other benign materials such as
carbon dioxide and water (Atlas and Cerniglia, 1995). However, concerns about the net benefit
of bioremediation strategies remain. This is attributed to lingering questions regarding the
potential production of toxic metabolic by-products, possible toxic components in the
formulation of bioremediation agents, and the ineffective degradation of the most toxic
components of residual oils (Hoff, 1991; Office of Technology Assessment, 1991).
To date, a single ideal biological method - both sensitive and efficient - for the assessment of
contaminant impacts to all sediment biota has not been identified. Two separate, yet
complimentary, approaches have evolved: bioassessment and bioassays. Bioassessments are
field-based analyses typically characterized by assessing the impacts of the contamination and
treatment activity on environmental populations such as benthic communities, intertidal flora and
fauna, etc. They are characterized as having limited experimental controls (Herricks and
Schaeffer, 1984). Bioassays are laboratory-based tests that incorporate rigorous experimental
protocols and controls. Both toxicity tests and bioaccumulation studies are bioassays (Chapman,
1989).
3.5.1 Bioassessment
Changes in benthic community structure can be used as a means of assessing ecosystem response
to contaminated sediments in aquatic ecosystems. Since most contaminants such as crude oil
within the aquatic ecosystem eventually bind to sediment particles, emphasis on benthic
organisms (bottom dwelling vertebrates and invertebrates) as a primary means of assessing
ecosystem response is warranted. Of particular importance are the macrobenthic invertebrates
(organisms retained on screens of mesh size >0.2 mm) because of their basic longevity,
68
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sedentary lifestyles, proximity to sediments, influence on sedimentary processes, and trophic
importance. Microinvertebrates such as rotifers and nematodes are of particular ecological
interest; however, their taxonomy is less well known. Hence, they have not been routinely
monitored in environmental assessments.
While there is a vast bioassessment database on the effects of oil spills, the effects of clean up
techniques have until recently been seldom addressed. Clearly, a database on the effects of
clean-up operations would have obvious potential for guidance. For example, in a follow up of
the Exxon Valdez oil spill clean up, Driskell el al. (1996) noted that total abundance, species
richness, species diversity, and abundance of several major taxa (polychaetes, bivalves, and
gastropods) were significantly lower in hot-water-washed beaches than in unoiled beaches.
Infauna at oiled sites that were not hot-water washed rebounded quickly following the
disturbance. Three years after the spill, recovery of infauna at sites that were cleaned still lagged
significantly behind the oiled sites. Principal component analysis (PCA), a multivariate
ordination technique, was used to track site recovery trends. Negative effects were indicated by
reduction in size or biomass, mortality, and reduced or failed reproductive success. Conversely,
the possibility of positive impacts was also identified (e.g., when oil tolerant species bloom
during the period of reduced competition-predation). Changes in epifauna and infauna were also
used to assess the rates of natural recovery and the impacts of intertidal clean-up activities on the
coast of Saudi Arabia following the 1991 Gulf oil spill (Watt et al., 1993).
Macroinvertebrate bioassessment has been limited in field trials evaluating the efficacy of
bioremediation strategies due to the amount of unrestricted surface area required for sample
collection. The use of bioassessments in this context will expand with the development of
bioassay protocols based on bioanalytical techniques (enyzmatic measurements, as well as
immunoassay and biosensor techniques) aimed at the subcellular or multicellular level of
biological organization. Application of these kinds of tests should be tailored for both the field
and laboratory (Lee et al., 1998).
Bioassessment can readily include potential impacts on vegetation. Field surveys demonstrated
that the 1991 Gulf War oil spill severely damaged intertidal vegetation along the Saudi Arabian
Gulf coast (Boer, 1994). Along a 45 km stretch of intertidal mangroves and salt-marshes,
Salicornia europaea was almost extinct. Dwarf mangrove (Avicennia marina) and salt-marshes
dominated by Arthrocnemum macrostachyum and Halocnemum strobilaceum were severely
damaged. Halopeplis perfoliata and Limonium axillare salt-marshes were relatively unaffected.
It was noted that natural re-establishment of the vegetation would be protracted unless active
measures were taken to aid recovery.
A bioassessment of vegetative growth was recently used to document the efficacy of
bioremediation strategies to enhance the rate of habitat recovery within a tidal freshwater marsh
located along the St. Lawrence River, Canada (Lee et al., 2001). Scirpuspungens, the dominant
plant species at the study site was found to be tolerant to the oil, and its growth was significantly
enhanced above that of the unoiled control by the addition of nutrients (Figure 3.5).
The aim of oil spill remediation is to restore a site to its pre-spill condition. In this context,
monitoring the recolonization of impacted areas should be a primary goal in bioassessments.
69
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Colonization and succession describe changes in the numbers and kinds of organisms making up
the community over time. They provide an integrated measure of a toxicant's effect on
immigration, emigration, competition and predation. Colonization is somewhat analogous to
reproduction in a single species: it reflects the ability of the community to replicate and organize
itself. Fleeger et al. (1996) showed that unweathered Exxon Valdez crude oil delayed, but did
not preclude, colonization by meiofauna (harpacticoids) into azoic sediment of Prince William
Sound.
To date, emphasis has been placed on the characterization of impacts on the macroinvertebrate
community and vegetation. Nevertheless, sediment-associated contaminants enter the non-
benthic environment and community through natural processes including resuspension,
desorption, ingestion of benthic organisms, ingestion of sediment, and adsorption to or uptake
through membranes during sediment contact. Due to mobility and sampling issues, it is
inherently much more difficult to work on pelagic organisms such as fish. Nevertheless, given
the holistic nature of toxicant perturbations on aquatic ecosystems and the multifaceted
interactions between the water and sediment compartments, consideration should also be given
to the bioassessment of fish and other nonbenthic community organisms (e. g., bacteria,
phytoplankton, cladocera, and amphibians).
140 a ¦ Max. ~ Min.
120
100
ao
60
?
— 40
'5 20
X
Ss Q
jP
Q.
60- b
till
40
20
ll
1 II 1
Nat- NH/ NOj NH^
Att. Intact Intact UnoiEed
Figure 3.5 Minimum and maximum height of the (a) dominant (Scirpus pimgens) and (b)
secondary (Eleocharis palustris) plant species at Weekl5. Treatment of the oiled plots included
natural attenuation (Nat. Att.); nutrient amendment with granular ammonium nitrate and super
triple phosphate (NH4+ Intact); nutrient amendment with sodium nitrate instead of ammonium
nitrate (NO3"). Unoiled plots were amended with granular ammonium nitrate and super triple
phosphate (NH4+ Unoiled).
70
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3.5.2 Bioassays
Bioassays provide a more accurate picture of ecosystem health at a contaminated site than
chemical analyses because their result is an integration of the interaction that occurs between the
contaminant and environmental variables. Bioassay endpoints are quantitative measures of
toxicity. They compliment biological surveys, which describe communities of organisms in the
field, and chemical analyses, which provide information on the nature of contaminants at a site,
the magnitude of the remediation problem, and potential methods of treatment. Resource
managers frequently use bioassays to identify the most toxic areas, thereby helping to prioritize
sites for more thorough evaluation, including the selection of methods for chemical analyses.
Sediment toxicity tests are generally classified as "acute" or "chronic". They are usually
performed on whole sediment (e.g., solid-phase), suspended sediment, sediment liquid phases
(pore water, interstitial water), or sediment extracts (elutriates, solvent extracts). In general,
assays using whole sediment samples are more sensitive than assays using elutriate or pore water
samples. The American Society of Testing and Materials (ASTM, 1991) currently defines an
acute toxicity test as a comparative study in which the organisms that are subject to different
treatments are observed for a short period, usually not constituting a substantial portion of their
life span. A chronic test is defined as a comparative study in which organisms that are subjected
to different treatments are observed for a long period or a substantial portion of their life span.
Acute tests often utilize mortality as the only measure of effect, while chronic tests usually
include measures of growth, morphology, reproduction, behavioral effects, or other sublethal
endpoints.
Plant and animal communities are diverse; their members differ in their sensitivity to toxicants.
A single species bioassay cannot represent the range of sensitivity of all biota within an
ecosystem. To improve ecological relevance, a test battery approach with species from different
trophic levels is required. Accountability for the influence of natural environmental factors in
sediment bioassays is assisted by the testing of reference and control samples. Reference
sediment may be defined as sediment collected from the vicinity of a study site, possessing
similar characteristics to the test sediment, but without anthropogenic contaminants. Sediment
characteristics, such as particle size distribution and percent organic carbon of the reference
sediment should simulate, as closely as possible, that of the test sediment. In some cases the
reference sediment might also show toxicity due to naturally occurring chemical, physical, or
biological properties. This factor can be addressed by determining the toxicity of control
sediments (natural or artificially prepared sediments known to be nontoxic) and the use of
positive controls (a sediment of known toxicity to the test organism under the conditions of the
test).
Bioassays have been developed and used extensively since the 1960s for the screening of
chemicals and regulatory compliance monitoring. Sediment bioassays have been used
extensively to diagnose the effects of oil spills (Teal, et al., 1992; Gilfillan et al., 1995; Neff and
Stubblefield, 1995; Randolph et al., 1998). Their application has now been extended to include
the documentation of effects and success of oil spill countermeasures like bioremediation (Lee et
al., 1995b; Mearns et al., 1995). While any living organisms can be used in theory, toxicity tests
with fish and macroinvertebrates have been standardized by environmental agencies to assess the
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hazards of industrial wastes to aquatic systems (Blaise et al., 1988). Rapid advances are now
being made in the development of cost-effective high-performance micro-scale procedures
involving bacterial, protozoan, microalgal, and microinvertebrate indicators (Wells et al., 1998).
Furthermore, the high demand for simple, rapid, and practical toxicological procedures has
resulted in the creation of commercial bioanalytical products such as the Microtox® Test (AZUR
Environmental Inc., USA).
Major criteria to consider in the selection of species for sediment toxicity testing include: (1)
their behavior in sediment (habitat, feeding habits, etc.), (2) their sensitivity to test material, (3)
their ecological and/or economic relevance, (4) their geographical distribution, (5) their
taxonomic relation to indigenous animals, (6) their acceptability for use in toxicity measurement
(standardized test method), (7) their availability, and (8) their tolerance to natural sediment
characteristics such as grain size.
The response of the test organisms to the toxicant or test sediment is often affected by its life
stage. Larval or juvenile life stages are generally more sensitive than adults.
3.5.2.1 Benthic invertebrates
In terms of benthic invertebrates, amphipods are among the most sensitive of benthic species.
They are among the first to disappear from benthic communities in sediments impacted by
pollution (Swartz et al., 1982; Mearns and Word, 1982). They have been used successfully to
characterize shoreline impacts following oil spill incidents (Teal et al. 1992; Gilfillan et al.,
1995; Wolfe etal., 1996).
Gilfillan et al. (1995) collected mussels from several locations for tissue hydrocarbon analysis to
estimate bioavailable hydrocarbon concentrations in epifaunal species. In a newer approach,
studies of sediment contamination and verification of laboratory bioassays involved controlled in
situ exposures (caged animals) to expand the level of ecological relevance. In this case, oysters
were used during a shoreline bioremediation experiment in Delaware Bay to document the loss
of oil from the study area and to determine how the overall oiling may have impacted offshore
resources (Mearns et al., 1997).
3.5.2.2 Microtox
Simple, sensitive, rapid, cost-effective, reproducible, and practical methods are needed for the
screening of toxic impacts during oil spill response operations. The Microtox® Test, a
commercial bioassay accepted by regulatory agencies, is based on the measurement of changes
in light emission by a nonpathogenic, bioluminescent marine bacterium (Vibrio fisheri) upon
exposure to test samples. The test has been used extensively worldwide over the last 18 years for
toxicity screening of chemicals, effluents, water and sediment, and for contamination surveys
and environmental risk assessment. Variations of this test have been applied to time-series
monitoring of sediment and water toxicity. Ho and Quinn (1993) identified strong rank
correlations between the Microtox response and polycyclic aromatic fractions of organic extracts
of sediments. Its application for monitoring the efficacy of oil spill remediation methods has
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been proven (Lee et al., 1995b; 1997b). Mueller et al. (1999) quantified the effectiveness of
intrinsic recovery within an oiled wetland by monitoring the rate of acute toxicity reduction
using the Microtox 100% Test on a water extractable phase. The observed decrease in toxicity
followed a pattern similar to the decrease seen in petroleum concentrations by GC/MS total
target analyte measurements.
3.5.2.3 Fish
Due to their economic, recreational, and aesthetic value, fish have been historically selected as a
primary bioassay organism. Difficulties in using fish as biomonitors of sediment contamination
arise from their preference for particular sediments or habitats and their residence time in or over
contaminated areas. Furthermore, their absence in a water body may more directly reflect water
quality.
Biochemical and physiological alterations, if severe enough or protracted, can lead to structural
alterations in organelles, cells, tissue, and organs. Detection of specific alterations using
anatomical and cytological endpoints may indicate both prior and current exposure to chemical
contaminants, so histopathology has been instrumental in assessing the toxicological impact of
contaminated sediments. The documentation of neoplasms in fish and other aquatic organisms
was perhaps the first use of histopathological indices in ecotoxicology.
Biomarkers (as distinguished from the oil biomarkers such as hopane discussed earlier) are used
by resource managers as a means to identify a toxicological response from fish populations.
Biomarkers can be defined as biochemical, physiological, or pathological responses measured in
individual organisms on exposure to environmental contaminants, and which also provide
information concerning sublethal effects arising from such exposures. The family of enzymes
referred to as cytochrome P-450s (or P450) act on the functional groups of lipophilic substrates
in a process referred to as mixed function oxidase (MFO) or monooxygenase reactions (Ortiz de
Montellano, 1986). MFO reactions are induced by polycyclic aromatic hydrocarbons (PAHs)
and a variety of halogenated hydrocarbons (notably certain chlorinated biphenyls, dibenzofurans,
and dibenzodioxins). The enzyme system is sensitive to these contaminants at levels
encountered in the environment. In fish, the most widely employed and readily performed
techniques are measurements of enzyme activities, particularly aryl hydrocarbon hydroxylase
(AAH) and ethoxyresorufin O-deethylase (EROD), that are highly associated with the P450 1A
proteins.
Hodson et al. (2001) has monitored changes in the bioavailability and toxicity of oil-derived
PAH to early life stages of fish in a field trial to evaluate the effectiveness of wetland
bioremediation and phytoremediation strategies. For over 1.5 years, sediments from
experimental plots were tested by bioassays of MFO (CYP1A) enzyme activity in livers of trout
as an index of PAH exposure. Oil alone, oil mixed with sediments in the lab, and oiled
sediments from the experimental plots all caused CYP1A induction relative to unoiled controls,
indicating the presence and bioavailability of PAH. Induction did not vary markedly among
treatments, but declined slowly with time. Concomitant chemical analysis suggested that PAHs
were depleted primarily by weathering or sediment dispersion rather than by bioremediation
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treatments. Sediments were also chronically toxic to developing stages of trout and medaka
(iOryzias latipes), causing increased rates of deformities and mortality.
3.5.3 Application of bioassays to assess bioremediation in marine environments
Bioassays were used to document the effectiveness of shoreline bioremediation in accelerating
toxicity reduction of a sandy shoreline at Fowler Beach, Delaware, USA, that had been oiled
with weathered Nigerian Bonny light crude (Mearns et al., 1995). The bioassay suite included
two solid phase (Amphipod Survival, Microtox Solid-Phase) and three pore water tests (Grass
Shrimp Embryo Survival/Growth, Microtox, Sea Urchin Fertilization). Treatment with nutrients
(sodium nitrate and sodium tripolyphosphate) or nutrients and oil degrading bacteria (isolated
from the study site) did not accelerate toxicity reduction. However, results of the high-frequency
test based on the hatching success of grass shrimp embryo suggested there may have been a
substantial delay in pore water toxicity reduction due to the addition of the nutrients themselves
during the first few weeks. The Sea Urchin Fertilization Test was least sensitive. The most
sensitive tests were the 10-day amphipod and grass shrimp embryo bioassays.
Bioremediation by nutrient enrichment was investigated as a method of treating a mixture of
Forties Crude Oil and Heavy Crude Oil stranded on Bullwell Bay, Milford Haven, UK, after the
grounding of the Sea Empress in 1996 (Swannell et al., 1999b). Experimental results showed
that the oil was significantly more biodegraded after two months as a result of application of the
fertilizer. Based on the results of a bioassay that involved monitoring the development of oyster
embryos, and the results of the Microtox Organic Solvent Basic Test, there was no evidence of
detrimental effects associated with the bioremediation treatments.
To date, detrimental effects from nutrient enrichment have not been observed following actual
field operations (Prince, 1993; Mearns et al., 1997) although the possibility of a future incident
still exists. As an example, oxygen depletion and production of ammonia from excessive
applications of a fish-bone meal fertilizer during one field experiment caused detrimental effects
that included toxicity and the suppression of oil degradation rates (Lee et al., 1995b).
Furthermore, in a subsequent bioremediation field trial it was reported that a commercial
bioremediation product suppressed the rates of toxicity reduction as it increased the retention of
residual oil within the sediments (Lee et al., 1997b). For safety assurance, future operational
guidelines must include ecotoxicological monitoring protocols.
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Chapter 4 TYPES OF AMENDMENTS AND CONSIDERATIONS IN THEIR
APPLICATION
The success of oil spill bioremediation depends on our ability to optimize various physical,
chemical and biological conditions in the contaminated environment. Existing amendments for
enhancing oil biodegradation in marine shorelines and freshwater wetlands include addition of
nutrients, addition of microbial cultures or enzymes, phytoremediation, and oxygen
enhancement.
4.1 Nutrient Amendment
As reviewed in Chapter 2, nutrient addition has been proven to be an effective strategy to
enhance oil biodegradation in various marine shorelines. Theoretically, approximately 150 mg of
nitrogen and 30 mg phosphorus are consumed in the conversion of 1 g of hydrocarbon to cell
material (Rosenberg and Ron, 1996). Therefore, a commonly used strategy has been to add
nutrients at concentrations that approaches a stoichiometric ratio of C:N:P of 100:5:1. Recently,
the potential application of resource-ratio theory in hydrocarbon biodegradation was discussed
(Head and Swannell, 1999; Smith et al., 1998). This theory suggests that manipulating the N:P
ratio may result in the enrichment of different microbial populations, and the optimal N:P ratio
can be different for degradation of different compounds (such as hydrocarbons mixed in with
other biogenic compounds in soil). However, the practical use of these ratio-based theories
remains a challenge. Particularly, in marine shorelines, maintaining a certain nutrient ratio is
impossible because of the dynamic washout of nutrients resulting from the action of tides and
waves. A more practical approach is to maintain the concentrations of the limiting nutrient or
nutrients within the pore water at an optimal range (Bragg et al., 1994; Venosa el al., 1996).
Commonly used nutrients include water-soluble nutrients, solid slow-release nutrients, and
oleophilic fertilizers. Each type of nutrient has its advantages and limitations. General
characteristics of these nutrients and important factors affecting their persistence in the field,
such as waves and tides, and physical intrusion effects, will be discussed in this section and
summarized in Table 4.1. More practical issues such as nutrient application strategies will be
discussed in Chapter 5.
4.1.1 Water-soluble nutrients
Commonly used water-soluble nutrient products include mineral nutrient salts (e.g. KNO3,
NaNC>3, NH3NO3, K2HPO4, MgNHiPO/t), and many commercial inorganic fertilizers (e.g. the
23:2 N:P garden fertilizer used in Exxon Valdez case). They are usually applied in the field
through the spraying of nutrient solutions or spreading of dry granules. This approach has been
effective in enhancing oil biodegradation in many field trials (Swannell et al., 1996; Venosa et
al., 1996). Compared to other types of nutrients, water-soluble nutrients are more readily
available and easier to manipulate to maintain target nutrient concentrations in interstitial pore
water. Another advantage of this type of nutrient over organic fertilizers is that the use of
inorganic nutrients eliminates the possible competition of carbon sources. The field study by Lee
et al. (1995a) indicated that although organic fertilizers had a greater effect on total heterotrophic
microbial growth and activity, the inorganic nutrients were much more effective in stimulating
crude oil degradation.
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However, water-soluble nutrients also have several potential disadvantages. First, they are more
likely to be washed away by the actions of tides and waves because of their water-solubility. The
field study in Maine demonstrated that water-soluble nutrients can be washed out within a single
tidal cycle in high-energy beaches (Wrenn, 2000, see section 2.6.2). Second, inorganic nutrients,
ammonia in particular, should be added carefully to avoid reaching toxic levels. Existing field
trials, however, have not observed acute toxicity to sensitive species resulting from the addition
of excess water-soluble nutrients (Mearns el al., 1997; Prince et al., 1994). Third, water-soluble
nutrients may have to be added more frequently than slow release nutrients or organic nutrients,
resulting in more labor-intensive, costly, and physical intrusive applications.
4.1.2 Granular nutrients (slow-release)
Many attempts have been made to design nutrient delivery systems that overcome the washout
problems characteristic of intertidal environments (Prince, 1993). Use of slow release fertilizers
is one of the approaches used to provide continuous sources of nutrients to oil contaminated
areas. Slow release fertilizers are normally in solid forms that consist of inorganic nutrients
coated with hydrophobic materials like paraffin or vegetable oils. This approach may also cost
less than adding water-soluble nutrients due to less frequent applications. Slow release fertilizers
have shown some promises from oil bioremediation studies and applications. For example,
Olivieri et a/.(1976) found that the biodegradation of a crude oil was considerably enhanced by
addition of a paraffin coated MgNH4P04. Another slow-release fertilizer, Customblen (vegetable
oil coated calcium phosphate, ammonium phosphate, and ammonium nitrate), performed well on
some of the shorelines of Prince William Sound, particularly in combination with an oleophilic
fertilizer (Atlas, 1995a; Pritchard et al., 1992; Swannell et al., 1996). Lee et al. (1993) also
showed that oil biodegradation rates increased with the use of a slow release fertilizer (sulfur-
coated urea) compared to water-soluble fertilizers.
However, the major challenge for this technology is control of the release rates so that optimal
nutrient concentrations can be maintained in the pore water over long time periods. For example,
if the nutrients are released too quickly, they will be subject to rapid washout and will not act as
a long-term source. On the other hand, if they are released too slowly, the concentration will
never build up to a level that is sufficient to support rapid biodegradation rates, and the resulting
stimulation will be less effective than it could be. The field trials on of the shorelines of Prince
William Sound showed that on certain beaches, Customblen granules were apparently washed
away before any significant enhancement of bioremediation was recorded (,Swannell et al.,
1996). Several recent studies have shown that a slow release nutrient (Max Bac, a product
similar to Customblen) failed to demonstrate enhancement of oil degradation because the
nutrient release rate was too slow to affect oil biodegradation (Croft et al., 1995; Sveum and
Ramstad, 1995).
4.1.3 Oleophilic nutrients
Another approach to overcome the problem of water-soluble nutrients being rapidly washed out
was to utilize oleophilic organic nutrients (Atlas and Bartha, 1973; Ladousse and Tramier, 1991).
The rationale for this strategy is that oil biodegradation mainly occurred at the oil-water
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interface; since oleophilic fertilizers are able to adhere to oil and provide nutrients at the oil-
water interface, enhanced biodegradation should result without the need to increase nutrient
concentrations in the bulk pore water. A well-known oleophilic fertilizer is Inipol EAP 22, a
microemulsion containing urea as a nitrogen source, lauryl phosphate (the phosphorus source), 2-
butoxy-l-ethanol as a surfactant, and oleic acid to give the material its hydrophobicity. This
fertilizer has been subjected to extensive studies under various shoreline conditions and was
successfully used in oil bioremediation on of the shorelines of Prince William Sound. Other
oleophilic organic fertilizers include polymerized urea and formaldehyde, and some organic
fertilizers derived from natural products such as fishmeal (Lee et al., 1995a; Rosenberg el al.,
1992; Sveum andRamstad, 1995).
Table 4.1 Major nutrient types used in oil bioremediation
Type of nutrients
Advantages
Disadvantages
Applications in the
field or field trials
Water soluble
Readily available
Easy to manipulate
for target nutrient
concentrations
No complicated effect
of organic matter
Rapidly washed out
by wave and tide
Labor-intensive, and
physical intrusive
applications
Potential toxic effect
Alaska (Pritchard et
al., 1992)
Delaware (Venosa et
al., 1996)
Slow release
Provide continuous
sources of nutrients
and may be more cost
effective than other
types of nutrients
Maintaining optimal
nutrient release rates
could be a challenge
Alaska (Pritchard et
al., 1992)
Nova Scotia (Lee et
al., 1993)
Oleophilic
Able to adhere to oil
and provide nutrients
at the oil-water
interface
Expensive
Effectiveness is
variable
Containing organic
carbon, which may
compete with oil
degradation and result
in undesirable anoxic
conditions
Alaska (Pritchard et
al., 1992)
Nova Scotia (Lee et
al., 1987, 1989,1995a
&b)
The effectiveness of oleophilic fertilizers also depends on the characteristics of the contaminated
environment such as action of wave and tide, and sediment types. Based on several earlier
studies, Sveum et al. (1994) indicated that oleophilic fertilizers proved to be more effective than
water-soluble fertilizers when the spilled oil resided in the intertidal zone. But they have no
advantages in enhancing oil biodegradation in the supralittoral zone where water transport is
limited. Inipol EAP 22 was found to be more effective in coarse sediments than in fine sediments
due to the difficulty in penetration for the oleophilic fertilizer in fine sediments (Sveum and
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Ladousse,1989). Variable results have also been produced regarding the persistence of oleophilic
fertilizers. Some studies showed that Inipol EAP 22 can persist in a sandy beach for a long time
under simulated tide and wave actions (Santas and Santas, 2000; Swannell et al. 1995). Others
found that Inipol EAP22 was rapidly washed out before becoming available to hydrocarbon-
degrading bacteria (Lee and Levy, 1987; Safferman, 1991). Another disadvantage with
oleophilic fertilizers is that they contain organic carbon which may be biodegraded by
microorganisms in preference to petroleum hydrocarbons (Lee et al., 1995a; Swannell et al.,
1996), and may also result in undesirable anoxic conditions (Lee et al., 1995b; Sveum and
Ramstad, 1995).
In summary, the effectiveness of these various types of nutrients will depend on the
characteristics of the contaminated environment. Slow-release fertilizers may be ideal nutrient
sources if the nutrient release rates can be well controlled. Water-soluble fertilizers are likely
more cost-effective in low-energy and fine-grained shorelines where water transport is limited.
And oleophilic fertilizers may be more suitable for use in high-energy and coarse-grained
beaches. However, successful application of bioremediation products will always require
appropriate testing and evaluation based on the specific conditions of each contaminated site.
4.2 Microbial Amendments
Addition of oil-degrading microorganisms or bioaugmentation has been proposed as a
bioremediation strategy. The rationale for this approach includes the contention that indigenous
microbial populations may not be capable of degrading the wide range of substrates that are
present in complex mixtures such as petroleum and that seeding may reduce the lag period
before bioremediation begins (Forsyth et al., 1995; Leahy and Colwell, 1990). For this approach
to be successful in the field, the seed microorganisms must be able to degrade most petroleum
components, maintain genetic stability and viability during storage, survive in foreign and hostile
environments, effectively compete with indigenous microorganisms, and move through the pores
of the sediment to the contaminants (Atlas, 1977; Goldstein etal., 1985).
There are many vendors of bioremediation products, who claim their product (most of them are
microbial agents) aids the oil biodegradation process. The U.S. EPA has compiled a list of
bioremediation agents (USEPA, 2000) as part of the National Oil and Hazardous Substances
Pollution Contingency Plan (NCP) Product Schedule, which is required by the Clean Water Act,
the Oil Pollution Act of 1990, and the National Contingency Plan. A current list of
bioremediation agents in NCP schedule is shown in Table 4.2. A product can be listed only when
its safety and effectiveness have been demonstrated under the conditions of a test protocol
developed by EPA (NETAC, 1993). However, listing does not mean that the product is
recommended or certified for use on an oil spill (USEPA, 2000). The efficacy test protocol uses
laboratory shake flasks to compare the degradation of artificially-weathered crude oil in natural
seawater with and without a bioremediation product. Similar test protocols for freshwater
conditions were recently proposed (Haines et al., 1999).
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Table 4.2 Bioremediation agents in NCP product schedule (Adapted from USEPA, 2000)
Type
Name or Trademark
Manufacture
Biological additives
(Microbial Culture or
Enzyme additives)
BET BIOPETRO
BIOGEE HC
BioEnviro Tech, Tomball, TX
RMC Bioremediation, Shreveport, LA
BR (formerly ENVIRO-
ZYME BR)
Enviro-Zyme, Inc., Stormville, NY
ENZYT
(LIQUID/CRY S T A)
Acorn Biotechnical Corporation
Houston, TX
MICRO-BLAZE
Verde Environmental, Inc., Houston, TX
OPPENHEIMER
FORMULA
Oppenheimer Biotechnology, Inc.
Austin, TX
PRISTINE SEA II
Marine Systems, Baton Rouge, LA
PRP ( Petroleum
Remediation Product)
Petrol Rem, Inc., Pittsburgh, PA
STEP ONE
B & S Research, Inc.
Embarrass, MN
SYSTEM E.T. 20
Quantum Environmental Technologies, Inc.
(QET), La Jolla, CA
WMI-2000
Waste Microbes, Inc., Houston, TX
Nutrient additives
INIPOL EAP 22
(oleophilic)
Societe, CECA S.A.
France
LAND AND SEA
RESTORATION
Land and Sea Restoration LLC, San
Antonio, TX
OIL SPILL EATER II
Oil Spill Eater International, Corporation
Dallas, TX
VB 5 91 ™ W ATER,
VB997™SOIL, AND
BINUTRIX
(partially encapsulated
& oleophilic)
BioNutraTech, Inc., Houston, TX
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As reviewed in Chapter 2, however, even though the addition of microorganisms may be able to
enhance oil biodegradation in the laboratory, its effectiveness has not been convincingly
demonstrated in the field. Actually, most field studies indicated that bioaugmentation is not
effective in enhancing oil biodegradation in marine shorelines, and nutrient addition or
biostimulation alone had a greater effect on oil biodegradation than the microbial seeding
(Jobson et al., 1974; Lee and Levy, 1987; Lee et al., 1997b, Venosa et al., 1996). The failure of
bioaugmentation in the field may be attributed to the fact that the carrying capacity of most
environments is likely determined by factors that are not affected by an exogenous source of
microorganisms (such as predation by protozoans, the oil surface area, or scouring of attached
biomass by wave activity), and that added bacteria seem to compete poorly with the indigenous
population (Tagger et al., 1983; Lee and Levy, 1989; Venosa et al., 1992). Therefore, it is
unlikely that externally added microorganisms will persist in a contaminated beach even when
they are added in high numbers. In short, those criteria mentioned above for a successful
colonization are very difficult to be met in the field.
Fortunately, oil-degrading microorganisms are ubiquitous in the environment, and they can
increase by many orders of magnitude after being exposed to crude oil (Atlas, 1981; Lee and
Levy, 1987, Pritchard and Costa, 1991). Therefore, in most environments, there is usually no
need to add hydrocarbon degraders. In certain circumstances that have not been well defined,
when the indigenous bacteria are incapable of degrading one or more important contaminants,
addition of microbial inocula may be considered. Genetically engineered organisms are not
likely to be used in the near or even distant future.
4.3 Plant Amendments (phytoremediation)
Phytoremediation has been defined as the use of green plants and their associated
microorganisms to degrade, contain, or render harmless environmental contaminants
(Cunningham et al., 1996). This technique is emerging as a potentially cost-effective option for
clean-up of soils contaminated with petroleum hydrocarbons (Frick et al., 1999). As
summarized by Macek et al. (2000), the main advantages of phytoremediation include less
disruption to the environment, potential to treat a diverse range of contaminants, and high
probability of public acceptance. Major concerns regarding this technology include dissolution
and migration of contaminants, limitation by the toxicity of the contaminated environments, and
it being a relatively slow process. Phytoremediation has been studied in a freshwater
environment in Quebec, Canada (Garcia-Blanco et al., 2000; Venosa et al., 2002 (submitted)).
These researchers found that addition of nutrients did not result in enhancement of
biodegradation of crude oil contaminating the plots, whether or not plants were left intact or
removed. It appeared that in a wetland environment, oxygen became limiting at depths within a
few mm from the surface.
4.3.1 Mechanisms of phytoremediation
Phytoremediation of petroleum hydrocarbons generally involves three major mechanisms: (1)
degradation, (2) containment and (3) the transfer of contaminants from soil to the atmosphere
(Cunningham et al., 1996; Frick et al., 1999).
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Degradation can be accomplished by both plants and their associated microorganisms. One of the
most important processes involved in the degradation is the interaction between plants and
microorganisms in the rhizosphere (root zone). Plants can stimulate the growth and metabolism
of soil microorganisms by providing root exudates of carbon, enzymes, nutrients, and oxygen,
which can result in more than 100-fold increase in microbial counts (Macek et al., 2000). This
process is also mutual beneficial. The microbes can reduce the phytotoxicity of contaminants so
that plants can grow in adverse soil conditions. Cometabolism may also play an important role in
phytodegradation. Ferro et al. (1997) suggested that plant exudates might have served as co-
metabolites in enhancing the biodegradation of pyrene in the rhizosphere.
Other major mechanisms of phytoremediation include containment of petroleum hydrocarbons
and their transfer from the soil to the atmosphere. Containment involves the accumulation of
contaminants within the plants, adsorption of contaminants onto roots, and binding of
contaminants in the rhizosphere through enzymatic activities (Cunningham et al., 1996; Frick et
al., 1999). Plants can also transport volatile petroleum hydrocarbons to the atmosphere through
leaves and stems. However, these effects are less important than the degradation mechanism in
phytoremediation of petroleum hydrocarbons (Ferro etal. 1997).
4.3.2 Considerations in application of oil phytoremediation
While phytoremediation had been used successfully within the terrestrial environment to
decontaminate soils (Banks and Schwab, 1993; Schnoor et al., 1995), the technique has not been
employed as an operational oil spill countermeasure. Until recently, only limited research had
been carried out on the effectiveness of phytoremediation in freshwater wetlands (Lin and
Mendelssohn, 1998). Most of the studies were greenhouse experiments rather than field studies.
Like bioaugmentation, studies on phytoremediation of petroleum hydrocarbons have produced
mixed results. The effectiveness of phytoremediation is site-specific, which can be affected by
factors as oil properties, types of plants, and environmental conditions.
Oil concentrations
Plants can tolerate oils with certain concentration ranges. When oil concentrations are too high,
toxic effects will lead to growth inhibition or death of plants. When oil concentrations are too
low, phytoremediation will not be effective either due to poor bioavailability. Longpre et al.
(1999) investigated the impact of oil concentrations on a freshwater wetland plant (Scirpus
pungens) along the shore of the St. Lawrence River. The results showed that the plant growth
was stimulated in the presence of crude oil at a concentration less than 4.56g/Kg sediment when
compared to the growth of the control plants. At higher oil concentrations, up to 27.4g/kg
sediment, growth inhibition or no growth increase was observed. When the oil concentration was
above 36.4 g/kg sediment, plant growth was significantly reduced. The study concluded that the
plants were likely to survive and grow in sediments contaminated with crude oil in a range of
concentrations comparable to oil spill incidents.
Plant species
Another important factor in considering phytoremediation is establishment of appropriate plants.
Lin and Mendelssohn (1996) studied the effect of oil spills on four freshwater marsh plant
species. Two of them (C. ordoratus, and A. teres) failed to survive in any of the oiled sods and E.
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quadrangulata could only persist at oil levels up to 8 L/m2. In contrast, the growth of Sagittaria
lancifolia was enhanced in response of oil addition up to 24 L/m2. Generally, legume and grass
species have been the choices for their potential use in phytoremediation of petroleum
hydrocarbons (Frick et al., 1999). Legumes are nitrogen-fixing plants, which may have
advantages in competing with non-legume species in oil-contaminated sediments. Native plants
should normally be selected since they have better chance to survive or out-compete non-
indigenous inocula (Cunningham et al., 1996).
Environmental factors
Similar to the environmental factors affecting microbial biodegradation discussed in chapter 2,
major environmental factors affecting phytoremediation of petroleum hydrocarbons include soil
types, nutrients, oxygen and temperature. Detailed description can be seen in Cunningham et al.,
1996 and Frick et al., 1999.
4.3.3 Applications in marine shoreline and freshwater wetlands
Current applications of phytoremediation in marine shorelines and freshwater wetlands have
been limited to accelerate recovery and restoration of oiled wetland. For example, mangroves
were successfully replanted to restore oil-killed mangrove forest in Panama after the 1986
Refineria Panama oil spill (Teas et al., 1989). Only a few field studies have been carried out on
the effectiveness of phytoremediation in enhancing oil degradation in marine shorelines and
freshwater wetlands.
Lin and Mendelssohn (1998) investigated the effects of biostimulation and phytoremediation in
enhancing habitat restoration and oil degradation in a coastal wetland environment (greenhouse
study). They found that application of fertilizer in conjunction with the presence of transplants
led to much higher oil degradation rates than phytoremediation alone. The results were attributed
to a higher microbial number and activity induced by the fertilizer. However, it was still not clear
whether this effect was due to biostimulation of soil microorganisms or due to phytoremediation
via fertilizer-increased plant biomass. These confounding effects perhaps could have been
distinguished by adding one more treatment (biostimulation with absence of transplants).
In 1999 and 2000, a major research study was conducted on the shoreline of the St. Lawrence
River (Garcia-Blanco et al., 2000; Venosa et al., 2002 (submitted)). The experimental design
was similar to the one used on the marine shoreline in Delaware Bay (Venosa et al., 1996). There
were 5 treatments: a no oil control and four oiled treatments. The oiled treatments included a
natural attenuation control plot with no amendments, a plot receiving ammonium nitrate and
orthophosphate nutrients but with the wetland plants continually cut back to ground surface to
suppress photosynthetic activity and growth, a plot receiving the same nutrients as Treatment B
but with the plants left intact, and a plot similar to Treatment C but with only nitrate (no
ammonium) serving as the nitrogen source. The no-oil control also received the same nutrients as
the oiled treatments receiving nutrients. Findings are summarized as follows: (1) alkane
degraders increased only marginally in all treatments while the PAH degraders were stimulated
to increase by 3.5 orders of magnitude in response to exposure to crude oil; (2) nitrogen in the
form of ammonium was partly adsorbed to negatively charged soil particles, partly taken up by
the root system of the wetland plants, and partly leached into the pore water. Nitrogen in the
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form of nitrate leached into the pore water, and some was taken up in the root system (Lee et a/.,
2002 (submitted); (3) the primary mechanism of oil mass loss from all the plots, regardless of
treatment, was physical rather than biological; (4) with respect to biodegradation of total alkanes
and PAHs during the first 21 weeks of the investigation as measured by GC/MS analysis, only
about 35% biodegradation occurred in all treatments on average, and no significant differences
among any of the treatments were observed (p > 0.05); (5) a substantial increase in plant biomass
was observed due to fertilizer addition; (6) better biodegradation occurred in surface samples in
plots where the plants had been removed than in any of the core samples because of the oxic
nature of the surface and the lack of competition for nutrients by the plant species. Enhanced
oxygen transfer to the rhizosphere by the plants through their roots did not appear to take place,
at least at the level needed by hydrocarbon degraders to metabolize the oil rapidly.
The major reason for the lack of biodegradation beyond only about 35% was ascribed to the fact
that the oil had been raked into the top 2-3 cm of sediment to make sure that penetration had
occurred. When such oil penetration occurs, little oxygen is available to allow significant
biodegradation to take place throughout the oiled zone. If oil contaminates only the surface
where more aerobic conditions exist, and if it does not penetrate deeply into the subsurface,
better biodegradation should take place, at least theoretically. The major conclusion reached
from this study was that bioremediation of an oil-contaminated freshwater wetland where
significant penetration of oil has taken place into the sediment has limited potential for enhanced
cleanup of the contamination.
In summary, the effectiveness of phytoremediation in enhancing oil degradation in freshwater
wetlands is highly site-specific and does not promise to be an effective oil cleanup technique.
However, it does show promise in accelerating the recovery and restoration of wetland
environments contaminated with oil and oil products.
4.4. Oxygen Amendment
Oxygen usually is not a limiting factor on many sandy beaches. However, oxygen limitation may
occur in wetlands and fine-grained shorelines as indicated by some field studies (Garcia-Blanco
et al., 2001b, Lee and Levy, 1991, Purandare, 1999). Under such circumstances, oxygen
amendment may be considered as a bioremediation strategy. Although oxygen supply has been
widely used for bioremediation of oil contaminated soils and groundwater, such as at many
subsurface fuel contaminated sites, this strategy has not been applied to enhance oil degradation
in marine shoreline and freshwater wetlands. This is because oxygen amendment usually
involves expensive and environmentally intrusive operations. For bioremediation of a large-scale
oil spill, use of this approach is probably not practical even when oxygen is a limiting factor.
However, under certain circumstance that involves high oil contamination in smaller scale and in
less sensitive habitats, oxygen amendment can still be considered as an alternative for oil
bioremediation. Commonly used oxygen supply techniques include tilling, forced aeration, and
chemical methods (Atlas, 1991; Brown and Crosbie, 1994; Riser-Roberts, 1998). These methods
are summarized below. They are mostly based on studies and practices in soil environments.
Special attention will be given to their potentials of application in marine shorelines and
freshwater wetlands.
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4.4.1 Tilling
Tilling has been a conventional physical method to accelerate natural oil removal by exposing
oiled sediments to a higher level of physical abrasion and biochemical degradation (Owens,
1998). This technique is also an effective means of aeration for surface layer of sediments. It has
been successfully used to accelerate biodegradation in landfarming (Atlas, 1991; Jerger et al.,
1994). Traditional tilling machines, such as disk harrows and rototillers, can aerate surface soils
to a depth of 6 to 24 inches. Sediments deeper than about 2 ft (60 cm) can be aerated by using
construction equipment, such as abackhoe (Riser-Roberts, 1998).
Currently, tilling has been recommended as a physical method to accelerate natural weathering
processes of oil in sandy or coarse-sediment beaches (Owens, 1998). The main purpose of this
practice is to increase physical abrasion of oils rather than to enhance aeration since oxygen is
usually not a limiting factor in these environments. However, based on existing experiences in
landfarming, this technique may have some potential in enhancing oil biodegradation in some
fine-sediment beaches where oxygen is limited. Tilling is also considered a low-cost technology
among the available aeration methods (Jerger et al., 1994).
Major concerns regarding this technique include disturbance of both the natural shape of
shorelines and local habitats and the potential of releasing of oil and oiled sediment into adjacent
locations. The experience from the St. Lawrence River field trial also suggests that the tilling of
surface soil may cause oil penetration deep into the shoreline sediments and may reduce the
overall oil biodegradation rates if the oil penetrates into anaerobic sediments (Garcia-Blanco et
al., 2001; Venosa et al., 2002 (submitted); see Section 5.5.2).
4.4.2 Forced aeration
Forced aeration techniques, including injection of aerated water, air and pure oxygen, are
expensive methods and commonly used for enhancing bioremediation in subsurface sediments
and groundwater contaminated with petroleum hydrocarbons (Brown and Crosbie, 1994; Riser-
Roberts, 1998). Oil contamination of coastal and wetland environments, however, usually occurs
near the surface, especially when the contamination is the result of an offshore spill.
Furthermore, crude oil rarely penetrates coastal sediments to depths of greater than one foot
(Gundlach, 1987). Therefore, these techniques of subsurface aeration are probably not
appropriate for use in bioremediation of oil spill in marine shorelines and freshwater wetlands.
4.4.3 Chemical methods
Chemical methods involve addition of alternative oxygen sources such as hydrogen peroxide
(H2O2), or alternative electron acceptors such as nitrate. Hydrogen peroxide can provide oxygen
at a rate up to two orders of magnitude faster than the forced aeration methods (Brown and
Norris, 1994). It also requires less equipment and capital cost. However, problems including too
rapid decomposition, gas blockage, and inefficient use were encountered in some sites when
using H2O2 (Brown and Norris, 1994). The chemical also can be toxic to microorganisms at high
concentrations (Riser-Roberts, 1998).
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Nitrate has received most attention as an alternative electron acceptor because it is relatively
inexpensive, very soluble in water, and does not decompose. Since nitrate is also commonly used
nutrient source for oil biostimulation, addition of nitrate may be a promising option for oil
bioremediation under oxygen limiting conditions. A potential disadvantage of this method is that
nitrate may be effective for degradation of fewer classes of compounds than oxygen. It has been
reported that nitrate-utilizing bacteria can degrade many aromatics but do not degrade aliphatic
compounds and benzene under denitrifying conditions (Brown et al., 1993). However, others
recently found that degradation of alkanes could take place under denitrifying conditions (Hess
et al., 1996). Sulfate is another potentially useful electron acceptor especially in certain marine
environments, such as salt marshes, where sulfate reduction is one of most important natural
processes (Mitsch and Gosselink, 2000). Some laboratory studies have shown that PAHs and
alkanes can be degraded under sulfate-reducing conditions at similar rates to those under aerobic
conditions in some marine sediment (Caldwell et al., 1998; Coates et al., 1997). However, these
high oil degradation rates under sulfate-reducing conditions have not been reported or
demonstrated in the field.
In summary, the potential of using oxygen amendment for enhancing oil biodegradation in
marine shorelines and freshwater wetlands is limited. Tilling may be considered as an aeration
strategy for enhancing oil biodegradation in the upper layer of sediments in less sensitive
habitats. Nitrate could be a potential alternative electron acceptor for use in a wide range of
environments, but its use as an effective enhancer of biodegradation is questionable. More
studies, particularly field trials, are still required to evaluate the effectiveness of these strategies.
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Chapter 5 GUIDELINES FOR BIOREMEDIATION OF MARINE SHORELINES
AND FRESHWATER WETLANDS: DECISION-MAKING AND
PLANNING
Existing research and applications have demonstrated that bioremediation is an effective
technology that can be used to treat certain oil-contaminated environments. Typically, it is used
as a polishing step after conventional mechanical cleanup options have been applied, although it
could also be used as a primary response strategy if the spilled oil does not exist as free product
and if the contaminated area is remote enough not to require immediate cleanup or not accessible
by mechanical tools. However, one of the major challenges in the application of oil
bioremediation is lack of guidelines regarding the selection and use of this technology. Although
extensive research has been conducted on oil bioremediation in the last decade, most existing
studies have been concentrated on either evaluating the feasibility of bioremediation for dealing
with oil contamination or testing favored products and methods (Mearns, 1997). Only a few
limited operational guidelines for bioremediation in marine shorelines have been proposed (Lee,
1995; Lee and Merlin, 1999; Swannell et al., 1996). The following two chapters will present a
more detailed and workable guidance for use by spill responders for the bioremediation of
marine shorelines and freshwater wetlands based on recent field studies and current
understandings on bioremediation processes.
As a result of recent field studies (Lee et al, 1997b; Venosa et al., 1996), we now know that there
is usually little need to add hydrocarbon degrading microorganisms because this approach has
been shown not to enhance oil degradation more than simple nutrient addition. Therefore, this
document will only present guidelines for oil bioremediation using biostimulation strategies,
mainly nutrient addition.
A general procedure or plan for the selection and application of bioremediation technology is
illustrated in Figure 5.1. The major steps in a bioremediation selection and response plan include:
1. Pre-treatment assessment - This step involves the evaluation of whether bioremediation is a
viable option based on the type of oil that has been spilled, its concentration, the presence of
hydrocarbon-degrading microorganisms, concentrations of background nutrients, the type of
shoreline that has been impacted, and other environmental factors (pH, temperature, presence
of oxygen, remoteness of the site, accessibility of the site and logistics, etc.).
2. Design of treatment and monitoring plan - After the decision is made to use bioremediation,
further assessments and planning are needed prior to the application. This involves selection
of the rate-limiting treatment agents (e.g., nutrients), determination of application strategies
for the rate-limiting agents, and design of sampling and monitoring plans.
3. Assessment and termination of treatment - After the treatment is implemented according to
the plan, assessment of treatment efficacy and determination of appropriate treatment
endpoints are performed based on chemical, toxicological, and ecological analysis.
This chapter covers operational guidelines for decision-making in the use of bioremediation and
describes the planning process for bioremediating marine shorelines and freshwater wetlands.
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The next chapter (Chapter 6) will present guidelines for assessment of field results and
establishment of appropriate treatment endpoints.
If bioremediation
is selected
Climate, prior oil
exposure and other
site characteristics
Background
nutrient content
Oil type and
concentration
Shoreline type
nutrient products
Nutrient application
strategy
Sampling and
monitoring plan
Toxicological and ecological
analysis
Analysis of oil biodegradation
and physical loss
Step 1:
Pretreatment Assessment
Step 2:
Bioremediation Planning
Step 3:
Assessment and Termination
Figure 5.1 Procedures for the selection and application of oil spill bioremediation
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5.1 Pre-treatment Assessment
Pretreatment assessment involves some preliminary investigations to assess whether
bioremediation is a viable option and determination of the rate determining process, which
include the evaluation of (1) oil types and concentrations, (2) background nutrient content, (3)
shoreline types, and (4) other environmental factors such as the prevalent climate and prior oil
exposures.
5.1.1 Oil type and concentration
Oil Type
As reviewed in Chapter 2, the biodegradability of different types of oils and petroleum products
varies greatly depending on the distribution of oil components. In general, the susceptibility to
microbial degradation for petroleum hydrocarbons is in the order of n-alkanes > branched
alkanes > low-molecular-weight aromatics > cyclic alkanes > high molecular-weight aromatics,
although this pattern is not universal (Perry, 1984). The degradation rate for the same oil
components may also vary significantly for different oils. It has been found that the rate and
extent of biodegradation of biodegradable components (e.g. n-alkanes) decreases with the
increase of non-biodegradable fractions (e.g., resins and asphaltenes) (Uraizee et al., 1998;
Westlake et al., 1974). Therefore, the heavier crude oils are less likely biodegradable than lighter
crude oils. McMillen el al. (1995) investigated the biodegradability of 17 crude oils with API
gravity ranging from 14° to 45°. They concluded that crude oil with greater than 30° API gravity
can be considered readily biodegradable, and those with less than 20° API gravity (heavier oils)
are slow to biodegrade. Similar results were obtained by other researchers (Hoff el al., 1995;
Sugiura, et al., 1997). Wang and Bartha (1990) also investigated the effects of bioremediation on
residues of fuel spills in soils. The results showed that the treatability by bioremediation for the
fuel residues are in the order of jet fuel > heating oil > diesel oil. However, more work is still
required to classify crude oils and refined products with respect to their theoretical amenability to
cleanup by bioremediation.
The biodegradation potential of oils also depends on the weathering processes that alter oil
compositions and properties. For example, evaporation leads to removal of the more toxic, lower
molecular weight components from the spilled oil. Therefore, there is less need to bioremediate a
spill of light petroleum products such as gasoline since it would evaporate rapidly. The formation
of water-in-oil emulsion may increase mass transfer limitation for oxygen and nutrients and
decrease the oil biodegradation rate. Interactions between oil and various types of shorelines also
play important roles in oil degradation, which will be discussed later. Field experience also
suggested that oils that have been subjected to substantial biodegradation might not be amenable
to bioremediation due to the accumulation of polar components in the oils (Bragg et al., 1994;
Oudet etal., 1998).
Oil concentrations
The concentration of oil is another important consideration in determining whether
bioremediation is a viable option. Very low concentrations of hydrocarbons in the environment
may be inefficiently attacked by microbes (Foght and Westlake, 1987). For sites contaminated
with oils at low concentrations, biodegradation is also less likely to be limited by nutrients or
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oxygen. Therefore, bioremediation may not be effective in enhancing biodegradation in these
cases. Natural attenuation may be a more viable option.
High concentrations of hydrocarbons may cause inhibition of biodegradation due to toxic effects,
although the inhibitory concentration varies with oil composition. Therefore, there should be an
optimum oil concentration range for bioremediation applications, below which degradation is not
easily stimulated, and above which inhibition occurs. However, this concentration range,
particularly the maximum concentration of oil amenable to bioremediation, has not been well
quantified. Field experiences in Prince William Sound, Alaska showed that less than 15 g oil/kg
sediments could be treated using bioremediation (Swannell et al., 1996). Xu el al. (2001)
recently investigated the effect of oil concentration in a microcosm study using weathered
Alaska North Slope crude oil. The results showed that crude oil concentrations as high as 80 g
oil/kg dry sand were still amenable to biodegradation. Favorable oil concentrations for
bioremediation are also related to background conditions, such as shoreline types, which will be
discussed later in this chapter.
5.1.2 Background nutrient content
Since nutrient addition has been chosen as the primary strategy to enhance oil biodegradation,
assessment of background nutrient concentrations becomes critical in determining whether
bioremediation is a viable option, whether natural attenuation should be considered, and/or
which nutrient (nitrogen or phosphorus) should be added for oil bioremediation.
As mentioned in Chapter 2, in marine environments, nutrients are generally limiting due to the
naturally low nitrogen and phosphorus concentrations in seawater (Floodgate, 1984). Nutrient
content is more variable in freshwater systems and is normally abundant in freshwater wetlands
(Cooney, 1984; Mitsch and Gosselink, 1993). However, background nutrients also depend on
other site characteristics such as local industrial and domestic effluents and agricultural runoff.
Recent field studies indicate that natural nutrient concentrations in some marine shorelines can
be high enough to sustain rapid intrinsic rates of biodegradation without human intervention
(Oudet et al., 1998; Venosa et al., 1996). The field trial in Delaware showed that although
biostimulation with inorganic mineral nutrients significantly accelerated the rate of hydrocarbon
biodegradation, the increase in biodegradation rate over the intrinsic rate (i.e. slightly greater
than twofold for the alkanes and 50% for the PAHs) would not be high enough to warrant a
recommendation to actively initiate a major, perhaps costly, bioremediation action in the event of
a large crude oil spill in that area (Venosa, et al., 1996). The high intrinsic biodegradation rate
was attributed to the high background nutrient concentrations (0.8 mg N/L on average) because
the Fowler Beach area of Delaware Bay was adjacent to farmland. The relatively high organic
content of both the Delaware Bay seawater and the underlying geology of the site and the
presence of a saltwater marsh several hundred meters landward from the beach could also
account for the high nitrogen levels encountered. The study investigators observed that
maintenance of a threshold nitrogen concentration of 3-6 mg N/L in the interstitial pore water
was biostimulatory for hydrocarbon biodegradation.
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A similar conclusion was also reached in a field trial to evaluate the influence of a slow-release
fertilizer on the biodegradation rate of crude oil spilled on interstitial sediments of an estuarine
environment in the bay of Brest, France (Oudet et al., 1998). Due to the high background levels
of N and P at the study site, no significant difference in biodegradation rates was detected
following nutrient addition. It was proposed that bioremediation by nutrient enrichment would be
of limited use if background interstitial pore water levels of N exceed 1.4 mg/L, which is close to
the finding from the Delaware study (Venosa et al., 1996).
Phosphorus is another essential nutrient related to microbial growth. Although no field study of
critical phosphorus concentrations on marine shoreline and freshwater wetlands has been
reported, it has been generally accepted that the optimal N:P ratio for microbial growth is in a
range of 5:1 to 10:1. Therefore, the threshold phosphorus concentration for maintaining optimal
hydrocarbon degradation can be derived based on this ratio and critical nitrogen concentrations
obtained from existing field studies. However, further research is still required in determining the
influence of phosphorus on oil bioremediation under various marine shoreline and freshwater
wetland environments.
These results suggest that, in the event of a catastrophic oil spill impacting a shoreline, one of the
first tasks in pretreatment assessment is to measure the natural nutrient concentrations within the
interstitial water in that environment. If they are high enough, further investigation is required to
determine whether such a nutrient loading is typical for that area and season (i.e., determine the
impact of chronic runoff from nearby agricultural practice and local industrial and domestic
effluents). The decision to use bioremediation by addition of nutrients should be based on how
high the natural levels are relative to the optimal or threshold nutrient concentrations.
5.1.3 Type of shorelines
The characteristics or type of the contaminated shoreline also play an important role in the
decision to use bioremediation. This preliminary investigation involves the assessment of the
need for bioremediation based on wave and tidal energy, the sediment characteristics, and
geomorphology of the shoreline.
Shoreline energy and hydrology
Oil can be removed rather rapidly under high wave and tide influence, typically in rocky
shorelines. In high-energy environments, bioremediation products are also more difficult to
apply successfully since they can be washed out rapidly. High wave energy will also scour
microorganisms attached to the sediment particles, and diminish the net oil biodegradation rate
that can be achieved. The Maine field study demonstrated that washout rate of nutrients from the
bioremediation zone will be strongly affected by the wave activity of the contaminated beach.
However, washout due to tidal activity alone is relatively slow, and nutrients will probably
remain in contact with oiled beach material long enough to effectively stimulate oil
biodegradation on low-energy beaches (Suidan and Wrenn, 2001; see Section 2.6.2).
However, many of the same characteristics that make low-energy beaches favorable for
bioremediation cleanup from a nutrient persistence perspective might make other conditions
unfavorable with respect to other important factors. For example, availability of oxygen is more
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favorable on high-energy beaches than on low-energy beaches. Aeration mechanisms for near-
surface coastal sediments involve exchange of oxygenated surface water with oxygen-depleted
pore water by wave-induced pumping and tidal pumping. For low energy beaches, tidal pumping
is the only likely aeration mechanism, and as a result, the surface sediments are more likely to be
anoxic than are similar depths on high-energy beaches (Brown and McLachlan, 1990). The
probability of moisture (or water activity) limitation is also higher on low-energy beaches,
because wave runup provides water to supratidal sediments on high-energy beaches during neap
tides (Suidan and Wrenn, 2001). Therefore, it is essential to thoroughly characterize the factors
that are likely to be rate limiting on each contaminated site before deciding and designing a
bioremediation response strategy.
Shoreline substrate
Although successful bioremediation application and field trials have been carried out on cobble,
medium sand, fine sand, and some salt marsh shorelines (Bragg et al., 1994; Lee and Levy,
1991; Swannell et al., 1999a; Venosa et al., 1996), different shoreline substrates or sediment
types will affect the feasibility and strategies of using bioremediation. In a 7-month field study,
Lee and Levy (1991) compared the bioremediation of a waxy crude oil on a sandy beach and a
salt marsh shoreline. Terra Nova crude oil was added at two concentrations, 3% (v/v) and 0.3%
(v/v) to beach sand and salt marsh sediments retained in in-situ enclosures in a low energy
environment. The results showed that at the lower oil concentrations (0.3%) within the sand
beach, oil biodegradation proceeded rapidly in both the fertilized plot and the unfertilized
control. The application of a bioremediation treatment provided no advantage. However, at the
higher oil concentrations (3%) on the sandy beach, oil biodegradation rates appeared to be
nutrient limited and were enhanced by nutrient addition. In contrast, the addition of nutrients to
the salt marsh sediments containing the lower (0.3%) oil concentration resulted in enhanced rates
of biodegradation. This additional need for nutrients at the lower oil concentrations is consistent
with the notion that nutrient demands within a salt marsh environment are higher, due to the size
of the microbial population within an organic-carbon rich environment. At the higher oil
concentration (3%) within the salt marsh sediments, insignificant rates of oil degradation were
reported following fertilization. The results clearly demonstrated that the success of
bioremediation depends on the characteristics of the shoreline.
On the sandy beach with low concentrations of oil, neither nutrient nor oxygen was a limiting
factor. Under these conditions, nutrient enrichment appears to provide little or no benefit, and
monitored natural attenuation can be considered as an alternative. However, at higher oil levels,
the microbial community within the sand beach may become nutrient-limited, and
bioremediation treatment could effectively enhance the rate of oil removal. In the salt marsh
environment, nutrient addition was only effective at low oil concentrations. Oxygen limitation
was more likely at higher oil concentrations due to the finer particle size and higher organic
content of the sediment in these environments. Similar results have been obtained in the field
study conducted in a freshwater wetland (Garcia-Blanco et al., 2001b; Venosa et al., 2002),
which also indicated that oxygen availability was likely a major rate-limiting factor in the
wetland environments. A field study sponsored by EPA and the Department of Fisheries and
Oceans-Canada was recently conducted on the shoreline of Nova Scotia to further investigate the
potential of using bioremediation in salt marshes. Guidelines for oil bioremediation in this type
of shoreline will be available upon completion of the data analysis from this investigation.
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5.1.4 Other factors
Prevalent climate
Prevalent climate, the ambient temperature in particular, is an important consideration when
assessing the feasibility of using bioremediation. As discussed in Chapter 2, the ambient
temperature of an environment affects both the properties of spilled oil and the activity or
population of microorganisms. At low temperatures, the viscosity of the oil increases, delaying
the onset of biodegradation (Atlas, 1981), and the volatility of toxic low-molecular-weight
hydrocarbons is reduced. Although the rates of biodegradation generally decrease with
decreasing temperature, bioremediation has been tested and applied successfully to enhance oil
biodegradation in cold arctic, alpine, and Antarctic environments (Margesin and Schineer, 1999).
This is probably because psychrophilic bacteria are plentiful and generally the dominant species
in these marine environments (Karrick, 1978).
A more important consideration regarding the effect of climate or weather on the use of
bioremediation perhaps is the seasonal factor. Significant seasonal differences in the size of
hydrocarbon degrader populations have been observed. The numbers of hydrocarbon degraders
may be much lower during winter than summer in some environments (Atlas, 1981). Oil
biodegradation slows significantly and even ceases when the contaminated sediments are frozen.
Therefore, oil bioremediation will be more effective during warmer seasons and probably should
only be considered during the summer for cold environments such as arctic regions.
Prior exposure to oil
Prior exposure of a microbial community to hydrocarbons either from natural sources (e.g.
chronic seeps and plant derived hydrocarbons) or as a result of pollution (e.g. spills and waste
disposal) may affect the rate at which subsequent hydrocarbon input can be biodegraded (Leahy
and Colwell, 1990). Those environments with a history of oil pollution or natural oil inputs have
been found to have a much higher percentage of hydrocarbon degraders and a generally greater
potential of hydrocarbon degradation than previously unpolluted areas (Atlas, 1981; Lee and
Levy, 1987, Pritchard and Costa, 1991). Therefore, for oil bioremediation in environments with
no prior oil exposure, there may be a lag and adaptation period before any significant oil
biodegradation occurs. This usually is not a concern since bioremediation itself is a relative slow
process and typically is used as a polishing step after conventional mechanical cleanup
operations. In contrast, those environments with prior exposure to oil need a shorter lag period
before initiation of biodegradation and thus will likely have a higher potential for oil
biodegradation. Thus, this type of environment is generally considered a favorable condition for
using bioremediation.
5.1.5 Summary of pretreatment assessment
In summary, the following pretreatment assessments should be conducted to determine whether
bioremediation is a viable option in response to a spill incident:
• Determine whether the spilled oil is potentially biodegradable - Light petroleum products
and light crude oils (API gravity > 30°) are relatively biodegradable; products rich in normal
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alkanes are relatively biodegradable; heavy crude oils (API gravity < 20°) and residual fuel
oils, which are high in polar compounds (asphaltenes and resins) are less biodegradable.
High concentrations of oil may also inhibit biodegradation.
• Determine whether the nutrient content at the impacted area is likely to be an important
limiting factor by measuring the background nutrient concentrations within the interstitial
water in that environment - The decision to use bioremediation by addition of nutrients
should be based on how high the natural levels are relative to the optimal or threshold
nutrient concentrations (e.g., > 2 mgN/L on sandy marine shorelines). It should also be
determined if the natural nutrient concentrations present are typical of the area or sporadic. If
sporadic, biostimulation may still be appropriate when the nutrient levels fall to limiting
values; if chronic, biostimulation may not be necessary.
• Determine whether the shoreline characteristics are favorable for using bioremediation -
High-energy rocky beaches and some low energy shorelines such as some wetlands are
considered not likely to be very amenable to nutrient addition.
• Determine whether climatic or seasonal conditions are favorable for using bioremediation -
bioremediation will be more effective during warmer seasons, particularly in cold
environments. Prior exposure to oil will also be a favorable but not a solely determinative
condition for selecting bioremediation.
5.2 Selection of Nutrient Products
After bioremediation is determined to be a potentially effective cleanup option based on the
preliminary investigations, further assessments and planning are needed before its application.
The first task is to select appropriate nutrient products through both screening tests and
assessments based on characteristics of the contaminated site.
5.2.1 Nutrient selection based on efficacy and toxicity
To assist response personnel in the selection and use of spill bioremediation agents, it is useful to
have some simple, standard methods for screening the performance and toxicity of
bioremediation products as they become available (Blenkinsopp, et al., 1995; Haines et al., 1999,
Lepo and Cripe, 1998a). One of the most comprehensive examples of such protocols is the tiered
approach developed by EPA, in cooperation with the National Environmental Technology
Applications Center (NETAC, 1993; Thomas et al., 1995). Conducting treatability tests using
micro- or mesocosms is another commonly used approach.
EPA/NET AC protocols
The NETAC/EPA protocols consist of five progressive tiers, which increase in environmental
cost and complexity with each tier of testing (Table 5.1). The approach begins with a Base Tier
in which basic information on the agent's toxicity is gathered based on a review of its
formulation. During this tier, the presence of chemicals or biological components that are
normally considered unacceptable (i.e. pathogens, carcinogens, or hazardous substances) would
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be identified. Tier I provides the basis for a preliminary evaluation of whether an agent could be
effective and safely applied, which includes a description of how the product will be used, and
information on previous usage. Tier II provides empirical evidence through the use of laboratory
shake flask treatability studies to estimate a product's effectiveness. This tier also provides
information on the relative changes in aliphatic and aromatic oil constituent concentrations over
time and the total hydrocarbon degrading microbial activity. Tier III proposes the use of flow-
through microcosm systems to study biodegradation effectiveness. Tier IV is the use of field
demonstrations to predict a product's potential effectiveness in the natural environment. Tiers III
and IV are no longer considered viable options when evaluating a product for use in an oil spill
due to overwhelming economic considerations.
It is clear that field studies can provide the most convincing demonstration of the effectiveness of
oil bioremediation because laboratory studies simply cannot simulate real world conditions such
as spatial heterogeneity, climate change, and mass transfer limitations. Since conducting a field
study just to determine that a product might work is unrealistic and economically burdensome,
the practical approach in selection of nutrient products for the bioremediation of an oil spill
would be through laboratory tests, microcosm tests in particular, as well as evaluations based on
existing field study results in similar environmental conditions.
Table 5.1 Bioremediation product test protocols developed by EPA and in cooperation with
NETAC.
Test Levels
Description
Base Tier
Collection and analysis of basic information on product safety
including formulation and acceptability of its chemical or biological
components
Tier I
Feasibility assessment concerning production capabilities, potential
effectiveness, and safety certification, including a description of how
the product will be used, and information on previous usage
Tier II
Efficacy and safety evaluation using shake flasks to compare the
degradation of artificially-weathered crude oil in natural seawater with
and without a bioremediation agent
Tier III
Efficacy and safety evaluation using microcosm systems to simulate
various environments (e.g. open water, beaches, and marshes)
Tier IV
Efficacy and safety evaluation through a field demonstration
Nutrient selection through microcosm tests
The laboratory treatability tests, especially well-designed microcosm tests, are most commonly
used approaches to determine the type and level of amendments, such as the types of fertilizer
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and the optimal nutrient concentrations. A good example of these microcosms is illustrated in
Figure 5.2, which has been used for various screening and treatability tests (Ahn, 1999, Du et al.,
1999, Xu et al., 2001). These microcosms have at least three advantages over the batch reactors
that are often used in this type of study: (1) they are connected to respirometers, allowing non-
destructive acquisition of kinetic data by continuously recording the oxygen consumption that
accompanies oil biodegradation; (2) they are open systems that can simulate the nutrient washout
that will occur in contaminated intertidal zones; and (3) they are designed to simulate tidal
flushing by filling and draining on a 12-hour cycle, thus simulating the periodic anoxia that can
occur due to tidal flooding. The potential for oxygen limitation in these reactors is a particularly
important advantage over more conventional microcosms. Because nutrient concentrations do
not limit the oil biodegradation rate when oxygen becomes sufficiently depleted, these systems
will provide more realistic estimates of maximum biodegradation rates than well-aerated shake
flasks will provide. Also, some fertilizers contain large amounts of readily biodegradable organic
compounds that can accentuate dissolved oxygen depletion. Although they might be very
effective in well-aerated microcosms, oxygen availability can limit their effectiveness in the field
(L QQetal., 1995a).
Connected to
Respirometer
Glass
Beads
Oil
Contaminated
Sands
Timer
Mineral Salts
Solution
-A
?
Pump
Sampling
Port
Figure 5.2 Schematic diagram of a beach microcosm for laboratory treatability testing of oil
spill bioremediation treatment.
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Using this microcosm system, Xu et al. (2001) investigated the effect of different nitrogen
nutrients on the bioremediation of weathered Alaska North Slope crude oil under simulated tidal
conditions. Three oil concentrations of 5, 20, and 80 g oil/kg dry sand were used. Two types of
nitrogen nutrients (KNO3 and NH4C1) were applied at a concentration of 100 mg-N/L. Oil
biodegradation was evaluated by monitoring C02 production, oxygen uptake, nitrogen
consumption, as well as oil constituent analysis. Results indicated that more biomass growth
occurred in the submerged (sometimes anoxic) portion of the sand, and better oil degradation
was observed in microcosms to which nitrate-nitrogen was applied. This result suggested that
nitrate might also have enhanced oil bioremediation by serving as an electron acceptor when
oxygen was limiting. However, the role of nitrate still requires further investigation. Ramstad
and Sveum (1995) also found that nitrate had the most pronounced effect in stimulating oil
degradation when comparing the effect of nitrate, ammonia, and an organic nitrogen product on
biodegradation of topped Statfjord crude oil in a continuous-flow seawater column system.
Effect of nutrient type may also depend on the properties of shoreline substrates. Jackson and
Pardue (1999) found that addition of ammonia as compared to nitrate appeared to more
effectively simulate degradation of crude oil in salt marsh soils in a microcosm study. The
ammonia requirement was only 20% of the concentration of nitrate to achieve the same increase
of degradation. The authors concluded that ammonia was less likely to be lost from the
microcosms by washout due to its higher adsorptive capacity to sediment organic matter.
However, in a microcosm study using sandy sediments, Suidan and Wrenn (2000) found that
there were no significant differences in the nutrient washout rates or the abilities of ammonium
and nitrate to support oil biodegradation. These results suggest that although cation-exchange
adsorption may be an important difference between ammonium and nitrate in sediments with
high cation-exchange capacities (CECs), such as marsh sediments, it is unlikely to be significant
in sediments with low CECs, such as sand.
Toxicity and other environmental impacts
In addition to demonstrating the efficacy of nutrient products in enhancing oil degradation, it is
also necessary to demonstrate that bioremediation products have low toxicity and do not produce
any undesired environmental and ecological effects. Various toxicity test protocols have been
developed (NETAC, 1993; Lepo and Cripe, 1998a, See Section, 3.5). For example, the EPA
Tier II safety evaluation consists of 7-day toxicity tests with the bioremediation product (without
oil) in natural seawater using a crustacean (Mysidopsis bahia, mysid), and a fish (Menidia
beryllina, the inland silverside). Additional Tier II toxicity tests evaluate the potential for
interaction between the product and the water-soluble-fraction of a weathered crude oil. Indirect
effects of nutrient products should also be evaluated, which include oxygen depletion through
increase in organic carbon or eutrophication, and enhanced production of toxic oil degradation
metabolites (Lepo and Cripe, 1998b).
So far, no detrimental effects from bioremediation by nutrient enrichment have been observed
following actual field operations (Prince, 1993; Mearns et al., 1997). However, the possibility
that harmful effects might occur remains. For example, oxygen depletion and production of
ammonia from excessive applications of a fish-bone meal fertilizer during a field study caused
detrimental effects, including a slowing in oil degradation rates and toxicity reduction rates
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measured by Microtox® Solid-Phase Test (Lee et al., 1995b). For safety reasons, proper
ecotoxicological assessment is always necessary in selecting nutrient products.
5.2.2 Environmental factors affecting nutrient selection
Nutrient selection also depends on environmental factors such as temperature, shoreline energy,
and substrate. A field study conducted by Lee et al. (1993) indicated that the effectiveness of
specific nutrient formulations might be influenced by temperature conditions. The study
investigated the efficacy of water-soluble inorganic fertilizers (ammonium nitrate and triple
superphosphate) and a slow release fertilizer (sulfur-coated urea) to enhance the biodegradation
of a waxy crude oil in a low energy shoreline environment. The results showed that at temperate
conditions above 15°C, the slow-release fertilizer appeared to be more effective in retaining
elevated nutrient concentrations within the sediments and more effective in enhancing oil
degradation than water-soluble fertilizers. However, lower temperatures were found to reduce
the permeability of the coating on the slow-release fertilizer and suppressed nutrient release
rates. Water-soluble fertilizers such as ammonium nitrate were then recommended under these
temperature conditions.
The action of wave, tide, and sediment type will also affect the selection of nutrients. Some
studies suggested that oleophilic fertilizers might be more suitable for use in high-energy and
coarse-grained beaches (Sveum et al., 1994; Sveum and Ladousse, 1989; See Chapter 4),
although stronger evidence is needed to confirm this suggestion. Therefore, for optimal
effectiveness, the nutrient selection should always take into account the environmental
conditions, the type of contaminated shoreline, and the methods of application, which will be
discussed later in this chapter (Lee et al., 1993; Prince, 1993; Swannell et al., 1996).
5.3 Determination of the Optimal Nutrient Loading and Application Strategy
After the initial selection of nutrient products that meet the requirements of efficacy and safety,
the next step is to determine the proper nutrient loading and the best nutrient application
strategies. Major considerations in this task include the determination of optimal nutrient
concentration, frequency of addition, and methods of addition. Finally, selection of appropriate
nutrient products should also be conducted in conjunction with this process.
5.3.1 Concentration of nutrients needed for optimal biostimulation
Since oil biodegradation largely takes place at the interface between oil and water, the
effectiveness of biostimulation depends on the nutrient concentration in the interstitial pore water
of oily sediments (Bragg et al., 1994; Venosa et al., 1996). The nutrient concentration should be
maintained at a high enough level to support maximum oil biodegradation based on the kinetics
of nutrient consumption. Higher concentrations will not only provide no added benefit but also
may lead to potentially detrimental ecological and toxicological impacts.
Studies on optimal nutrient concentrations have been conducted both in the laboratory and in the
field. Boufadel et al., (1999a) investigated the optimal nitrate concentration for alkane
biodegradation in continuous flow beach microcosms (Figure 5.2) using heptadecane as a model
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alkane immobilized onto sand particles at a loading of 2g heptadecane/kg sand. They determined
that a continuous supply of approximately 2.5 mg N/L supported maximum heptadecane
biodegradation rates. Du et al. (1999) also investigated the optimal nitrogen concentration for oil
biodegradation using weathered Alaska North Slope crude oil in the same microcosms with an
oil loading of 5g/kg sand. The results showed that nitrate concentrations below approximately 10
mg N/L limited the rate of oil biodegradation. The higher nutrient requirement was attributed to
the more complex substrate (crude oil) compared to the pure heptadecane of Boufadel et al.
(1999a). The more complex substrate (crude oil) of Du et al. (1999) also likely selected a
different population of degraders than those that grew on the pure heptadecane (Boufadel et al.,
1999a), which might have contributed to the different growth rate characteristics observed.
Ahn (1999) further studied the effect of nitrate concentrations under tidal flow conditions instead
of continuous flow. He used the same beach microcosms as Du et al. (1999) filled with sand
loaded with weathered Alaska North Slope crude oil at 5g/kg sand. A nutrient solution with
nitrate concentrations ranging from 6.25 to 400 mg N/L was supplied semi-diurnally to simulate
tidal flow. The results indicated that the optimum nitrate concentration for maximum oil
biodegradation rate was over 25 mg N/L. Some laboratory studies have reported that greater than
100 mg N/L was required to stimulate maximum biodegradation rates (Atlas and Bartha, 1992;
Reisfeld et al., 1972), but this observation probably reflects a stoichiometric rather than a kinetic
requirement, since these experiments were conducted in closed batch reactors.
Compared to the results from laboratory studies, nutrient concentrations that supported high oil
biodegradation rates were found to be lower in field studies. For example, the field tests that
were conducted after the Exxon Valdez oil spill in Prince William Sound, Alaska showed that the
rate of oil biodegradation was accelerated by average interstitial nitrogen concentrations of about
1.5 mg N/L (Bragg et al., 1994). A similar result was obtained from the study conducted in the
Bay of Brest, France, in which nitrogen was not a limiting factor when the interstitial pore water
concentrations exceeded 1.4 mg N/L (Oudet et al., 1998). The Delaware field trial also showed
that the background nitrate concentration (0.8 mg N/L) was sufficient to support fairly rapid
natural (but not maximal) rates of alkane and PAH biodegradation (Venosa et al., 1996).
Increasing the average nitrate concentration in the bioremediation zone of the experimental plots
to between 3 and 6 mg N/L resulted in a moderate increase in the oil biodegradation rate.
Observations from the referenced field studies suggest that concentrations of approximately 1 to
2 mg/L of available nitrogen in the interstitial pore water is sufficient to meet the minimum
nutrient requirement of the oil degrading microorganisms for the approximately 6-hour exposure
time to the contaminated substrate during a tidal cycle. However, laboratory microcosm results
as well as the Delaware field study suggest that higher concentrations of nitrogen can lead to
accelerated hydrocarbon biodegradation rates. Since the minimum nitrogen concentration needed
to satisfy the nitrogen demand in a tidal cycle is 1 to 2 mg N/L, and since concentrations of
nitrogen in pore water that lead closer to maximum rates of biodegradation can be several-fold to
as much as an order of magnitude higher, it is recommended that biostimulation of oil impacted
beaches should occur when nitrogen concentrations of at least 2 to as much as to 5-10 mg N/L
are maintained in the pore water with the decision on higher concentrations to be based on a
broader analysis of cost, environmental impact, and practicality. In practice, a safety factor
should be used to achieve target concentrations, which will depend on anticipated nutrient
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washout rates, selected nutrient types, and application methods. For example, in the Delaware
study, since nitrate in the interstitial pore water was quickly diluted to background levels
whenever the incoming tide completely submerged the plots, water-soluble nutrients were
applied every day using a sprinkler system. A 100-fold safety factor to account for dilution was
used to achieve the 3-6 mg/L average interstitial pore water concentrations experienced at
Delaware. A lower safety factor may be needed when using slow release nutrients.
5.3.2 Nutrient application strategies
Once the optimal nutrient concentrations have been determined, the next task is to design
nutrient application strategies, which include nutrient application frequency and delivery
methods.
Frequency of nutrient addition
The frequency of nutrient addition to maintain the optimal concentration in the interstitial pore
water mainly depends on shoreline types or nutrient washout rates. On marine shorelines,
contamination of coastal areas by oil from offshore spills usually occurs in the intertidal zone
where the washout of dissolved nutrients can be extremely rapid. Oleophilic and slow-release
formulations have been developed to maintain nutrients in contact with the oil, but most of these
rely on dissolution of the nutrients into the aqueous phase before they can be used by
hydrocarbon degraders (Safferman, 1991). Therefore, understanding the transport of dissolved
compounds in intertidal environments is critical in designing nutrient addition strategies, no
matter what type of fertilizer is used.
The Maine field study on nutrient hydrodynamics (See Section 2.6) has demonstrated that during
spring tide, nutrients can be completely removed from a high-energy beach within a single tidal
cycle. But it may take more than two weeks to achieve the same degree of washout from a low-
energy beach. Washout during the neap tide can be much slower because the bioremediation
zone will be only partially covered by water in this period. Since nutrients may be completely
washed out from high-energy beaches within a few days, and remain in low energy beaches for
several weeks, the optimal frequency of nutrient application should be based on observations of
the prevalent tidal and wave conditions in the bioremediation zone. For example, a daily nutrient
application may be needed for a high-energy beach during spring tide. But weekly or monthly
additions may be sufficient for low-energy beaches when the nutrients are applied during neap
tide. Nutrient sampling, particularly in beach pore water, must also be coordinated with nutrient
application to ensure that the nutrients become distributed throughout the contaminated area and
that target concentrations are being achieved. The frequency of nutrient addition should be
adjusted based on the nutrient monitoring results.
Methods of nutrient addition
Nutrient application methods should be determined based on the characteristics of the
contaminated environment, physical nature of the selected nutrients, and the cost of the
application.
Shoreline energy and geometry are important factors in determination of nutrient application
methods. The study in Maine suggested that surface application of nutrients may be ineffective
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on high-energy beaches because wave action in the upper intertidal zone may cause nutrients
from the surface layers of the beach to be diluted directly into the water column, resulting in their
immediate loss from the bioremediation zone. Daily application of water-soluble nutrients onto
the beach surface at low tide could be a feasible approach (Venosa et al., 1996), although this
method is highly labor-intensive. Nutrients that are released from slow-release or oleophilic
formulations will probably behave similarly to water-soluble nutrients with respect to nutrient
washout. Formulations with good long-term release characteristics probably will never achieve
optimal nutrient concentrations in environments with high washout rates. Therefore, they will not
be effective on high-energy beaches unless the release rate is designed to be high enough to
achieve adequate nutrient concentrations while the tide is out.
Another potentially effective strategy is the subsurface application of nutrients onto high-energy
beaches. Wise et al. (1994) found that application of nutrients through a trench or subsurface
drain placed above the high-tide level, rather than directly on the beach by sprinklers, would
result in significantly longer retention times. However, since nutrients move downward and
seaward during transport through the intertidal zone of sandy beaches, nutrient application
strategies that rely on subsurface introduction must provide some mechanism for insuring that
the nutrients reach the oil-contaminated area near the surface. The approach that was proposed
by Wise et al. (1994) assumes that nutrients dissolved in the freshwater plume will be brought
into contact with the oiled beach material periodically by the rising tide because the freshwater
plume should float on top of the saltwater. However, the finding of freshwater trapped between
two saltwater wedges by Boufadel et al. (1999) indicates that subsurface injection of nutrients
above the high tide level would likely not be an effective method for providing nutrients to the
bioremediation zone.
Compared to high-energy shorelines, application of nutrients on low-energy beaches is much less
problematic. Since washout due to tidal activity alone is relatively slow (Suidan and Wrenn,
2001), surface application of nutrients is an effective and economical bioremediation strategy on
low-energy beaches.
The physical forms of fertilizers are also critical in determination of appropriate nutrient
application methods. Generally, available fertilizers can be classified into four types in terms of
their physical forms: (1) slow release fertilizer briquettes, (2) dry granular fertilizers, (3) liquid
oleophilic fertilizers, and (4) water-soluble inorganic nutrient solutions (Glaser et al., 1991;
Swannell et al., 1996). The application of the briquette forms is problematic in regards to
buoyancy of the briquettes and redistribution by tide and wave action (Glaser et al., 1991). The
method used during the Exxon Valdez spill involved packing the briquettes in mesh bags tethered
to steel bars driven into the beach subsurface. The poor distribution problem occurs by
channeling of nutrients vertically down the beach rather than lateral spreading.
Dry granular fertilizers can be slow-release (e.g., Customblen in Alaska) or water-soluble, solid
granules (e.g., prilled ammonium nitrate). Granular fertilizers are easier and more flexible to
apply using commercially available whirlybird-type hand spreaders. Although this type of
fertilizer is also subject to washout by wave and tidal action, dry granular fertilizers are probably
the most cost-effective way to control nutrient concentrations. Liquid oleophilic nutrients, such
as Inipol EAP 22, are also relatively easy to apply by using hand-held or backpack sprayers. One
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of the problems when using Inipol EAP 22 during the Exxon Valdez spill cleanup was its low
pour point (11°C), which led to clogging of spraying nozzles and poor uniformity of application
(Glaser et ah, 1991). The problem was later resolved by warming the product. This type of
fertilizer is significantly more expensive than granular fertilizers. Water-soluble nutrient
solutions are normally delivered to the beach by a sprinkler system after dissolving nutrient salts
in a local water source. Although this type of nutrient may be easier to manipulate to maintain
target concentrations in interstitial pore water, its application may require more complicated
equipment such as large mixing tanks, pumps, and sprinklers.
Based on current experiences and understandings, application of dry granular fertilizer to the
impact zone at low tide is probably the most cost-effective way to control nutrient
concentrations.
5.4 Sampling and Monitoring Plan
To properly evaluate the progress of bioremediation, a comprehensive and statistically valid
sampling and monitoring plan should be developed before the application of bioremediation. The
sampling and monitoring plan should include important efficiency and toxicity variables,
environmental conditions, and sampling strategies.
5.4.1 Important variables
Important variables to be monitored in an oil bioremediation project include limiting factors for
oil biodegradation (e.g., interstitial nutrient and oxygen concentrations), evidence of oil
biodegradation (e.g., concentrations of oil and its components), microbial activity (e.g., bacterial
numbers), environmental effects (e.g., ecotoxicity levels) and other environmental conditions
(e.g., temperature and pH). A comprehensive monitoring plan proposed for a bioremediation
field study is listed in Table 5.2. A monitoring plan for a full-scale bioremediation application
should be similar to this and at least should include those critical measurements.
Since oil biodegradation in the field is usually limited by availability of nutrients, nutrient
analysis, particularly the nutrient concentrations in the pore water, is one of the most important
measurements in developing proper nutrient addition strategies and assessing the effect of oil
bioremediation. The frequency of nutrient sampling must be coordinated with nutrient
application, making certain that the treatment is reaching and penetrating the impact zone, target
concentrations of nutrients are being achieved, and toxic nutrient levels are not being reached.
Otherwise, nutrient application strategies should be adjusted accordingly. The location from
which nutrient samples are collected is also important. Recent research on solute transport in the
intertidal zone has shown that nutrients can remain in the beach subsurface for much longer time
periods than in the bioremediation zone (Wrenn et ah, 1997a). Nutrient concentration profiles
along the depth of the oil-contaminated region can be monitored by using multi-port sample
wells or sand samples collected from the oil-contaminated region.
The sampling depth should be determined based on the preliminary survey of oil distribution. It
can be established by determining the maximum depth of oil penetration, then adding a safety
factor, which will be chosen based on the observed variation in oiled depth, to ensure that the
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samples will encompass the entire oiled depth throughout the project. The safety factor will be
modified if observations during the bioremediation application suggest that the depth of oil
penetration has changed.
The success of oil bioremediation will be judged by its ability to reduce the concentration and
environmental impact of oil in the field. As discussed in Chapter 3, to effectively monitor
biodegradation under highly heterogeneous conditions, it is necessary that concentrations of
specific analytes (i.e., target alkanes and PAHs) within the oil be measured using
chromatographic techniques (e.g., GC/MS) and are reported relative to a conservative biomarker
such as hopane. On the other hand, from an operational perspective, more rapid and less costly
analytical procedures are also needed to satisfy regulators and responders on a more real time,
continual basis. Existing TPH technologies are generally not reliable and have little biological
significance (See Chapter 3). TLC-FID seems to be a promising screening tool for monitoring oil
biodegradation (Stephens et al., 1999).
Table 5.2 Monitoring plan for a bioremediation field study
Analysis
Matrix
Recommended Methods
* dissolved nitrogen
Sediment (interstitial
pore water)
extract in acidified 0.1% NaCl
4500-Norg D (persulfate oxidation to NO3")
45OO-NO3" F (automated Cd-reduction)
dissolved phosphorus
Sediment (interstitial
pore water)
extract in acidified 0.1% NaCl
4500-P B.5 (persulfate oxidation)
4500-P E (ascorbic acid method)
•residual oil
Sediment
extract into methylene chloride
analyze components by GC/MS-SIM
dissolved oxygen
Aqueous
Hach high range assay
pore-water pH
Aqueous
potentiometric with combination electrode
microbial populations
Sediment
MPN for alkane and PAH degraders
metabolic activity
Sediment
CO2 production from sand slurries
''toxicity of residual oil
Sediment
Microtox® Solid-Phase Test
toxicity of residual oil
Sediment
10-day amphipod survival bioassay
toxicity of pore-water
Aqueous
Microtox® Acute Test
toxicity of pore-water
Aqueous
sea-urchin fertilization bioassay
bioaccumulation potential
SPMDs
2-week adsorption into SPMDs
beach profile
surveying using fixed markers (e.g., wells,
plot boundary markers) in intertidal zone
beach profile
surveying relative to fixed benchmarks in
the supratidal zone
* Critical measurements
** Semi-permeable membrane devices
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In addition to monitoring the treatment efficacy for oil degradation, the bioremediation
monitoring plan should also incorporate reliable ecotoxicological endpoints to document
treatment effectiveness for toxicity reduction. Commonly used ecotoxicity monitoring
techniques, such as the Microtox® assay and an invertebrate survival bioassay, have been
summarized in chapter 3. These microscale bioassays may provide an operational endpoint
indicator for bioremediation activities on the basis of toxicity reduction (Lee et al., 1995b).
Considerations for selecting an appropriate bioremediation endpoint based on both oil
degradation and toxicity reduction will be presented in detail in the next chapter.
Other important variables in a comprehensive monitoring plan include site background
conditions (e.g., oxygen, pH and temperature) and shoreline profiles. Oxygen availability is
crucial for rapid bioremediation, because hydrocarbon biodegradation is primarily an aerobic
process. Therefore, dissolved oxygen (DO) in the pore water should be monitored on a regular
basis. The frequency of DO sampling should also be coordinated with nutrient application,
particularly when organic nutrients are used (Lee et al., 1995b; Sveum and Ramstad, 1995; See
Section 4.1.3), to insure that anoxic conditions do not result. When oxygen does become limited,
the nutrient dosage and application frequency should be adjusted accordingly. Alternatively,
oxygen amendment strategies, such as tilling or addition of chemical oxygen sources, may be
considered, although these approaches are likely to be expensive and potentially environmentally
hazardous (see Chapter 4).
Measurement of pH in the pore water is also important in monitoring oil bioremediation.
Biodegradation of oil in marine environments is optimal at a pH of about 8 (Atlas and Bartha,
1992). The pH of seawater is usually around 8.5, which is adequate to support rapid oil
biodegradation.
5.4.2 Statistical considerations in the sampling plan
To ensure that monitored results reflect the reality in a highly heterogeneous environment, it is
important that a bioremediation sampling plan be designed according to valid statistical
principles such as randomization, replication, and proper control.
To minimize bias, a random sampling plan should be used to evaluate treatment effects and their
variance within the bioremediation zone. For samples with a high degree of spatial
heterogeneity, which will be the case for most oil spill sites, stratified sampling strategies should
be used. For example, the sampling field on a marine shoreline can be divided into a number of
sectors or quadrants based on the homogeneity of geomorphology within each sector (e.g., upper
and lower intertidal zones), and independent samples should be taken in each sector according to
the rule of proportionality (e.g., taking more samples in more heavily oiled sites).
Although economic factors may be restrictive, efforts should be made to ensure that an adequate
number of samples are taken to reach a given accuracy and confidence. Power analysis should be
used to assist in the determination of sample replications required in a monitoring plan. A
statistical power test was performed in the Delaware field study to determine the number of
replicates that would be needed in future studies to detect significant treatment differences under
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similar conditions (Venosa et al., 1996). The study indicated that the required replicates to detect
treatment effects depend on expected variance and expected treatment differences. For example,
if oil distribution and shoreline characteristics are highly heterogeneous, variance will be high,
thus requiring more replicates to detect significant treatment effects. If background nutrients are
high, treatment differences will be low, and more replicates will also be required. By comparing
three shoreline assessment designs used for the Exxon Valdez oil spill, Gilfillan et al. (1999) also
proposed several strategies to increase power (i.e., the probability that significant differences
between two or more treatments are detected when indeed they exist). One of the approaches to
increase power is to select sampling sites from only the most heavily oiled locations. This
strategy may not be feasible for assessing the oil degradation within the whole bioremediation
zone, although it may be useful for evaluating the effect of bioremediation on ecological
recovery since the ecological injury most likely occurs at the heavily oiled locations.
A control area normally refers to a set-aside untreated site, which has similar physical and
biological conditions as the treated site. Although on-scene coordinators prefer not to leave oiled
sites uncleaned, it is difficult to assess the true impact of a treatment without control or set-aside
areas (Hoff and Shigenaka, 1999). When selecting control areas, one must consider not only the
similarity of the conditions but also the effect of sand and nutrient exchanges between the treated
and untreated areas (See Section 6.1.2).
5.5 Considerations for Freshwater Wetland Bioremediation
Guidelines proposed in earlier sections are mostly derived from studies and practices on marine
shorelines. However, freshwater conditions or habitats may differ sufficiently from marine
situations so that a simple transfer of response strategies may not be necessarily the most
appropriate. Special considerations for oil bioremediation in freshwater wetlands are summarized
here based on current understandings, particularly the findings of the St. Lawrence River field
study (Garcia-Blanco et al., 2001b; Venosa etal., 2002; Lee etal., 2001a).
5.5.1 Characteristics of freshwater wetlands
Wetlands occupy the interface between terrestrial and aquatic systems. They have been defined
by EPA and the U.S. Army Corps of Engineers as: "Those areas that are inundated or saturated
by surface or groundwater at a frequency or duration to support, and that under normal
conditions do support, a prevalence of vegetation typically adapted for life in saturated soil
conditions. Wetlands generally include swamps, marshes, bogs, and similar areas'' (EPA, 40
CFR 230.3 and Corps, 33 CFR 328.3). Wetlands normally should have the following three
characteristics or diagnostic parameters: (1) at least periodically, the land supports predominantly
hydrophytes (i.e., plants adapted to the flooded conditions), (2) the substrate is predominantly
undrained hydric soil (i.e., a soil with unique physical and chemical characteristics, such as
highly reduced conditions, due to repeated and prolonged saturation), and (3) the substrate is
saturated with water or covered by shallow water for a significant part of the growing season
each year (Greene, 2000; Mitsch and Gosselink, 1993).
Wetland ecosystems have enormous ecological and environmental value, contributing to aquifer
recharge, water quality improvement, flood mitigation, and shoreline erosion protection. They
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also provide unique and extensive habitats for a wide variety of flora and fauna. Furthermore,
wetlands play an important role in the global cycles of nitrogen, sulfur, methane, and carbon
dioxide (Mitsch and Gosselink, 1993).
The lower 48 states contained an estimated 100 million acres (400,000 square kilometers) of
wetlands in the mid-1980s, an area about the size of California, among which freshwater
wetlands are estimated to make up 80 to 95 percent (Mitsch and Gosselink, 1993, Greene, 2000).
Freshwater wetlands also cover extended areas of Alaska and Canada. Their proximity to areas
of human activity makes them susceptible to contamination by petroleum hydrocarbons, via
leakage, runoff or spill. Necessary measures need to be taken to protect these ecosystems since
they are among the environments most sensitive to oil and clean-up activities (Hayes et al., 1995,
1997). Furthermore, there are reports that the application of traditional oil spill cleanup
techniques in wetland habitats caused more damage than the oil itself (Baker et al., 1993). When
looking for both inexpensive and environmentally friendly technologies for wetland cleanup and
restoration, bioremediation and phytoremediation have potential for being the most suitable
options (Atlas, 1995; Cunningham et al, 1996).
The limiting conditions for oil biodegradation in freshwater wetlands may be significantly
different from most marine environments. In terms of nutrient supply, freshwater wetlands can
be divided into eutrophic wetlands (e.g., tidal marsh) and oligotrophic wetlands (e.g., cypress
domes (Mitsch and Gosselink, 1993). Many freshwater wetlands also exhibit a seasonal pattern
of uptake and release of nutrients. During the growing season (i.e., late spring and summer),
there is a high rate of nutrient uptake by vegetation from both the water and sediments. And
when higher plants die in the fall, a substantial portion of nutrients will be released to the water
and sediments. The amount of inorganic nutrients or nutrients available for oil biodegradation
also depends on many other processes, such as nutrient mineralization, denitrification, anaerobic
release of phosphorus, and wetland hydrodynamics (Mitsch and Gosselink, 2000).
When wetland soils are inundated with water, oxygen diffusion rates through the soil are
drastically reduced. Available oxygen in the soil and in the interstitial water is quickly depleted
through metabolism by aerobic organisms. Below a few centimeters—and sometimes only a few
millimeters—of the soil surface, the environment becomes anaerobic (Gambrell and Patrick,
1978). When the metabolic demand for oxygen by soil organisms exceeds that of supply, the
redox potential in the soil drops and other ions (nitrate, manganese, iron, sulfate, and carbon
dioxide) are used as electron acceptors (Mitsch and Gosselink, 1993). Therefore, in freshwater
wetland environments, petroleum biodegradation is likely to be limited by oxygen availability.
Because wetland sediments are generally more fine-grained and, more importantly, saturated
with water, the extent of oil penetration is expected to be much lower in freshwater wetlands
than on a porous sandy marine beach. In a microcosm study, Purandare (1999) investigated the
penetration of weathered light Arabian crude oil in freshwater wetland sediment under two
different water levels. For the case of low water level, where the sediment was saturated but not
covered with water, the oil was found to penetrate only about 2.5 cm in 16 weeks. For the case of
high water level, where the water level was 10 cm above the sediment surface, most of the oil
was floating on the surface of the water and the penetration depth in the sediment for some
settled and dissolved oil was also about 2.5cm during this study. The depth of oil penetration in
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the case of the St. Lawrence River study was higher (about 9 cm), due to the initial raking of the
wetland sediments after they were oiled. This depth is still much lower than the depth of oil
penetration that occurs in marine sandy beaches (up to 30 cm, Gundlach, 1987) and cobble
beaches (up to 1 m, Wolfe et al., 1994).
Another important feature of wetlands is that at least periodically, the land supports
predominantly hydrophytes, or plants "growing in water, soil, or on a substrate that is at least
periodically deficient of oxygen as a result of excessive water content." (Greene, 2000). These
wetland plants may play important roles in oil bioremediation and wetland restoration. On the
one hand, they may be involved in degradation, containment, and transfer of petroleum
hydrocarbons from the soil to the atmosphere (Frick, et al, 1999). On the other hand, these
wetland plants may also compete with hydrocarbon-degrading microorganisms for nutrients.
5.5.2. Bioremediation strategies in freshwater wetlands
Although the same decision-making and planning principles that were described earlier in this
chapter for bioremediation of marine shorelines should also apply to freshwater wetland
environments, the feasible bioremediation strategies are likely to be different due to the distinct
characteristics of wetlands. The potential effectiveness of different amendments is discussed as
follows mainly based on the findings of St. Lawrence River field study (Garcia-Blanco et al.,
2001b; Venosa et al. 2002; Lee etal., 2001a).
Nutrient amendment
Since nutrients could be limited in wetland sediments during the growing season in particular,
addition of nutrients would seem to have some potential for enhancing oil biodegradation in such
an environment. However, the results from the St. Lawrence River field study (See Section 2.5)
showed that no significant enhancement was observed in terms of the oil biodegradation
following biostimulation through addition of nutrients (either ammonium or nitrate). After 21
weeks, reduction of target parent and alkyl-substituted PAHs averaged 32% in all treatments.
Reduction of target alkanes was of similar magnitude. The removal of PAHs in nutrient-amended
plots was only slightly better than natural attenuation after 64 weeks of treatment. Oil analysis
from the top 2 cm sediment samples showed that the plots amended with ammonium nitrate and
with Scirpus pungens plants cut back demonstrated a significant enhancement in target
hydrocarbon reduction over natural attenuation as well as all other treatments. This suggests that
biostimulation may be effective only in the top layer of the soil, where aerobic conditions are
greater, and when hydrocarbon-degrading microorganisms do not have to compete for nutrients
with the growing wetland plants.
Another potential problem with respect to the use of biostimulation in wetlands is that some
plant communities may be sensitive to nutrient additions. Repeated and excessive nutrient
additions may alter the nature of the wetland ecosystem as indicated by the effects of chronic
nutrient additions to the Everglades in Florida (Davis, 1994).
Oxygen amendment
Since oxygen has been found the most likely limiting factor in oil biodegradation in freshwater
environments, oxygen amendment may be considered. However, an appropriate technology for
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increasing the oxygen concentration in such environments, other than reliance on the wetland
plants themselves to pump oxygen down to the rhizosphere through the root system, has yet to be
developed. Existing oxygen amendment technologies developed in terrestrial environments, such
as tilling, forced aeration, and chemical methods (See Section 4.4), are not likely to be cost-
effective for bioremediation of freshwater wetlands since they often involve expensive and
environmentally intrusive practices.
During the St. Lawrence River field trial (Garcia-Blanco et al., 2001b; Venosa el al., 2002), after
the first nutrient and oil applications, the top 1-2 cm surface soil in all plots was manually raked
using cast iron rakes. This was done to minimize loss of oil from the plots due to tidal action and
to uniformly incorporate the nutrients and the oil into the soil. However, the oil analysis results
suggested that the tilling of surface soil might have slowed the overall oil biodegradation rates by
enhancing oil penetration deep into the anaerobic sediments. Based on these observations,
surface tilling will not be an effective strategy for increasing the oxygen concentration in
freshwater wetlands (although this was not the intent of the raking). The slightly better but
statistically insignificant performance in both alkane and PAH degradation with addition of
nitrate compared to ammonium after 64 weeks of treatment implied that nitrate may have served
as an alternative electron acceptor in enhancing oil biodegradation when oxygen was limiting.
However, the limited increase in biodegradation rate over natural attenuation would not warrant
a recommendation to use nitrate as an oxygen amendment agent in such an environment.
Phytoremediation
Since plants cover wetlands at least periodically, the use of phytoremediation becomes a natural
option for wetland cleanup and restoration. Phytoremediation is emerging as a potentially viable
technology for cleanup of soils contaminated with petroleum hydrocarbons (Frick et al., 1999
See Section 4.3). However, this technique has not been used as a wetland oil spill
countermeasure. Only limited studies have been carried out on the effectiveness of
phytoremediation in enhancing oil degradation in marine shorelines and freshwater wetlands. Lin
and Mendelssohn (1998) found in a greenhouse study that application of fertilizers in
conjunction with the presence of transplants (S. alterniflora and S. patens) significantly enhanced
oil degradation in a coastal wetland environment. In the case of freshwater wetlands, the St.
Lawrence River study suggested that although application of fertilizers in conjunction with the
presence of a wetland plant (Scirpus pungens) may not significantly enhance oil degradation, it
could enhance habitat recovery through the stimulation of vigorous vegetative growth and
reduction of sediment toxicity and oil bioavailability (Lee et al., 2001a). The effectiveness of oil
phytoremediation in freshwater wetland environments still requires further study.
Natural attenuation
Natural attenuation has been defined as the reliance on natural processes to achieve site-specific
remedial objectives (USEPA, 1999b). This approach has been increasingly recognized as a
possible viable option for oil spill cleanup with more understanding gained over the past decade
about the advantages and disadvantages of active treatment versus natural attenuation (Owens et
al., 1999). As indicated by Sell et al. (1995), the decision-making should focus more on a
preference for natural attenuation except when a large amount of viscous oil is present, where
natural removal will be slow, or when non-ecological factors are of greater importance.
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The St. Lawrence River Study demonstrated that the availability of oxygen, not nutrients, is
likely to be the limiting factor for oil biodegradation in freshwater wetlands. However, no
feasible technique is currently available for increasing oxygen concentration under such an
environment. As a result of this study, natural attenuation is recommended as the most cost-
effective strategy for oil spill cleanup in freshwater wetlands when the oil concentration is not
high enough (e.g., less than 30 g/kg soil; Longpre etal., 1999) to destroy wetland vegetation.
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Chapter 6 GUIDELINES FOR ASSESSMENT OF FIELD RESULTS AND
TERMINATION OF TREATMENT
After the treatment is implemented according to the bioremediation plan, the next or final step in
an oil bioremediation project is to assess the treatment efficacy and terminate the bioremediation
action at appropriate treatment endpoints. Key questions to be answered in this task include
"what are the measurements of oil bioremediation success?" and "how clean is clean?" or "when
should bioremediation efforts be terminated?". Actually, these issues should be dealt with during
the bioremediation planning stage, when proper treatment objectives, strategies, and sampling
protocols should be established. On the other hand, more definite answers to these questions can
only be reached during the bioremediation actions and based on the findings of comprehensive
monitoring programs. Cost-benefit analysis (e.g., net environmental benefit analysis, Baker,
1995&1999) and sometimes political considerations should also be taken into account in this
process, however, which are beyond the scope of this document. From a technological point of
view, the measurements of bioremediation success and establishment of operational endpoints
should be based on both the efficacy of oil biodegradation and the evidence of ecotoxicity
reduction and ecological recovery, each of which will be discussed in this chapter.
6.1 Assessment of Oil Biodegradation Efficacy
6.1.1 Verification of oil biodegradation
Evidence for the effectiveness of oil bioremediation in terms of oil biodegradation should
include: (1) faster disappearance of oil in treated areas than in untreated areas, and (2) a
demonstration that biodegradation was the main reason for the increased rate of oil
disappearance. As described earlier in this document, assessing the effectiveness of oil
bioremediation in oil spill sites is difficult due to the heterogeneous conditions of contaminated
sites and lack of control over the oil distribution. Nevertheless, the success of bioremediation can
be verified through well-designed monitoring programs and proper data interpretation.
Distinguishing biodegradation from abiotic loss
Oil constituents can be lost from a shoreline by physical washout, dissolution, volatilization, and
biodegradation. To demonstrate the effectiveness of a bioremediation treatment, biodegradation
should be identified as the main mechanism for the increased rate of oil disappearance. As
described in Chapter 3, to effectively distinguish biodegradation from abiotic loss, specific oil
components or analytes should be analyzed using GC/MS techniques and then these analytes
should be normalized to a conserved biomarker, such as hopanes and chrysenes. This approach
has been successfully used to distinguish between biodegradation and the physical or chemical
loss of oil in recent bioremediation field studies (Bragg et al., 1994; Venosa et al., 1996; Lee et
al., 1997b). It should also be noted, however, that hopane normalization is most useful for
reducing the variability associated with heterogeneous oil distribution. The use of hopane
normalization to distinguish biodegradation from physical loss of oil is valid mostly when losses
due to dissolution and volatilization are negligible. Biodegradation can also be verified as the
main removal mechanism by examining the relationships between the degradation rates and the
substrate structure such as the relative degradation rates for homologous series of alkylated
PAHs. These relationships, when used in conjunction with other oil analysis data, can be very
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useful in accurate assessment of the extent and progress of oil biodegradation (Venosa et al.,
1997a; Wang et al., 1998; See section 3.3).
In addition to the demonstration of oil biodegradation based on chemical analysis, the
effectiveness of oil bioremediation can also be verified by monitoring the changes in growth and
activity of oil degrading microorganisms. Growth of hydrocarbon degrading bacteria can be
determined by Most Probable Number (MPN) techniques, particularly the procedure proposed by
Wrenn and Venosa (1996), which can separately enumerate aliphatic and aromatic hydrocarbon-
degrading bacteria and is simple enough for use in the field. Because many viable
microorganisms are unculturable (Atlas & Bartha, 1987, Macnaughton et al., 1999), the
emergent culture-independent molecular techniques, such as the PLFA-DGGE techniques (See
section 3.1.1.2), are becoming important tools to identify the diversity and composition of
uncultivated microbial communities and to enumerate bacteria in more precise ways. Other
useful tools in monitoring biological activities include in-situ measurement of microbial C02
production by the use of respirometric or radiorespirometric methods (Sugai et al., 1997;
Swannell etal., 1997).
Assessing treatment significance
To show that a treatment increases the rate of oil biodegradation, the concentrations of the target
analytes (e.g., hopane normalized total resolvable alkanes or aromatics) should be significantly
lower in treated than untreated areas within the time frame of bioremediation applications (e.g.,
several months to a year). Convincing demonstration of an increased rate of oil degradation
requires taking a sufficient number of true replicate samples that are randomly interspersed
throughout the sampling domain (see section 5.4). The field results should then be interpreted
using proper statistical analysis (Venosa et al., 1996), including analyzing field data using
standard statistical models and analysis of variance (ANOVA) techniques, which are usually
done at each sampling event. In addition to the ANOVAs, the entire data set should be analyzed
by non-linear regression analysis to estimate the rate of decline in biodegradable analytes for
each treatment over the entire course of the bioremediation treatment. As demonstrated in the
Delaware field trial (Venosa et al., 1996), the interpretation of treatment effectiveness will be
affected by reaction kinetics, background nutrient concentration, the variance of analytical
results, the number of replicates, and statistical significance required to demonstrate the
differences between treated and control areas.
An example of a likely treatment effect under different kinetic conditions is shown in Figure 6.1.
The performance of oil biodegradation expected on an oiled shoreline can be estimated based on
the first-order kinetic models described in Section 2.5.3. In the Delaware field study, first-order
biodegradation rate coefficients ranging between 0.026 and 0.056 day"1 for total resolvable
alkanes and 0.021 to 0.031 day"1 for total resolvable aromatics were observed (Venosa et al.,
1996). These first-order biodegradation rate coefficients are also a function of the nutrient
concentration (See equation 2.4). The half-saturation concentration Kn for nitrate is
approximately 0.5 mg N/L (Boufadel et al., unpublished). In this example, the background
nitrogen concentration is assumed to be 0.1 mg N/L. The assumption of low background nutrient
concentrations is reasonable since this is one of the prerequisites that bioremediation actions
should be selected. Two cases of oil biodegradation rates (i.e., 0.026 and 0.056 day"1) in the
treated areas are examined, which covers both the high (treated) and low (natural attenuation)
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ends of what was observed in the Delaware study (Venosa et al., 1996). A theoretical control rate
of 0.0093 day"1 was assumed based on a background pore water nitrogen concentration of 0.1
mg/L and a Kn of 0.5 mg/L. This is the natural attenuation rate that would have occurred had the
background nitrogen concentration been 0.1 mg/L rather than the actual 0.8 mg/L.
•S
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180
60 -
40 -
20 -
0
\
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X
k0„troi = 0 0093 day1
ktreatl = 0 026 daV"'
\
= 0.056 day"
_L
_L
0
10
70
80
90 100
20 30 40 50 60
Time (days)
Figure 6.1 Influences of biodegradation rates a on detectability of treatment effect
It can be seen from Figure 6.1 that differences between treated and control plots are prominent
very early in the project period when first order decay coefficient was 0.056 day"1. Differences
were also evident when the decay rate was only 0.026 day"1, but they were not as prominent.
Thus, had the background nitrogen concentration been closer to 0.1 mg/L, the overall conclusion
from the study would have been to recommend bioremediation since a substantial enhancement
would have been evident with nutrient addition. This would have been true even considering the
aromatic fraction, which had a lower decay rate but still higher than the theoretical control.
To determine the effect of variance on the ability to detect differences between treatments,
Figure 6.2 was developed. This figure was based on the power analysis conducted by Venosa et
al. (1996). Using 5 replicates and at a statistical power of 80% (i.e., the probability that
significant differences between two or more treatments are detected when they actually exist),
the minimum detectible difference clearly increases linearly with variance. Thus, if the variance
doubles, the ability to detect a difference between treatments lessens inversely. Figure 6.3 was
developed to show how the minimum detectable difference varies at a constant variance but at an
increasing number of replicate plots. Obviously, the fewer the number of replicate plots
established, the more difficult it is to detect statistically significant treatment differences. These
figures point out why it is so important to minimize variance by hopane normalization and
increasing the number of replicate plots.
Ill
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variance
Figure 6.2 Minimum detectible treatment difference as a function of variance.
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6.1.2 Assessment of physical loss
To better evaluate oil bioremediation performance, one should be able to distinguish between
physical loss and biodegradative loss. This requires a mass balance. However, little information
is available in this regard for most oil bioremediation field studies and applications due to the
lack of comprehensive monitoring programs, reliable measurement tools, and proper data
interpretation.
So far, the most complete mass balance of any major oil spill was the Exxon Valdez incident. In
the early 1990s, Wolfe et al. (1994) undertook a comprehensive monitoring program. They
estimated that about 3 years after the spill, approximately 20% of the spilled oil had evaporated
and undergone photolysis in the atmosphere; approximately 14% was recovered and disposed of;
approximately 2% remained on intertidal shorelines and 13% in subtidal zones. Approximately
30%) of the spilled oil was biodegraded in the water column, and nearly 20%> was biodegraded on
the shorelines. However, the paper did not provide a mass balance for the 113 km of shorelines
in Prince William Sound where bioremediation applications (nutrient additions) were performed.
A methodology was proposed by Venosa et al. (1996) to conduct a mass balance and to
distinguish biodegradation from physical loss of oil in the Delaware field study. As mentioned
earlier, spilled oils can be lost from a shoreline by physical washout, dissolution, volatilization,
and biodegradation. The method assumes that a nonbiodegradable component of oil (namely,
hopane) can be used to estimate the first three loss rates, and that the actual biodegradation rate
of an analyte can be estimated from the difference between its total loss rate and its physical loss
rate. For this approach to be valid the physical washout rate of the oil must be dominant, and the
oil loss due to dissolution and volatilization must be negligible. Volatilization and dissolution
can cause some preferential loss of oil components particularly in the early stage of an oil spill
(see Chapter 2). However, these factors are unlikely to be important during the bioremediation
treatment since bioremediation is typically used as a polishing step after conventional
mechanical cleanup options have been applied and is often initiated weeks to months after an oil
spill. Wolfe et al. (1994) reported that evaporation was no longer an important loss mechanism
three months after the Exxon Valdez spill.
Figure 6.4 shows the overall first order disappearance of hopane and total extractable organic
matter (EOM) based on the results of the Delaware study. Over 90%> of the spilled oil was
removed from the shoreline through physical washout based on the rate of hopane loss. The
EOM first-order rate coefficient was higher than the hopane disappearance rate. The difference
in loss rates between hopane and EOM was attributed to biodegradation because EOM includes
both biodegradable and nonbiodegradable components. However, because EOM and other total
petroleum hydrocarbon (TPH) measurements are not sensitive enough, no differences between
bioremediation treatments and control could be determined using this approach in the study. This
observation suggests that losses due to bioremediation may not be detectable using TPH analysis.
This again demonstrates that the success of bioremediation should be judged by analyzing
variables of biological significance, such as the reduction of concentrations of oil components of
ecological concern (e.g., PAHs), toxicity of the oil, and ultimately the ability to accelerate the
recovery of the oil contaminated ecosystem.
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Hopane: k = 0.025
EOM: k = 0.034
0.8
0.6
O
Biodegradation
O
O
0.4
Physical loss of oil
Total oil loss
0.2
0.0
0
10
20
30
40
50
60
70
80
90
100
Time, days
Figure 6.4 Fate of oil disappearance from shoreline during the Delaware study
Another important consideration in assessing physical loss of oil and bioremediation success in a
marine shoreline is the effect of shoreline substrate transport. Suidan and Wrenn (2001) found
that substantial sand transport occurs over short time scales on marine shorelines, particularly on
high-energy beaches, although the net amount of shoreline sediment did not change significantly
during the tracer study. Sediment transport will affect oil transport since oiled sand will move
with the bulk sand. As discussed in Section 3.2.2, the physical redistribution of oil between the
inside and outside of experimental plots can affect interpretation of results from bioremediation
field studies that are conducted on shorelines contaminated by real oil spills. Studies of this type
often involve treatment of oiled sand in small plots that are surrounded by large areas of
untreated, contaminated beach. Because the treated sand will become mixed with untreated sand
from surrounding areas of the beach by wave-induced sediment transport, the apparent
effectiveness of the treatment will be reduced. Hopane normalization cannot correct for this
underestimation of the treatment effect since the hopane concentration inside the plots will not be
affected by sand transport. However, this effect will be less important in cases of an actual
bioremediation application when the entire contaminated shoreline is treated. The potential effect
of shoreline substrate transport on data interpretation in these cases, however, will be to
overestimate the treatment effect due to the sand transport from large treated areas to small
untreated control areas. Therefore, the control area set aside for assessing oil bioremediation
should be either large enough to reduce this effect or relatively isolated from the treated area to
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minimize the exchange of sand and nutrients between the two areas. The influence of shoreline
substrate transport should always be taken into account in both the design of the sampling plan
and the interpretation of data from the field, particularly for high-energy beaches.
6.1.3 Operational endpoints based on oil biodegradation
Bioremediation endpoints for many soil and groundwater sites contaminated with petroleum
products are often selected based on predetermined remediation target concentrations, such as
10-10,000 mg/kg TPH and 0.1-500 mg/kg BTEX, which are adopted by various regulatory
agencies (King, et al., 1998; Salanitro el al., 1997). However, for reasons discussed previously in
regard to the inadequacies of TPH analysis, such targets are not appropriate for protecting the
environment. It is a good idea, however, for parties involved in a remediation project to have
some measurable endpoints for management and regulatory purposes. Based on existing
experiences, the following criteria are suggested in determining bioremediation endpoints with
respect to oil biodegradation.
• Bioremediation treatment should be terminated when the extent of oil degradation tends to
level off based on oil analysis results. Cost-benefit analysis should be used in establishing
target bioremediation levels. It is unrealistic and uneconomical to remove all traces of oil
hydrocarbons using bioremediation technologies.
• The concentrations of target oil analytes can also be used as endpoint indicators, particularly
when the treatment is highly effective. Emphasis should be given to those chemicals of
environmental concerns, such as PAHs. The target concentrations should be agreed upon in
the treatment plan and can be determined based on existing standards used for other
environments (e.g., oil contaminated surface and subsurface soils, Bell etal., 1994).
• The change in oil composition may also help to establish the bioremediation endpoint. As oil
becomes more biodegraded, the fraction of less biodegradable components (e.g., resins and
asphaltenes) in the oil become enriched. Studies following the Exxon Valdez spill showed
that oil biodegradation slowed substantially when the polar content of the North Slope crude
oil reached 60-70% of the total mass (Bragg et al., 1994). Therefore, the polar fraction as a
percent of the total oil mass remaining may potentially serve as an endpoint indicator
although more research is still required to establish quantitative criteria. If rapid removal of
the resin and asphaltene fractions of the oil is the desired endpoint, the only way to achieve
this is by excavation and hauling to a contained or secure facility. That is because removal of
these constituents in nature is known to occur only through dispersion and dilution. The rate
of these processes can be very site specific and depend heavily on the type of substrate (i.e.,
cobble, sand, etc.) and wave energy.
6.2 Environmental Assessments
Bioremediation is among the least intrusive of the current operational physical and chemical oil
spill countermeasure options available. Nevertheless, apprehension remains about the
environmental impact of bioremediation agents released into the environment (Hoff, 1993;
Holloway, 1991; Lee et al., 1995a; Office of Technology Assessment, 1991). In addition to
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potential effects on wildlife and humans, there is concern that the by-products of enhanced oil
biodegradation may be more toxic than the parent compounds. For general acceptance of
bioremediation as an oil spill countermeasure, we must demonstrate that it does not induce
negative effects that suppress the rates of natural habitat recovery. Environmental assessment
should be an integral part of guidelines governing the application of bioremediation treatments to
ensure protection of the environment.
Both ecosystem structure and function must be considered in environmental assessments.
Ecosystem structure is studied by examining species abundance, biomass, and diversity and other
components at one point in time. Bioassessment field surveys (Chapter 3.5.1) provide this basic
information on ecosystem structure. Ecosystem function describes the dynamics or changes in
the system over time. Information on ecosystem function is provided by the quantification of
rates of biological processes like production, respiration, mineralization, and nutrient
regeneration. In addition, bioassays (Chapter 3.5.2) provide a means of quantifying the potential
effects of toxicants on ecosystem function.
6.2.1 Operational guidance from environmental assessments for treatment application
Controlled studies suggest that during remedial operations, optimal rates of degradation in
sediments can be achieved by sustaining elevated interstitial nutrient levels that do not elicit an
adverse effect (Lee et al., 1993; Swannell et al., 1996; Lee et al., 1997b; Venosa et al., 1996;
Boufadel et al., 1999b). Bioassessment field surveys can be used to guard against detrimental
effects such as the stimulation of toxic algal blooms associated with eutrophication.
Environmental assessments can also provide guidance for bioremediation by pinpointing the
optimal time for the onset of treatment. It has become clear in numerous studies that treatments
have limited or no success as long as the residual oil is retaining its most toxic compounds,
which are typically low molecular weight compounds that are removed through natural
weathering processes. In such a case, it is better to wait one to two weeks before treatment to
allow for toxicity levels to decline, which can be delineated by time-series monitoring of
sediment or water toxicity using bioassays.
The results of toxicity tests have been used to explain the mode of action and performance results
of commercial bioremediation agents containing biostimulation (nutrients) and bioaugmentation
(bacterial inocula) properties. Lee et al. (1997b) observed that increases in microbial activity
following treatment are not necessarily correlated with toxicity reduction or habitat recovery.
Natural attenuation was more effective because the bioremediation agent inhibited the physical
loss of residual oil from the sediments.
Standard bioassay test protocols are now being developed by regulatory agencies for toxicity
evaluation of oil spill bioremediation agents (Thomas et al., 1995; Blenkinsopp et al., 1995).
Field application of bioremediation products should be limited to those that have passed
regulatory screening procedures for performance and toxicity. For a conservative approach, it is
recommended that feasibility studies be conducted prior to full-scale operations. This can be
accomplished by conducting contaminant biodegradability studies in the laboratory and
monitoring the effectiveness of the proposed treatment on several untreated but oiled shoreline
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segments. If the desired treatment endpoint (reduction of residual oil concentration or enhanced
rate of habitat recovery) has been identified, chemical analysis and bioassays can be used to
quantify the efficacy of bioremediation treatments.
6.3 Case Study: Environmental Assessment of Bioremediation Treatments in a Tidal
Freshwater Marsh
Since environmental assessment is a relatively new approach in evaluating the effectiveness of
oil bioremediation treatments, a case study is presented in detail here to help spill responders
better understand and conduct this type of assessment. To evaluate the efficacy of nutrient
amendment and phytoremediation as bioremediation strategies, a controlled oil spill field trial
was recently conducted in a tidal freshwater marsh on a shoreline of the St. Lawrence River,
Canada to determine if nutrient enrichment would enhance the rates of residual oil loss and
habitat recovery (Blanco-Garcia et al., 2001b; Venosa et al., 2002; Lee et al., 2001a). The
experimental design and bioremediation performance with respect to oil biodegradation has been
reviewed in Chapter 2 (See Section 2.5.2). Environmental assessment of the extent of habitat
recovery, which included a suite of bioassays for the identification of possible detrimental
treatment effects (e.g. toxicity of the bioremediation agent or oil degradation by-products), is
described as follows.
6.3.1 Alterations in ecosystem structure
Vegetative recovery of the predominant plant species (Scirpus pungens) was monitored by
determining changes in species composition (predominant S. pungens and secondary Eleocharis
palustris), total biomass, height, and percent cover. During the first growing season S. pungens
tolerated the experimental oil concentrations, but suffered oil-induced growth inhibition (Figure
6.5), which is consistent with the results of recent greenhouse studies (Longpre et al., 1999).
The effects of growth inhibition by oil were still evident within natural attenuation plots at Week
65. The enhanced recovery of the vegetation in nutrient amended plots during the second field
season could not be attributed to the concentration of nutrients in the porewater, as they had
diminished to background levels before the previous winter. Analysis of nitrogen content within
the roots of plants suggested that the abundant growth was attributable to the recycling of
organic nitrogen stored by these plants over the first season.
With the reduction of residual oil concentrations (primarily by physical removal) and regrowth
of the predominant vegetative species, extensive recolonization by indigenous invertebrate
species such as the mystery snail (V georgianus) was also observed during the second field
season in all experimental plots (Week 45).
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140
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Figure 6.5 Average height of the predominant plant species (Scirpus pungens) at Week 15 in
three oiled treatments (Nat. Attn., NH4+, N03") and the unoiled but fertilized (NH4+) treatment.
6.3.2 Alterations in ecosystem function
The environmental impact of contaminants in aquatic ecosystems cannot be assessed accurately
with a single species bioassay, which cannot represent the range of sensitivity of all biota. To
address this issue, a test battery with organisms from several trophic levels was used: (1)
microbial response; (2) Microtox solid phase; (3) algal solid phase; (4) cladoceran survival; (5)
amphipod survival; (6) gastropod survival/histopathology; and (7) acute and chronic effects in
fish.
6.3.2.1 Microbial response
Concentrations of bacteria are up to seven orders of magnitude higher in the surface sediments
than in the water column. This concentration of microbial activity makes sediments the most
active site for transformations of organic carbon, nitrogen, phosphorus, magnesium, and sulfur.
If the processes of decomposition, mineralization, and nutrient regeneration are disturbed by
contaminants, the nature of the ecosystem will be changed. Despite the importance of these and
other processes concentrated in the sediments, relatively few assessments have focused on
functional changes attributable to sediment-associated contaminants.
Detailed studies on the microbial response to bioremediation treatments within experimental
enclosures were conducted by Greer et al. (2000). The viable bacterial population density
showed a slight increase during the first 4 weeks following oil and fertilizer application. The
increase was clearly due to the fertilizer, as evident from the contrasting population densities in
untreated areas which remained relatively unchanged throughout the monitoring period. Patterns
of hydrocarbon mineralization activity and distribution of hydrocarbon-degrading
microorganisms can be used as an indication of in-situ biodegradation of petroleum (Braddock et
al., 1996). The microbial populations demonstrated a rapid and sustained increase in
naphthalene mineralization activity in the plots that were both oiled and fertilized. Activity was
somewhat lower in unfertilized/oiled and fertilized-only plots. Hexadecane mineralization
118
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activity increased in response to fertilizer application, especially ammonium nitrate, in
comparison to sodium nitrate: activity in the unfertilized/oiled plots and unoiled reference
control areas remained relatively low. Laboratory assays to monitor various pathways in the
nitrogen cycle (nitrification, denitrification, nitrogen fixation) have been developed (Pritchard
and Bourquin, 1985) and can be used to assess changes in function due to contaminant stress.
Field and laboratory evaluation of nitrogen metabolism indicated significant denitrification
activity in sediments following fertilizer application, which was not adversely affected by oiling.
In contrast to the results of chemical analysis that showed no treatment effect on oil
biodegradation rates, the results demonstrated that the application of fertilizers stimulated the
activities of indigenous hydrocarbon degrading and denitrifying bacteria, and the presence of oil
did not have a detrimental effect on these activities.
6.3.2.2 Microtox solid phase test
In the Microtox® Solid Phase Test (Microbics Corporation 1995), the bacterium, Vibrio fisheri,
was exposed to oiled sediment. A significant decrease in bioluminescence relative to water-only
controls was indicative of sediment toxicity. Toxicity levels were calculated as the concentration of
sample that would result in a 50% reduction in luminescence ('effective concentration,' EC50). To
account for interferences from differences in sample grain size distribution, turbidity, and to a lesser
extent, color of the sample dilutions, sample test results were compared with results from unoiled
sediments from the immediate study area.
Oil toxicity was evident on comparison of oiled with unoiled plots (Figure 6.6). While the
fertilized plots showed a trend towards a reduction in toxicity, there was no such evidence in the
natural attenuation plots over 65 weeks. Major treatment effects were not observed in the first
field season. However, in the second field season the relative toxicity levels in nutrient amended
plots were similar to values of the unoiled controls.
6.3.2.3 Algal solid phase assay
An algal solid-phase assay (ASPA) was used to assess the toxicity of sediments recovered from
the experimental test plots (Blaise and Menard, 1998). The endpoint for this assay is based on the
concentration of sediment causing 50% inhibition (IC50) of esterase enzyme activity in
Selenastrum capricornutum due to toxicants. This test is an excellent biomarker for
environmental assessments, as esterases are a key group of ubiquitous enzymes found in both
plants and animals (Dorsey et al. 1989 ; Gala and Giesy 1990).
Results of the algal solid-phase assay (Figure 6.7) showed that there was no toxicity in the
unoiled reference plots throughout the course of the experiment (IC50 > 7%), and that nutrient
additions to unoiled sediment appear to cause no detrimental effects. For the first 21 weeks,
elevated levels of toxicity were observed in all the oiled plots. Significant reduction in sediment
toxicity was not observed until the second field season. At this point, a marked decrease in
toxicity was observed for the oiled plots amended with nutrients (either ammonium nitrate or
sodium nitrate). Nutrient amended plots were deemed non toxic by Week 65. While nutrient
treatments were terminated at the end of the first field season, treatments resulted in significant
positive effects that were not observed until the second field season. This lag suggests that
toxicity reduction may be correlated with enhanced vegetative growth associated with nutrient
enrichment.
119
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Figure 6.7. Algal solid phase assay (ASPA) as IC50 for sediment samples collected at Weeks 0,
1,6, 12, 21, 49 and 65. Bars represent means ± standard deviation.
120
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6.3.2.4 Cladoceran survival test
The cladoceran Daphnia magna ('water flea') was used to assess the toxicity of elutriates from
sediment samples (Environment Canada, 1990). Treatment of unoiled plots with ammonium
nitrate and super triple phosphate appeared to cause negligible impacts on the survival of D.
magna (Figure 6.8a). In contrast, immediate, but limited, toxic responses were observed in all
oiled plots. Ammonium nitrate and triple super phosphate additions appeared to reduce residual
oil toxicity to background levels by Week 2. Amendments with sodium nitrate appeared less
effective. Since the sensitivity of this assay appeared to be limited — the effect of residual oil
was deemed negligible by Week 6 — this assay was excluded from the second field season.
6.3.2.5 Amphipod survival test
The Amphipod Test measured the effects of sediment samples on survival of sediment-dwelling
Hyalella azteca neonates, 2-9 days old (Environment Canada, 1997). Both the mean percent
survival and the mean weight of animals in each treatment were compared with mean percent
survival and mean weight of amphipods in reference control sediments to determine if the
treatments caused a significant decrease in organism survival or growth.
Hyalella azteca mortality (Figure 6.8b) was a more sensitive bioassay endpoint (i.e. higher
response) than Daphnia mortality (Figure 6.8a). This may be due to species differences and the
fact that the Amphipod Survival Test is a direct-contact sediment test. Oil derived toxicants
within the sediment may not have been effectively transferred to the elutriate used in the
cladoceran test. Furthermore, in a laboratory test, the uptake of sediment-associated anthracene
by the freshwater amphipod Hyalella azteca was reported to occur at a rate much higher than
predicted. It was concluded that selective feeding of H. azteca on smaller particles results in a
diet of fines containing the highest organic matter concentration and, hence, contaminant levels
(Landrum and Scavis, 1983). H. azteca mortality was consistently lower in the unoiled nutrient
amended plot. High mortality was observed during the first week in all oiled plots. The presence
or absence of plants appeared to have no significant effect.
At the latter part of the first field season there appeared to be a pronounced increase in sediment
toxicity in the oiled plots amended with prilled ammonium nitrate. The sensitivity of H. azteca
to ammonia (LC50 = 14.9 mg-N/L) was verified in the laboratory. Many samples, notably
Weeks 12 and 21, had overlying water ammonia levels exceeding the established toxicity limits,
with the highest values corresponding to the oiled plots amended with ammonium nitrate. The
fertilized unoiled site had lower elevated ammonia levels. This and a possible synergistic effect
between oil and elevated NH4+ concentrations may explain the observation of little or no toxic
response observed in the fertilized unoiled plots (Figure 6.8b). In response to a reduction in
residual oil concentrations, percentage mortality was near background levels by the second field
season. The results of this bioassay suggested that the experimental treatments offered little
advantage on an operational scale.
121
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Figure 6.8 Time-series changes in sediment toxicity as quantified by: a) Cladoceran Survival
Test; b) Amphipod Survival Test. Error bars = 1 standard deviation
6.3.2.6 Gastropod survival/histopathology
Although many organisms have been used as sentinels or biomonitors of environmental
contaminants (LeBlanc and Bain, 1997), there is still a need to identify and exploit alternative
species that are sensitive and amenable to ecotoxicological testing. Molluscs are abundant and
widely distributed and their use as in situ biomonitors has been on the rise (Lagadic and Caquet,
1998). In this study (Lee et al., 2001b), gastropod survival and histopathology assays were
conducted with the mystery snail, Viviparus georgianus. They were specifically selected for use
as an in situ biomonitor as they feed on sediment detritus, algae and decaying organic matter
within the wetland. Snails (n = 50/treatment/sampling time) were caged within 20 x 20 x 22 cm
open mesh polypropylene baskets moored to the sediment surface of experimental plots, and in
designated 'untreated control' areas within the vicinity of the plots. Cages were recovered after
30, 60 and 90 days of exposure to evaluate effects on survival at the end of the second field
122
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season (Week 65). Healthy snails were also exposed for a 30 day period to test sediments
recovered from the plots, under laboratory conditions, to determine survival rates.
Significant growth (p < 0.001) of the snails was observed during the study, but no significant
differences (p > 0.05) were observed in tissue weight, shell size or shell thickness between the
experimental treatments. Under adverse conditions V. georgianus is known to retract into its
shell and show no motility for prolonged periods, so animal vitality was assessed by the presence
of an operculum. No surviving snails were found in any of the continuous exposure cages after
three months (Figure 6.9). To factor out the influence of stress from the caging of animals for
extended periods, data at each sampling event were normalized to unoiled control plots adjacent
to the test blocks, which received no oil or nutrient amendments (4%, 47% and 100% mortality
at 30, 60 and 90 days, respectively). The only oiled treatment that gave a higher percentage
survival after 30 days exposure was the ammonium nitrate amended plots with intact plants. The
other treatments were more toxic than the natural attenuation plots. Elevated nutrient
concentrations for 60 days did not reduce the toxic effects of the oiled sediments to V.
georgianus. Indeed, exposure to the treatment probably exacerbated stress.
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Figure 6.9 Percent survival of the mystery snail, Viviparus georgianus, following exposure
to test treatments for 30, 60 and 90 days. Results normalized to survival in unoiled sediment
without nutrient amendment.
Toxicity of nitrogenous compounds at high levels has been noted with various organisms.
Among invertebrates, ammonia toxicity was assessed with a freshwater snail (Hickey and
Vickers 1994). Acute values were derived for the snail to be 0.15 g/m ammonia compared to the
EPA value of 0.52 for salmonids, and they found that the snail was more sensitive than the
normally accepted sensitive species such as mayflies and stoneflies. After 30 days of continuous
exposure, the highest mortality was reported in the fertilized plots with plants cut back. Direct
observations suggested that these animals suffered from harsh conditions in the absence of
natural ground cover. All treatments containing nutrients with or without oil caused higher
mortalities than no oil or oil alone. Laboratory exposures using sediment recovered from the
experimental plots at Week 65 (which had less oil, and of lower bioavailability) showed that the
application of treatments in the first season had insignificant long-term effects (Figure 6.10).
123
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Based on the results of the amphipod and snail bioassays, bioremediation using nutrients in situ
will need to take the above into consideration.
100 -,
80 -
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nh4+
no3-
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NH4+
Figure 6.10 Percent survival of the mystery snail, Viviparus georgianus under laboratory
conditions following 30 days of exposure to sediments sampled at Week 65. Error bars = 1
standard deviation
When organisms are exposed to xenobiotics, cellular changes have been observed to occur.
Evaluation of histological changes can provide important information as to the stress of the
organism and mode of action of pollutants. Specimens of live snails were preserved for
histological analysis. All tissues showed degenerative changes over time in snails exposed to
fertilizers. Pathological changes were most obvious in epithelial cells lining such tissues as gills,
intestine and digestive gland, but some degenerative changes were also observed in other organs
such as gonads. The most dramatic changes were observed in snails from plots treated with
sodium nitrate (Figure 6.11). Effects of treatment on reproductive success of V. georgianus were
noted. Gravid females contained on average 10 young (range 3 to 15) visible with the naked eye.
At 1-week post oil exposure, gravid females showed some degenerating embryos in all but the
control treatments. This was also true for some of the 1-month post exposure gravid females, but
number of gravid females with macroscopic embryos decreased.
6.3.2.7 Acute and chronic effects on fish
Experience with the Exxon Valdez demonstrated that oil deposition in shoals where pink salmon
(Oncorhynchus gorbuscha) and Pacific herring (Clupea pallasi) spawn, caused blue sac disease
(BSD) of newly hatched larvae (Hose et al., 1996; Marty et al., 1997). BSD is characterized by
yolk sac and pericardial edema, hemorrhaging, deformities, and induction of mixed function
oxygenase enzymes (Marty et al., 1997; Billiard et al., 1999; Fragoso et al., 1998). As a
consequence, there is considerable interest in the application of bioremediation strategies that
accelerate the removal of oil from intertidal beaches, which provide the nursery habitat for many
fish species. The success of these technologies should be judged not simply on how quickly oil
disappears, but on a demonstrated reduction in risk, i.e. on how quickly toxicity (hazard) and
exposure to oil are reduced.
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Figure 6.11. Representative sections through the intestine of Viviparus showing normal villi
seen in control snails (a) and degenerated villus with hemocytic infiltration seen in snails from
the sodium nitrate treatment after 1 month exposure (b). Bar = 75 (am.
Establishing the exposure of fish to oil is difficult. While chemical measures of oil in sediments,
water and tissues are routine, there is no guarantee that fish accumulate oil or its components
equally or in proportion to environmental concentrations. Further, many of the components of
oil (e.g. alkanes, polynuclear aromatic hydrocarbons (PAH)), are metabolized by fish, so that
chemical analyses of fish tissues may not represent the true dose or dose rate.
Assessing hazard is equally difficult. Oil is a complex mixture, and many of its components
have different modes of toxicity. At acutely lethal concentrations, mortality is rapid and is likely
due to narcosis caused by monoaromatics (benzenes, toluenes, xylenes, etc.) and alkanes. At
lower concentrations (Billiard et ah, 1999), chronic toxicity (BSD) may be linked to the
concentrations of alkyl PAH such as 7-isopropyl-l-methylphenanthrene (retene, a C-4
phenanthrene). This mechanism may involve metabolism of alkyl PAH to more toxic forms by
CYP1A enzymes. Toxicity and rate of excretion of phenanthrene and retene can be modulated
by inhibiting or inducing CYP1A activity (Hawkins et ah, 2000). Delayed responses, such as the
long-term onset of cancer, may also follow brief exposures to pro-carcinogens such as
benzo(a)pyrene (BaP). In this case, the mechanism involves oxygenation of BaP by CYP1A
enzymes to carcinogenic diols and epoxides, which cause genotoxicity by forming adducts with
bases of DNA (Varanasi et ah, 1989).
In this case study, bioassays of CYP1A induction in juvenile rainbow trout (Oncorhynchus
mykiss). BSD in embryo-larval stages (ELS) of trout, and reproductive developmental studies
125
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with Japanese medaka (Oryzias latipes) were used to evaluate the presence, bioavailability, and
toxicity of PAH (and parent mixture of oil) in contaminated beach sediments (Hodson et al.,
2001). Changes in bioavailability and toxicity with time were monitored by analysis of time-
series samples.
The key to sediment assessment is bioavailability. Although sediments might contain relatively
high concentrations of toxic compounds, this condition does not necessarily lead to adverse
effects on organisms living in the sediments (Payne et al., 1988). The only means of measuring
bioavailability is by measuring or determining biological response. Such testing has often
involved measures of bioaccumulation (the ability of an organism to accumulate contaminants in
tissues. However, because bioaccumulation is a phenomenon, not an effect (and can be
relatively expensive due to costly chemical analyses), emphasis has shifted towards indicative
endpoints that are based on sediment toxicity tests, which are effects-based and relatively
inexpensive.
Bioavailability was assessed by the extent of CYP1A induction in fingerling trout after a 4 d
exposure to sediments. S-9 fractions were prepared from liver homogenates and activity of
ethoxyresorufin-o-deethylase (EROD, CYP1A enzyme) was measured by a kinetic microplate
fluorescence method following the protocols outlined by Hodson et al. (1996). Each bioassay
included negative controls (water only), unoiled sediment controls, and positive water controls
(fish exposed to B-naphthoflavone, a model inducer). Results indicated that one day after oiling,
EROD activity of fish exposed to oiled sediments was on average 25-fold higher than that in the
control sediments (range = 13-35-fold). Over a 17 month period, there was about an 80%
decline in induction potency by oiled sediments sampled from the beach. The decline in EROD
activity paralleled declines in total hydrocarbons, total PAH, and total alkyl-PAH, measured by
GC/MS with correlation coefficients between log of EROD activity and log of hydrocarbon
concentrations of 0.96 or higher (Hodson et al., 2001).
Blue sac disease of trout was assessed by exposing 50 eyed-eggs (about 15 d post fertilization) to
sediments until they had hatched, resorbed the majority of their yolk sac, and begun to swim up
(Zambon et al., 2000). After 32 d of exposure, larvae were removed and the % survival and
prevalence of the symptoms of BSD (edema, hemorrhaging, deformities) were measured
(Guiney et al., 1997). Results showed that trout embryos and larvae exposed to oiled sediments
from Ste. Croix exhibited a low prevalence of symptoms of BSD, although rates were higher
than observed for fish exposed to the reference sediments. All oil-exposed fish examined
histologically demonstrated intense staining for CYP1A protein, indicating a significant
exposure to CYPlA-inducing compounds such as PAH. These preliminary results suggested
that symptoms of BSD were more frequent in fish exposed to oiled than to un-oiled sediments,
indicating a risk to early life stages of species that spawn on tidal beaches.
Medaka larvae exposed to oiled sediments for 90 days (maturity) experienced a higher mortality
rate (42.5± 4.2%) than those exposed to un-oiled sediments (9.1±7.4). The growth of surviving
medaka in the oiled sediment treatments was impaired (Figure 6.12), as shown by their smaller
size compared to controls (20.6±1.3 vs 16.4±3.3 mm). In addition, nearly all medaka from the
oiled sediment treatments had deformed or missing fins. Preliminary histological examination of
126
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fixed medaka indicated a degenerative liver lesion, spongiosis hepatis, in >95% of the fish from
the oiled-sediments, but not in controls. There was also a low incidence of male medaka (<10%)
with intersex gonads (testis-ova) in fish exposed to oiled sediments.
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Figure 6.12. Effect of exposure to Ste. Croix oiled sediment (Oct. 1999) on growth of medaka.
Treatments sharing the same letters are not statistically different.
The laboratory fish bioassay represents a 'worst-case' scenario as the fish could not avoid
exposure. Furthermore, mixing of the sediments in the test chambers might have disrupted the
top-most sediment layer (that might have been depleted of oil in the field by weathering) and
exposed more oily layers of sediment. The ratio of water to sediment is also fixed, which is very
different from the situation in well-flushed tidal beaches. Finally, the test organisms were not
beach spawning species of fish such as smelt, capelin, and herring.
Semi-permeable membrane devices (SPMDs) were used to assess in-situ changes in the
bioavailability and toxicity of residual oil components following treatment. These passive in-situ
samplers, developed by Huckins et al. (1990), contain purified triolein, a substance that
constitutes a major fraction of the neutral lipid of fish. When immersed in water, SPMDs absorb
non-ionic, organic chemicals having a log K0W>1, a size <1 A, a molecular weight of about 600
or less, and possibly neutral organo-metal complexes. These characteristics correspond to that of
known mixed-function oxygenase (MFO) inducing compounds, including PAHs. The
operational advantage of SPMDs is that they can be deployed within the environment to provide
an integrated sample over time. This characteristic makes the assay highly advantageous for
field use. In terms of ecological relevance, the diffusion of dissolved neutral organic chemicals
into the triolein through the pores of the polyethylene membrane within SPMDs simulates the
diffusion of compounds across a live fish gill membrane. The lipid can be analyzed by
traditional chemical techniques to provide a list of chemicals absorbed, their concentrations in
127
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the SPMDs, and, by back calculation, their concentrations in water or sediment. In this study
SPMDs were used as concentrating devices for biological testing with the Microtox® Assay.
SPMD units (12.5 cm) enclosed in protective cases were deployed at the water sediment
interface at predetermined intervals and recovered for analysis after one week of exposure.
SPMD concentrated samples were diluted with organic solvents for toxicological analysis by
Microtox. The results showed that by 4 weeks, the toxicity of sediments within the oiled plots
had declined to the levels of the unoiled but fertilized control as the result of a reduction in
bioavailability (Johnson et al., 2000).
6.4 Ecotoxicological Tests for Risk Assessment
The application of in-situ bioremediation operations is not expected to generate a large volume
of waste materials like ex-situ operations do. If the program is effective, any residual
hydrocarbons will elicit little or no biological effect due to physical, chemical, and biological
processes that reduce their bioavailability. As illustrated in the case study, sediment quality can
be assessed by a number of methods that tend to fit into one of five categories: sediment
chemistry, sediment toxicity, community structure, tissue chemistry, and pathology. Ideally, all
five components would be utilized to assess sediment quality. However, in reality,
environmental managers are faced with limitations in both resources and time. They must
optimize use of their resources by selecting the information that will have the greatest utility.
Thus, monitoring programs should be focused on the measurement of variables that allow
quantification of treatment success against a pre-defined endpoint. Project coordinators must
strive to strike a balance between level of effort and type or quality of information needed to
make effective decisions.
While detrimental effects have not been linked to the application of bioremediation strategies
based on nutrient enrichment in actual spill response operations (Mearns et al., 1997; Prince,
1993), the results of recent field trials clearly demonstrate that the possibility exists. In the case
study presented (Section 6.3), improper application of bioremediation agents (the addition of
excess fertilizer) was detrimental to the environment. For example, synergistic effects between
ammonia and the test oil were observed in the Amphipod Survival Test. However, the nutrient
additions had no effect on the unicellular algal species, Selenastrum capricornutum, and actually
enhanced the growth and productivity of the dominant plant species within the oiled plots
relative to those that received no treatment. Rapid recovery of vegetation is critical within an
impacted wetland for erosion control. Depending on the desired endpoint of the remedial
operation, a balance must be made among the many positive (e.g. enhanced recovery of
vegetation) and negative effects of treatment (e.g. changes in productivity, species composition,
and diversity among the remaining wetland plants).
With refinement, bioassays can be used in oil spill operations to provide real-time guidance to
the treatment operations (e.g. determining the optimal nutrient concentration that does not elicit a
detrimental effect), to verify the success of countermeasures and to quantify the extent of habitat
recovery. The demonstration of species dependent responses in previous investigations and the
current case study suggest that future environmental risk assessments be based on a multi-
species, multi-trophic level test battery approach. The results of the ecotoxicological tests should
128
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be used to build an ecological risk assessment — an estimate of the probability of harm to the
aquatic environment derived from the synthesis of results of separate exposure and effects
components in a scientific manner (Gentile et a I., 1989). It is often stated that the objective of
oil spill countermeasures are to return an impacted site to its immediate pre-spill condition. This
is an unrealistic goal as the environment is a dynamic rather than a static system. On an
operational scale, our goal should be to return the structure and function of an ecosystem to
within the limits of pre-defined, acceptable criteria.
6.5 Ecotoxicological Tests to Identify Operational Endpoints
The effectiveness of oil spill countermeasures ultimately must be judged by their ability to
reduce injury to aquatic life. The worse case scenario would be the use of a response method
that is not effective in reducing exposure and increases injury to aquatic life. In bioremediation
operations the application of ecotoxicological monitoring protocols may be used to verify the
efficacy for toxicity reduction over that of no treatment.
Bioremediation treatments should be terminated when it is deemed that: (1) treatments offer no
operational advantage over natural recovery, (2) the contaminant concentrations and toxicity
values are reduced to acceptable levels, or (3) detrimental effects from the treatment strategy are
identified. Cost-benefit analysis should be considered in the decision of the acceptable level. It
is futile to expect bioremediation techniques to remove all traces of residual hydrocarbons. In
terms of ecological relevance, declaration of habitat recovery can be made when toxicity limits
are within regulatory guidelines and there is evidence for the return of the original community
structure (Lee et al., 1995b; Mearns etal., 1995).
129
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