EPA/600/R-04/074
July 2004
GUIDELINES FOR THE BIOREMEDIATION OF
OIL-CONTAMINATED SALT MARSHES
by
^ueqing Zhu, 2Albert D. Venosa, ^akram T. Suidan, and 3Kenneth Lee
University of Cincinnati
Cincinnati, OH 45221
2U.S. Environmental Protection Agency
National Risk Management Research Laboratory
Cincinnati, OH 45268
3Department of Fisheries and Oceans-Canada
Bedford Institute of Oceanography
Dartmouth, Nova Scotia B2Y 4A2
EPA Contract No. 68-C-00-159
Task Order No. 8
Task Order Manager
Albert D. Venosa
Land Remediation and Pollution Control Division
National Risk Management Research Laboratory
Cincinnati, OH 45268
National Risk Management Research Laboratory
Office of Research and Development
U.S. Environmental Protection Agency
Cincinnati, OH 45268
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Disclaimer
The information in this document has been funded by the United States Environmental Protection
Agency (U.S. EPA) under Task Order No. 8 of Contract No. 68-C-00-159 to the University of
Cincinnati. It has been subjected to the Agency's peer and administrative reviews and has been
approved for publication as an EPA document. Mention of trade names or commercial products does
not constitute an endorsement or recommendation for use.
11
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Foreword
The U.S. Environmental Protection Agency (EPA) is charged by Congress with protecting the
Nation's land, air, and water resources. Under a mandate of national environmental laws, the
Agency strives to formulate and implement actions leading to a compatible balance between human
activities and the ability of natural systems to support and nurture life. To meet this mandate, EPA's
research program is providing data and technical support for solving environmental problems today
and building a science knowledge base necessary to manage our ecological resources wisely,
understand how pollutants affect our health, and prevent or reduce environmental risks in the future.
The National Risk Management Research Laboratory (NRMRL) is the Agency's center for
investigation of technological and management approaches for preventing and reducing risks from
pollution that threaten human health and the environment. The focus of the Laboratory's research
program is on methods and their cost-effectiveness for prevention and control of pollution to air,
land, water, and subsurface resources; protection of water quality in public water systems;
remediation of contaminated sites, sediments and ground water; prevention and control of indoor air
pollution; and restoration of ecosystems. NRMRL collaborates with both public and private sector
partners to foster technologies that reduce the cost of compliance and to anticipate emerging
problems. NRMRL's research provides solutions to environmental problems by: developing and
promoting technologies that protect and improve the environment; advancing scientific and
engineering information to support regulatory and policy decisions; and providing the technical
support and information transfer to ensure implementation of environmental regulations and
strategies at the national, state, and community levels.
This publication has been produced as part of the Laboratory's strategic long-term research plan. It
is published and made available by EPA's Office of Research and Development to assist the user
community and to link researchers with their clients.
Lawrence W. Reiter, Acting Director
National Risk Management Research Laboratory
in
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EXECUTIVE SUMMARY
Salt marshes are among the most sensitive ecosystems and, therefore, the most difficult to clean.
Applications of some traditional oil spill cleanup techniques in wetland habitats have caused more
damage than the oil itself. The objective of this document is to present a detailed technical guideline
for use by spill responders for the cleanup of coastal wetlands contaminated with oil and oil products
by using one of the least intrusive approaches - bioremediation technology. This manual is a
supplement of the previously published "Guidelines for the Bioremediation of Marine Shorelines
and Freshwater Wetlands" (Zhu et a/., 2001), which has focused on the bioremediation of sandy
marine shorelines and freshwater wetlands. This guidance document includes a thorough review and
critique of the literature and theories pertinent to oil biodegradation and nutrient dynamics and
provides examples of bioremediation options and case studies of oil bioremediation in coastal
wetland environments. It also evaluates current practices and state-of-the-art research results
pertaining to the bioremediation of hydrocarbon contamination, and presents a procedure for the
design and evaluation of bioremediation processes applicable to the cleanup of oil contaminated
coastal wetlands. Special attention is given to oil bioremediation of salt marshes since they are the
most prevalent type of coastal wetland and have been the subject of the most extensive studies.
The document consists of two major parts. Part I presents the background and overview of
bioremediation options, which include the characteristics of coastal wetlands, oil spill threats and
countermeasures in salt marshes, and relevant state-of-the-art research. Part II provides guidelines
for design and planning of oil bioremediation in salt marshes, which includes site characterization
and evaluation, the selection of appropriate bioremediation technologies, and the design of sampling
and monitoring programs.
The overall conclusions reached by the guidance manual are as follows. Unlike sandy beaches, oil
biodegradation on marine wetlands is often limited by oxygen, not nutrient availability. Natural
attenuation is increasingly becoming the preferred strategy for the restoration of oil-contaminated
wetlands. However, field studies also show that on some coastal wetlands, nutrients might still be a
limiting factor for oil biodegradation, particularly if the oil does not penetrate deeply into the anoxic
zone of the wetland sediment. When biostimulation is selected, it is recommended that nitrogen
concentrations of at least 2 to as much as 10 mg N/L should be maintained in the pore water to
achieve optimal oil biodegradation, with the decision on higher concentrations to be based on a
broader analysis of cost, environmental impact, and practicality. Furthermore, if ecosystem
restoration is the primary goal rather than oil cleanup, at least one 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 resulting from the
addition of extra nutrients. Therefore, the decision to bioremediate a site should depend on cleanup,
restoration, and habitat protection objectives and other pertinent factors that may have an impact on
success.
No effort was made to determine the quality of secondary data reviewed in the literature and the
conclusions made from these data.
IV
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TABLE OF CONTENTS
Introduction and Overview of Bioremediation Options 1
1.1 Coastal Wetlands in the U.S 1
1.2 Oil Spills in Salt Marshes: Threats and Countermeasures 2
1.2.1 Threats of oil spills 2
1.2.1.1 Impact to wetland plants 2
1.2.1.2 Impact to wildlife and ecosystems 3
1.2.2 Response to oil spills in salt marshes 4
1.2.2.1 Physical Methods 4
1.2.2.2 Chemical methods 5
1.2.2.3 In-situ burning 5
1.2.2.4 Restoration 6
1.3 Bioremediation of Oil Spills in Salt Marshes 6
1.3.1 Environmental factors affecting oil biodegradation in salt marshes 6
1.3.2 Laboratory studies 8
1.3.3 Full-scale demonstrations 10
1.3.3.1 Nova Scotia, Canada, 1989 10
1.3.3.2 San Jacinto Wetland Research Facility (SJWRF), Texas, 1994-1997 10
1.3.3.3 Terrebonne Parish, Louisiana, 1998 11
1.3.3.4 Gladstone, Australia, 1997-1998 11
1.3.3.5 Nova Scotia, Canada, 2000-2001 12
1.3.4 Kinetics of oil biodegradation 15
1.3.5 Monitoring biological responses to quantify the efficacy of remediation treatment 16
1.3.5.1 Bioassessment 16
1.3.5.2 Bioassays 17
1.3.6 Bioremediation options on salt marshes 18
1.3.6.1 Nutrient Amendment 18
1.3.6.2 Microbial amendments 19
1.3.6.3 Oxygen amendment 20
1.3.6.4 Plant amendment (phytoremediation) 20
1.3.6.5 Monitored natural attenuation 21
Recommended approaches to bioremediation IN SALT MARSHES 23
2.1 Pre-treatment Assessment 25
2.1.1 Oil penetration and oxygen availability 25
2.1.2 Background nutrient content 26
2.1.3 Summary of pretreatment assessment 28
2.2 Treatment Selection and Design 28
2.2.1 Nutrient selection 29
2.2.2 Concentrations of nutrients needed for optimal bio stimulation 30
2.2.3 Nutrient application strategies 31
2.2.3.1 Frequency of nutrient addition 31
2.2.3.2 Methods of nutrient addition 31
2.2.4 Sampling and Monitoring Plan for Bioremediation Operations 32
2.2.4.1 Important variables and recommended measurements 32
2.2.5 Environmental assessment of an oil-contaminated salt marsh: a case study 35
2.2.5.1 Bioassessments 36
2.2.5.2 Risk assessment 46
2.3 Summary and Recommendations 47
REFERENCES 49
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1 INTRODUCTION AND OVERVIEW OF BIOREMEDIATION OPTIONS
1.1 Coastal Wetlands in the U.S.
Coastal wetlands are subjected to the influence of tidal action. They provide natural barriers to
shoreline erosion, habitats for a wide range of wildlife including endangered species, and key
sources of organic materials and nutrients for marine communities (Boorman, 1999, Mitsch and
Gosselink, 2000). Coastal wetlands can be classified into tidal salt marshes, tidal fresh water
marshes, and mangrove swamps (Mitsch and Gosselink, 2000).
• Tidal Salt Marshes — Salt marshes are those halophytic grasslands found in the middle and
high latitudes along protected coastlines. They are subjected to tidal action as well as high
salinities. In the United States, they are often dominated by the grass Spartma alterniflora in
the low intertidal zone, and Spartma patens with the rush Juncus in the upper intertidal zone.
Most of these wetlands are distributed along the Gulf of Mexico and the Atlantic coast.
• Tidal Freshwater Marshes ~ These wetlands are found inland from the salt marshes but
still close enough to the coast to experience freshwater tidal effects. Since these wetlands
lack the salinity stress of salt marshes, they are often very productive ecosystems and
dominated by a variety of grasses and by perennial and annual broad-leafed aquatic plants.
• Mangrove Swamps — Mangroves are subtropical and tropical coastal wetlands dominated
by halophytic trees and shrubs. In subtropical and tropical regions of the world, tidal salt
marshes give way to mangrove swamps. In the United States, they are mostly distributed
along the southern coast of Florida and generally dominated by the red mangrove
(Rhizophord) and the black mangrove tree (Avicennid).
It the early 1990s, it was estimated that the total area of coastal wetlands in the United States was
approximately 3.2 million ha (32,000 km2), with about 1.9 million ha or 60 percent of the total
coastal wetlands as salt marshes and 0.5 million ha as mangrove swamps (Mitsch and Gosselink,
2000). Coastal wetlands are no longer viewed as intertidal wastelands, and their ecological and
economic values have been increasingly recognized. Major benefits and functions of coastal
wetlands include:
• Shoreline Protection - Coastal wetlands provide a buffer between land and sea, protecting
marine shorelines from the ravages of storms and erosion by wave action. Salt marshes,
which sustain little damage from ocean storms, can shelter inland developed areas and
reduce potential storm damage to coastal buildings and structures.
• Support of Coastal Fisheries - Tidal marshes provide spawning site and nursery areas for
many fish and shellfish species. Due to their high productivity, coastal wetlands produce
great volumes of detrital organic materials and nutrients, on which many small invertebrates
and fish feed. It is estimated that over 95 percent of the commercial fish and shellfish species
in the United States are wetland dependent (Feierabend and Zelazny, 1987).
• Wildlife Habitat - Coastal wetlands are the primary habitat for many plant and animal
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species and provide food, water, and shelter to indigenous and migratory species. More
importantly, wetland habitats are essential for the survival of a large percentage of
endangered species. For example, of the 209 animal species listed as endangered by the U.S.
Department of Interior and the U.S. Fish and Wildlife Service in 1986, about 50 percent
depend on wetlands for their survival (Mitsch and Gosselink, 1993).
• Water Quality Management - Coastal wetlands maintain and improve water quality by
acting as sediment and chemical sinks (Baker et a/., 1989). Under favorable conditions,
wetland sediments, plants, and their associated microorganisms are able to contain, take up,
and degrade various environmental contaminants, such as excess fertilizers, pesticides, and
heavy metals.
Wetlands have suffered dramatic losses as a result of human activities, such as drainage for
agricultural use. Overall, more than 50 percent of the wetlands in the continental U. S. were lost from
the 1780s to the 1980s (Mitsch and Gosselink, 2000), and at a rate 7,300 ha/year from 1950s to
1970s (Tiner, 1984). Such losses have greatly diminished the nation's wetlands and their benefits.
The rates of loss have been declining since the mid 1970s with the enactment of wetland protection
laws and increased public appreciation. However, threats to coastal wetlands remain, including
conversion for agricultural, industrial, and residential development, mean sea level rise, and
chemical contamination from excessive nutrient inputs, chemical accumulations, and oil spills.
The threat of crude oil contamination to coastal wetlands is particularly high in certain parts of the
U.S., such as the Gulf of Mexico, where oil exploration, production, transportation, and refineries
are extensive (Lin and Mendelssohn, 1998). Oil and gas extraction activities in coastal marshes
along the Gulf of Mexico have been one of the leading causes of wetland loss in the 1970s (Mitsch
and Gosselink, 2000). Despite more stringent environmental regulations, the risk of an oil spill
affecting these ecosystems is still high because of extensive coastal oil production, refining, and
transportation.
1.2 Oil Spills in Salt Marshes: Threats and Countermeasures
1.2.1 Threats of oil spills
Marine wetlands are especially vulnerable to oil spills because the inherently low wave energy of a
wetland does not physically remove oil effectively. They are flooded at high tide and their complex
surface can trap large amounts of oil. Impacts of oil spills to coastal wetland ecosystems have been
described and reviewed extensively (Baker etal., 1989; Fingas, 2001; NAS, 1985; Pezeshki etal.,
2000). A brief summary on these impacts is provided in the following text.
1.2.1.1 Impact to wetland plants
Oil spills have been known to cause acute and long-term damage to salt marshes and mangroves
(Baker et a/., 1989; Burns et a/., 1993; Lin and Mendelssohn, 1996; Pezeshki et al., 2000). These
impacts include reduction in population and growth rate or abnormal growth and regrowth after
initial impact. Mangroves are generally more vulnerable to oil spills than salt marshes because oil on
the partially submerged roots of mangroves interferes with respiratory activity (Duke etal., 1997;
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Evans, 1985).
The degree of oil impact also depends on various factors, such as the type and amount of oil, the
extent of oil coverage, the plant species, the season of the spill, the soil composition, and the
flushing rate. For example, No.2 fuel oil has been found to cause much higher mortality and damage
to Spartina alterniflora., a dominant salt marsh grass along the Atlantic coast and Gulf of Mexico,
than Arabian crude oil, Libyan crude oil, and No.6 fuel oil (Alexander and Webb, 1983 & 1985).
Growth of Spartina alterniflora was not significantly affected by oil contamination at low to
moderate concentrations (less than 5 mg crude oil/g sediment, Alexander and Webb, 1987; less than
50 mg crude oil/g sediment, DeLaune et a/., 1979) and sometimes was even stimulated (Li et a/.,
1990). However, heavy contamination by light oil can lead to widespread mortality, and plants may
require a decade or more to recover. Different wetland plants also respond differently to oil spills.
Lin and Mendelssohn (1996) examined the effects of south Louisiana crude oil on three different
types of coastal marshes and found that the sensitivity of these marshes to the crude oil increased in
the order of S. lancifolia (freshwater marsh plant), S. alterniflora (salt marsh), and S. patens
(brackish marsh). Plants are more sensitive to oiling during the growing season than other periods
(Pezeshki et a/., 2000). The sediment type also plays an important role. In general, oil remains
longer in soils with higher organic matter and, therefore, has greater impact on resident plants. Some
wetland sediment can act as a reservoir absorbing oil and leaching it out into adjacent coastal
habitats, causing chronic impacts on biota (Levings et a/., 1994).
1.2.1.2 Impact to wildlife and ecosystems
Oil spills on coastal wetlands not only damage plants but also have serious consequences for the
wildlife and other organisms that rely on the wetlands as habitats and nursery grounds. These
impacts include obvious immediate consequences, such as widespread animal mortality due to
smothering and toxic effects, and more subtle long-term effects. Oil can affect the fish population by
both direct toxicity and by a reduction in the benthic species on which they feed (NAS, 1985).
Seabirds that congregate on the salt marshes suffer from the destruction of their feeding grounds. Oil
can also change an animal's feeding and reproductive behaviors. A light oiling of some birds can
inhibit egg laying (Fingas, 2001). Furthermore, heavy mortality of seabirds is often observed
because oiling effectively diminishes the natural water-repellant and insulation value of feathers.
The extent of the impacts also depends on many factors, such as the life cycle and the life habit of
organisms, the time and season of oil spills, the type and amount of oil, and the duration of oil
exposure (NAS, 1985). Sediment feeders could be more vulnerable to oil than epibenthic filter
feeders. Larval fish are more vulnerable to oil than juveniles and adults. Avian mortality would be
exacerbated by a spill occurring during their feeding and nesting season.
Considering the different sensitivity of wetland species and populations to oil, spills can
significantly affect the overall balance of wetland ecosystems, especially if damage occurs to a
dominant species. On the other hand, some long-term studies have suggested that many oiled marine
wetlands could recover naturally after a long time (Baker, 1999; Hester & Mendelssohn, 2000; Sell
et a/., 1995). Recovery times vary from a few years for some salt marshes to over a decade for
mangroves (NAS, 1985; Sell etal., 1995). In a few extreme cases, salt marsh ecosystems have not
fully recovered decades after the initial oil spills (Baker etal., 1993; Teal etal., 1992). The effects
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of oil on wetland ecosystems and recovery still require further investigation.
1.2.2 Response to oil spills in salt marshes
Since oil spills can cause serious damage to marine wetland ecosystems, effective countermeasures
are essential to minimize these ecological impacts. Major oil spill response options in marine
shorelines and freshwater environments have been briefly reviewed in Chapter 1 of the sandy
shoreline and freshwater wetland guidance document (Zhu etal., 2001). However, saltmarshes are
among the most sensitive ecosystems and, therefore, the most difficult to clean. Applications of
some traditional oil spill cleanup techniques in wetland habitats have caused more damage than the
oil itself (Baker 1999; Owens and Foget, 1982, Sell etal., 1995). Considering the characteristics of
wetland ecosystems, a number of cleanup and treatment techniques have been proposed and tested to
deal with oil contamination in coastal wetlands. The feasibility of these methods also depends on
various factors, such as the type and amount of spilled oil, season of the year, and environmental
conditions of the spill site.
1.2.2.1 Physical Methods
Many oil spill countermeasures based on physical clean up procedures, such as mechanical oil
removal, high pressure or hot water flushing, and sediment relocation, have been reported to do
more harm than good to wetland habitats. All physical methods that remain as options for use on
marine wetland environments require some caution during deployment to minimize environmental
damage.
• Booming and sorbents - Use of booms to contain and control the movement of floating oil
at the edge of the wetland and removal of the oil by adsorption onto oleophilic materials
placed in the intertidal zone. This method can be an effective strategy to prevent floating oil
from reaching sensitive habitats with minimal physical disturbance if traffic of the cleanup
crew is strictly controlled.
• Low pressure flushing - Oil is flushed with ambient-seawater at pressures less than 200 kpa
or 50 psi to the water edge for removal (NOAA, 1992). This technique can be used
selectively for quick removal of localized heavy oiling with minimal damage to wetland
vegetation. However, the potential for oil release into the sediments and adjacent water
bodies should be considered including appropriate containment measures.
• Cutting vegetation— Cutting vegetation may be a useful cleanup technique to remove oils
that form a thick coating on the vegetation and to prevent oiling of sensitive wildlife (Baker,
1989, NOAA 1992). However, the feasibility of this method depends strongly on the season
in which the spill occurs. In general, winter cutting of dead standing vegetation has little
effect on subsequence growth, but summer cutting could cause great damage to the regrowth
of wetland plants and result in shoreline erosion. The use of cutting should also be avoided
immediately prior to an anticipated rise in water levels because cutting followed by flooding
could cut off necessary oxygen to plant roots (Pezeshki et a/., 2000). Efforts should also be
made to minimize the inevitable damage due to traffic.
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• Stripping - Stripping of surface sediments can cause severe environmental impacts and may
only be considered in the case of extremely oiled wetlands where the oil in the sediments is
likely to kill the vegetation and prevent plant regrowth. To minimize erosion and habitat
loss, it is critical to follow the stripping by the restoration of sediment elevation and
replanting of the wetland species (Krebs & Tanner, 1981).
1.2.2.2 Chemical methods
Chemical methods have not been widely used in the United States mainly due to the concerns over
their toxicity and long-term environmental impacts. However, with the development of less toxic
chemical agents, the potential for their application will increase.
• Dispersants - Dispersants are chemicals that promote the dispersion of floating oil from the
water surface into the water column. Fields studies have shown that application of
dispersants in near shore waters can significantly reduce the retention of oil within the
intertidal zone and, therefore, the impacts to wetland plants (Duke et al., 2000; Getter &
Ballou, 1985). However, the use of dispersants in near shore water could have short-term
toxic effects on adjacent coastal habitats, such as subtidal animal communities. Direct
spraying or contact of dispersants with wetland plants may also have harmful effects on
vegetation (Wardrop et al., 1987).
• Cleaners - Cleaners are chemicals that help wash oil from contaminated surfaces. These
formulations have been used with low-pressure flushing operations to facilitate oil removal
from wetland vegetation. Studies have shown that the application of cleaners can prevent
mortality of salt marshes and mangroves (Pezeshki etal., 1995; Teas etal., 1993). However,
their use has been limited because of the paucity of data available with respect to their long-
term effects on wetland habitats. Also, concern has been expressed over the transfer of oil to
the nearshore waters.
1.2.2.3 In-situ burning
In-situ burning involves controlled burning of the oil and oiled vegetation at the contaminated site.
This technique is capable of rapidly removing large amounts of oil with limited equipment and
personnel. However, the technique may result in severe damage to wetland habitats, temporary air
pollution, and possibly toxic combustion residues. The degree of impact to salt marshes is seasonally
dependent. Like cutting, the likelihood of damage is greatest during the summer and least during the
period of dormancy in late fall and winter (Baker, 1989). In fact, fall burning of marshes has been a
commonly used management strategy for controlling wetland overgrowth in many areas. The
temporary air pollution caused by the airborne emissions are generally not considered a serious
health threat or environmental concern, especially at distances greater than a few kilometers from the
fire (Fingas, 2001). Limited data and applications have indicated that in-situ burning can be a viable
option for removing a large volume of pooled oil at the right season when sediments are saturated
(Mendelssohn etal, 1995; Pahl etal, 1997).
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1.2.2.4 Restoration
In cases of coastal wetlands being catastrophically damaged, plant and animal species have been re-
introduced as restoration strategies. (Bergen etal., 2000; Frink and Gauvry, 1995; Teas etal., 1989).
S. alterniflora was successfully replanted to restore the salt marshes in New Jersey after the 1990
Arthur Kill oil spill (Bergen et a/., 2000). Three years after the replanting, over 70% of the plant
coverage was restored as compared to only 5% by natural recolonization at the unplanted reference
sites. Mangroves were also successfully replanted to restore oil-killed mangrove forest in Panama
after the 1986 Refineria Panama oil spill (Teased a/., 1989). However, this approach may also upset
the ecological balance or natural succession processes if it is not carried out appropriately (Fingas,
2001).
1.3 Bioremediation of Oil Spills in Salt Marshes
Bioremediation is an emerging technology that involves the addition of materials (e.g. nutrients or
other growth-limiting cosubstrates) to contaminated environments to accelerate the natural
biodegradation processes (OAT, 1991). This technology has been recognized as one of the least
intrusive methods and has been shown to be an effective tool for the treatment of oil spills in
medium and low-energy marine shorelines (Lee et al, 1997; Swannell et al., 1996; Venosa et al.,
1996, Zhu etal., 2001). However, until a few years ago, only limited information was available on
the effectiveness and impacts of the bioremediation of oil spills in coastal wetlands (Lee etal., 1991;
Wood et a/., 1997; Wright et a/., 1997). Recently, several long-term field studies on oil
bioremediation in coastal wetlands have been carried out. These studies provide better understanding
of the potential of oil remediation in such environments (Burns et a/., 2000; Garcia-Blanco and
Suidan, 2001; Jackson and Pardue, 1999; Shin et a/., 1999). In this section, an in-depth review of
current practices and research on oil bioremediation in coastal wetland environments is presented
with emphasis on the findings of these field trials.
1.3.1 Environmental factors affecting oil biodegradation in salt marshes
The success of oil spill bioremediation depends on our ability to establish and maintain conditions
that favor enhanced oil biodegradation rates in the contaminated environment. Environmental factors
affecting oil biodegradation include temperature, nutrients, oxygen, pH, and salinity. These factors
have been discussed in general in Chapter 2 of the sandy shoreline and freshwater wetland guidance
document (Zhu etal., 2001) and in Section 5.5 of the document with respect to freshwater wetlands.
The limiting conditions for oil biodegradation in salt marshes can be significantly different from
other marine shorelines and even freshwater wetlands. A brief summary of these conditions in salt
marshes is given here.
In terms of nutrient supply, coastal marshes are considered high-nutrient wetlands (Mitsch and
Gosselink, 2000), but most of the nutrients, and nitrogen in particular, are present in the form of
organic matter and not readily available for microbial or plant uptake (Cartaxana et a/., 1999).
Figure 1.1 illustrates the nitrogen cycling that occurs in a wetland environment. The amount of
inorganic nitrogen or available nitrogen for oil biodegradation will depend on many processes, such
as nitrogen fixation, nutrient mineralization, plant uptake and release, denitrification, and wetland
hydrodynamics. Studies also show that the concentration of inorganic nitrogen (mostly ammonium)
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in salt marsh sediments exhibits a seasonal pattern with a concentration peak during the summer
months probably due to higher mineralization rates associated with elevated temperatures (Cartaxana
et a/., 1999). A similar trend was also reported for available phosphorus (Nixon et a/., 1980).
Therefore, when a major oil spill occurs in salt marshes, it is still likely that nutrient availability
becomes a limiting factor for oil degradation, depending on the type of sediment, the season, and
quantity of oil spilled. Hence, nutrient addition may be an important remedy to contemplate when
considering any or all of these factors.
N inflows via
Surface water"
groundwater
tidal exchange
Aerobic
soil layer
Anaerobic
soil layer
N 9 Fixation
N9 & N9O
V
N outflow via
Tidal exchange
Algae & Bacteria
Organic N
NFL
Mineralization Nitrification
Plant Release NO,- ^ Denit
Plant uptake
ification
Figure 1.1 Major processes involved in nitrogen cycles in a coastal wetland
Unlike other marine shorelines, the substrates of coastal wetlands are saturated or flooded with
water, and oxygen diffusion rates through these hydric soils are very slow. As a result, available
oxygen in the soils and in the interstitial water is quickly depleted through metabolism by aerobic
organisms and chemical oxygen demand due to reduced chemical species. Typically, there is a thin
layer of oxidized soil (a few millimeters) at the surface, below which the environment becomes
anaerobic (Gambrell and Patrick, 1978). The thickness of this oxidized layer depends on the
population of oxygen utilizers, the rate of photosynthetic oxygen production by algae and plants, and
the rate of oxygen transport through the sediments, which is related to wind, tide and wave action. A
study of oxygen demand in an oil contaminated salt marsh sediment indicated that oxygen
availability could be a limiting factor for oil degradation (Shin etal., 2000). These authors reported
that significant biodegradation occurred only when the tidal cycle exposed the surface of the salt
marsh to the atmosphere. The dominant electron acceptor in the anaerobic soil layer in salt marshes
is also different from most freshwater wetlands. In freshwater wetlands, methanogenesis is often the
dominant process for the oxidation of organic carbon in the reduced soil layer, while in marine
wetlands, sulfate reduction is usually the most important process when oxygen is limited since
seawater contains abundant sulfate (Mitsch and Gosselink, 2000). Studies have shown that in some
marine sediment, PAHs and alkanes can be degraded under sulfate-reducing conditions at similar
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rates to those under aerobic conditions (Caldwell etal., 1998; Coates etal., 1997). The importance
of this process in the biodegradation of oil in salt marsh environments still requires further
demonstration, especially in the field. The inherent ability of many wetland plant species to transfer
oxygen to the rhizosphere may also play a role in reducing the effect of oxygen limitation. However,
little research has been conducted on the capacity of this mechanism in enhancing oil
biodegradation.
Other important environmental factors affecting biodegradation of petroleum hydrocarbons include
pH and salinity. The optimal pH for oil biodegradation is between 6 and 9 (Atlas and Bartha, 1992).
The pH of wetland sediments and the overlying water depend on both soil type and hydraulic
condition. Sediments in salt marshes and mangroves are mostly organic and often acidic. In addition,
the anoxic condition and high sulfur content in coastal wetlands often lead to the production of
hydrogen sulfide. When exposed to air, sulfide can be reoxidized to sulfate and result in a further
drop in sediment pH. However, in areas with frequent tidal inundation, the pH of wetland sediments
and pore water is determined by seawater and is near neutral or slightly alkaline. The salinity of pore
water in coastal wetlands may also vary dramatically and depends on the frequency of tidal
inundation, rainfall, coverage of vegetation, groundwater and freshwater inflow, and soil type
(Mitsch and Gosselink, 2000). In areas adjacent to coastlines or receiving frequent tidal inundation,
the salinity of wetland water is close to that of seawater. Elevated salinity levels are frequently
reported at higher elevations or areas with little rainfall and other freshwater supply. Brackish
marshes are frequently found in estuarine areas and other sites that extend away from the open
ocean. Salinity can significantly affect the rates of oil biodegradation. Most marine microorganisms
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). Studies have also shown the rates of hydrocarbon degradation to decrease with
increased salinity above that of seawater (Rhykerd et a/., 1995; Ward and Brock, 1978).
Although many factors can affect oil biodegradation, not many environmental factors can be easily
manipulated to enhance this natural process. For example, it is not practical to alter wetland salinity,
and nothing can be done to change the climate. There are two main approaches to oil spill
bioremediation. (1) bioaugmentation, in which oil-degrading microorganisms are added to
supplement or augment 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
co-substrates. Extensive studies have been carried out recently on the bioremediation of oil
contaminated coastal wetlands both on a laboratory scale and in the field (see next section).
1.3.2 Laboratory studies
Most of the laboratory studies have focused on the potential of using nutrient amendments to
enhance oil biodegradation in salt marsh environments. This is because studies conducted in other
shoreline environments have demonstrated that the microbial population is rarely a limiting factor,
and nutrient addition alone had a greater effect on oil biodegradation than did the addition of
microbial products (Leeetal, 1997; Venosae^a/., 1996).
Jackson and Pardue (1999) conducted microcosm and mesocosm studies to investigate the effect of
different nutrient types on enhancing biodegradation of a Louisiana crude oil in Louisiana salt marsh
sediment. The microcosms contained a 60:1 (water/soil) slurry produced from a salt marsh sediment
-------
at an oil concentration of 0.7 g oil/g soil and were operated in a completely mixed and aerated mode,
where oxygen limitation was non-existent. Nutrient species examined included phosphate,
ammonium, nitrate, and phosphate plus ammonium. The results showed that oil degradation was
limited by nitrogen but not phosphorus under these conditions. Optimal nitrogen concentrations in
pore water were in the range of 100-670 mg N/L. Among the nitrogen species, ammonium was
found to be generally more effective in stimulating oil degradation than nitrate. Thus, ammonium
might be advantageous in the enhancement of the degradation of certain oil components because
ammonium is less likely to be lost from the system by washout due to its higher adsorptive capacity
to organic matter. Ammonium is also the more toxic species of nitrogen in the environment.
In a follow-on mesocosm study in the same paper (Jackson and Pardue, 1999), large intact cores
(900 cm2) of salt marsh sediments were contaminated with crude oil and treated with various
concentrations of ammonium salts. The results showed that ammonium amendments had limited
success in enhancing oil biodegradation. Even at the highest ammonium loading (10 mg/m2), the
nutrient amendment was only able to increase the degradation of lower chain length alkanes (
-------
supplemented with sodium tripolyphosphate and was added to microcosms at a N:P ratio of 5:1. The
study showed that the addition of nutrients did not enhance the rate of degradation over the natural
attenuation rate. The extent of microbial degradation of No. 2 fuel oil in all the microcosms averaged
only 20% for the total aliphatic hydrocarbons and 12% for the total PAHs. Degradation was greater
in all cases in the top layers than in the bottom layers of the columns, suggesting that oil degradation
may have been limited by oxygen availability under the conditions of this study.
Because these laboratory studies seem to suggest that adding nutrients may be effective under non-
oxygen limiting conditions, or during certain seasons, further field experiments are necessary to
determine the potential of oil bioremediation in coastal wetlands.
1.3.3 Full-scale demonstrations
From north temperate salt marshes to tropical mangroves, several field studies on the performance of
oil bioremediation have been carried out in recent years. These have provided more convincing
demonstrations of the effectiveness of oil bioremediation since laboratory studies may not consider
many real world conditions such as spatial heterogeneity, biological interactions, and mass transfer
limitations.
1.3.3.1 Nova Scotia, Canada, 1989
Lee and Levy (1991) conducted one of the first field trials on oil bioremediation in a salt marsh
environment. The study involved periodic addition of water-soluble fertilizer granules (ammonium
nitrate and triple super phosphate) to enhance biodegradation of waxy crude oil in a salt marsh
dominated by Spartina alterniflora and located in Nova Scotia, Canada (Lee and Levy, 1991). Two
levels of oil concentrations were used (0.3 and 3.0% v oil/v sediment) and two concentrations of the
NH4NO3 were tested (0.34 and 1.36 g/L sediment). In this study, pristane was used as a biomarker
for evaluation of biodegradation of crude oils. Results showed that the effectiveness of nutrient
addition was related to oil concentration. Enhancement by fertilizer was significant at the 0.3%
contamination level, but no enhancement occurred at 3%, which was attributed to the penetration of
the oil at higher concentrations into the reduced soil layers where little degradation is expected. This
study indicated that bioremediation might have a role in the cleanup of coastal wetlands lightly
contaminated with oil.
1.3.3.2 San Jacinto Wetland Research Facility (SJWRF), Texas, 1994-1997
To evaluate the effectiveness of various bioremediation options, a series of field trials were carried
out in a Texas coastal wetland by a research group from Texas A&M University (Mills et a/., 1997;
Mills et a/., 2003; Simon et a/., 1999; Townsend et a/., 1999; Mills et a/., 2003). This brackish
wetland was set aside for a long-term research program after an oil spill from ruptured pipelines in
1994. The 21-plot site, named San Jacinto Wetland Research Facility (SJWRF), has been used for a
series of studies on oil spill countermeasures. Studies on oil bioremediation included three phases.
Phase I of the research evaluated the intrinsic bioremediation or natural attenuation process after the
initial oil spill. The effect of biostimulation was investigated in phase II by evaluating the use of
diammonium phosphate and diammonium phosphate plus nitrate. Phase III involved the evaluation
of two commercial bioaugmentation products and a repeated diammonium phosphate treatment. The
10
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21 5 x 5 m plots were arranged in a randomized complete block experimental design, and Arabian
light and medium crude oils were used in phases II and III, respectively. Oil constituents were
determined using gas chromatography/mass spectroscopy (GC/MS) and were normalized to 17oc(H),
21(3(H)-hopane to reduce the effects of sample heterogeneity and physical losses. The results of
phase II showed that the diammonium phosphate treatment significantly enhanced the
biodegradation rates of both total resolved saturates and total resolved PAHs, and the addition of
diammonium phosphate plus nitrate only enhanced the biodegradation of total resolved saturates
(Mills et a/., 1997). The field trial on oil bioaugmentation in phase III showed as with other
shoreline types (Leeetal, 1997a; Venosae^a/., 1996), that the addition of microbial products does
not significantly enhance oil biodegradation rates (Simon etal., 1999). However, the performance of
the nutrient treatment in this phase also failed to demonstrate the enhancement observed in Phase II.
1.3.3.3 Terrebonne Parish, Louisiana, 1998
Due to the mixed results from the earlier trials, field studies were conducted to verify the feasibility
of oil bioremediation and to determine the limiting factors in oil biodegradation in coastal wetland
environments. Shin et al. (1999 & 2000) investigated the effect of nutrient amendment on the
biodegradation of a Louisiana "sweet" crude oil and oxygen dynamics in a Louisiana salt marsh,
which has a tidal range of 20 cm and is vegetated by Spartina alterniflora. Four treatments (unoiled
control, oiled control, oil plus ammonium nitrate, and oil plus a slow release fertilizer) were
examined in forty field plots arranged in a randomized complete block design. Oil components were
measured by GC/MS, and hopane was used as a biomarker. Oxygen dynamics were investigated by
monitoring sediment oxygen demand (SOD), and the importance of sulfate reduction was
determined using a 35SO42" radiotracer technique. Overall, the nutrient amendments did not
significantly stimulate oil biodegradation, which might have been related to the high background
nutrient concentrations at this site. Throughout this study, the background nitrogen concentrations in
the interstitial pore water were higher than the threshold nitrogen concentration of 1 - 2 mg N/L
required for maximum hydrocarbon biodegradation as found by Venosa etal. (1996) in an unrelated
field trial on a sandy beach. The addition of oil and fertilizers did increase the SOD and sulfate
reduction rates in marsh soils. About 2/3 of the oxygen demand was due to aerobic respiration with
the majority of this demand exerted by hydrocarbon degrading organisms, indicating aerobic
biodegradation of the crude oil was the main mechanism. The remaining 1/3 of the oxygen demand
was attributed to sulfide oxidation. Data also showed that significant biodegradation of crude oil in
the salt marshes occurred only when the tidal cycle exposed the surface of the marsh to air (Shin et
a/., 2000). This study indicated that oxygen availability appears to control the oil biodegradation
process in salt marshes.
1.3.3.4 Gladstone, Australia, 1997-1998
A field study on the performance of oil bioremediation in both mangrove and salt marsh ecosystems
was carried out recently in a tropical marine wetland located at Gladstone, Australia (Burns et al.,
2000; Duke et al., 2000; Ramsay et a/., 2000). This study evaluated the influence of a
bioremediation protocol on the degradation rate of a medium range crude oil (Gippsland) and a
Bunker C oil stranded in a tropical Rhizophora sp. mangrove environment and in Haloscarcia sp.
salt marshes behind the mangroves. The bioremediation strategy used in this study involved
pumping air beneath sediment that was supplemented with a slow release fertilizer (Osmocote™
11
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Tropical fertilizer) for the mangrove sites, and nutrient addition alone for the salt marsh sites. No
aeration was tested in the salt marsh experiments because the sediment of the salt marshes was much
less anoxic than that of the mangroves based on preliminary investigations. Four oiled treatments
(two types of oils with and without the bioremediation treatments) and two unoiled controls
(enclosure and ambient controls) were tested. Each treatment was studied with replicates of three
plots for the mangrove sites and replicates of four plots for the salt marsh sites. The oils were added
to mangrove plots and salt marsh plots at target loadings of 5 L/m2 and 2 L/m2, respectively. The
fertilizer was added at a loading of 0.15 kg/m2 40 hours after oiling for both mangrove and salt
marsh plots and then again after three months in mangrove plots only. Aeration in mangrove plots
started 40 hr after oiling and lasted for about four months. In this study, other than total
hydrocarbons (THCs), only individual alkanes were analyzed using GC-FID, and phytane was used
as a biomarker. Oil analysis for the mangrove sediments over 13 months showed that no significant
change in oil composition due to biodegradation was observed until two months after oiling, and by
that time 90% of the THCs were removed from the sediments through evaporation and dissolution.
The remediation strategy did not significantly enhance the degradation of either the remaining
Gippsland oil or the Bunker C oil. A similar lag phase before the start of oil biodegradation was
observed in the 9-month salt marsh experiment. The addition of the fertilizer to the salt marshes did
show a stimulation of the degradation of the lighter Gippsland oil and resulted in about 20% more oil
loss as compared with the untreated plots. However, the nutrient amendment did not significantly
impact the rate of loss of Bunker C oil in the salt marsh plots. Microbial analysis for the mangrove
sediments showed that the bioremediation treatment had a significant effect on alkane degraders and
increased the population size by one to three orders of magnitude, as compared to the oil only plots.
However, the population of aromatic degraders only increased slightly (one order of magnitude).
Due to the limitation of the experimental design, the study could not distinguish whether nutrient
addition or aeration stimulated the microbial growth.
In the same study, Duke et al. (2000) investigated the ecological effects of the bioremediation
strategy in the mangroves and compared the results to a previous field trial involving use of a
dispersant at the same site. Although the authors suggested that the dispersant (but not
bioremediation) significantly reduced the mortality of mangrove trees, the data appeared to show
that both treatments had some positive effects on the wetland habitats. The increase in the tree
mortality in bioremediation plots occurred only months after aeration and nutrient addition was
stopped. Even though some aspects of the design are questionable (e.g., lack of independent tests of
the effect of nutrients and aeration in mangrove environments, different oiling and treatment
conditions for the salt marsh and mangrove experiments), the Gladstone field trial did provided
useful insights on the potential of oil bioremediation, particularly in tropical marine wetland
environments.
1.3.3.5 Nova Scotia, Canada, 2000-2001
A comprehensive field trial conducted on oil bioremediation in a salt marsh environment was carried
out recently by the U.S. Environmental Protection Agency, University of Cincinnati, and Fisheries
and Oceans Canada in a coastal salt marsh site situated on the Eastern Shore of Nova Scotia, Canada
(manuscript not published at the time of this writing). This study explored various options for
restoring salt marshes heavily contaminated with petroleum hydrocarbons under north-temperate
conditions. Treatment options included natural attenuation, phytoremediation, and/or bioremediation
12
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by nutrient amendment and disking (gentle tilling). Like most North American salt marshes, this
tidal salt marsh was dominated by Spartina alterniflora. Tides were semi-diurnal with a range of
about 2 m. The commencement of the experiment coincided with the spring tide. Test plots for the
study were set up throughout the wetland in a way that all of them were exposed to the same tidal
inundation.
A randomized block design was used in the study. Eighteen 3 m x 3 m plots were set up in three
replicate blocks. Each block contained six treatments randomly distributed: (a) unoiled, no-nutrient
control; (b) unoiled with nutrient amendments (NfLJSTOs + CatfLiPO^iHiO}; (c) oiled with no
nutrient amendments and plants intact (naturalattenuation); d) oiled with nutrient amendments and
plants intact; (e) oiled with nutrient amendments and vegetation continually cut back to the ground
surface and removed to suppress the influence of plants and anaerobiosis associated with the
accumulation of plant detritus; (f) oiled with nutrient amendments and disked daily to introduce
oxygen into the rhizosphere.
A weathered Mesa crude oil was applied to the plots at a rate of 35 mg oil/g dry sediment during the
first two days of the study. Granular nitrogen and phosphorus nutrients were initially added to each
of the treated plots at a dosage of 450 g-N and 135 g-P per plot. Subsequent applications took place
on an as-needed basis as determined by residual nutrient analysis in the interstitial pore spaces.
When the nitrogen levels fell below a specified concentration range of 5-10 mg N/L in the pore
water, another application of the same magnitude was made. The effectiveness of various treatments
was determined by monitoring the reduction of oil constituents in both soil and grass samples using
GC-MS techniques. Hopane was used as a biomarker to reduce the effects of sample heterogeneity
and to distinguish bioremediation removal from physical losses. In addition to these detailed
chemical analyses, this project also used biological endpoints such as evidence of wetland plant
recovery and reduction of toxic responses to verify the success of the
bioremediation/phytoremediation treatments (discussed in Section 2.2.4).
The study showed that the biodegradation of targeted aliphatic hydrocarbons and PAHs took place to
a very high extent at this north-temperate salt marsh. After 20 weeks, the extent of degradation of
target n-alkanes within the experimental plots averaged 87% and 97 %, respectively, for the oil in
sediment and the oil associated with emergent vegetative growth. Reduction of parent and alkyl-
substituted PAHs was about 69% in the soil samples and 88% in the plant samples. However,
targeted alkanes and PAHs only represent a small fraction (less than 10%) of the total petroleum
hydrocarbons (TPHs). Biodegradation of TPH averaged only 3 5% in the soil samples and 42% in the
grass samples (very little of this TPH was comprised of high molecular weight plant waxes). More
than half of the applied oil remained in the marsh 20 weeks after oil application. Based on the extent
of oil washout (measured as mg hopane per kg dry soil) and the total oil loss (g TPH per kg dry soil),
the main mechanism for oil disappearance was attributed to biodegradation. These results contrast
with those reported in the St. Lawrence River freshwater wetland bioremediation study (Venosa et
a/., 2002) and the study on a tropical marine wetland at Gladstone, Australia, in which most of the
TPH was removed through physical mechanisms (Burns et al., 2000).
In this study, as in other reported field studies, no significant differences were observed among
treatments, either in the degradation of alkanes or PAHs. No significant enhancement of
biodegradation through the addition of nutrients or the use of disking was observed. The average
13
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nutrient concentrations in the interstitial pore water for various treatments during this study are
summarized in Table 1.1. Data indicated that the background nitrogen (mostly ammonium nitrogen)
and phosphorus concentrations in the interstitial pore water were always far in excess of 2-5 mg N/L
and 1 mg P/L, suggesting that nutrients may not be a limiting factor for oil biodegradation in this salt
marsh. Enhanced oxygen transfer to the rhizosphere by the plants through their roots or by disking
did not appear to take place either, at least to the level needed by hydrocarbon degraders to
metabolize the oil rapidly. From the extent of degradation of the targeted aliphatic and aromatic
hydrocarbons, it can be inferred that there was no oxygen or nutrient limitation in this particular salt
marsh site.
Table 1.1 Average nutrient concentrations in pore water for different treatments during
Nova Scotia study
NH4-N
mgN/L
NO3-N
mgN/L
PO4-P
mgP/L
Treatment A
Background
9.43
0.09
1.25
Treatment
B
Nutrient
Control
60.63
28.57
12.42
Treatment
C
Natural
Attenuation
18.08
0.07
2.17
Treatment D
Phyto-
remediation
92.49
37.11
5.66
Treatment E
Nutrient
Amendment
80.83
30.56
9.42
Treatment F
Disking
104.6
28.89
14.75
However, nutrient addition did stimulate microbial growth, as in the Gladstone study (Ramsay etal.,
2000). Alkane degraders in this wetland seemed to be nutrient limited, since the addition of nutrients
without oil led to an increase in number of about two orders of magnitude relative to background
levels. However, when oil was added to the plots without any nutrient amendment (natural
attenuation plots), the increase in alkane degraders was also on the same order, suggesting the
existence of different populations that can degrade alkanes under different conditions. However,
PAH degraders were clearly limited by their carbon sources. Only oiled plots showed an increase in
the number of PAH degraders. These populations did not seem to be limited by nutrients since the
addition of nitrogen and phosphorus did not have an effect on either the number of microorganisms
or on the rate of PAH degradation.
In summary, these field trials suggest that nutrient amendments may be less effective in stimulating
oil biodegradation rates in coastal wetlands than sandy beaches. Oil biodegradation on marine
wetlands is often limited more by oxygen than by nutrient availability. A large fraction of the total
nutrients in wetland sediments is bound in organic matter (i.e., plants and detritus) that is not readily
available for microbial uptake. In such cases, natural ongoing mineralization processes may provide
an effective means to overcome this restraint. However, field experience also suggests that some
coastal wetlands may be nutrient limited, and in these cases, biostimulation with nutrient enrichment
may still be an appropriate countermeasure treatment if the oil does not penetrate deeply into the
anaerobic layer of the wetland sediments (Lee and Levy, 1991; Venosa et a/., 2002; Mills et a/,
1997).
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1.3.4 Kinetics of oil biodegradation
Knowledge of the kinetics of oil biodegradation is important for assessing the potential fate of
targeted compounds, evaluating the efficacy of bioremediation, and determining appropriate
strategies to enhance oil biodegradation. The rates of biodegradation vary greatly among the various
components of crude oils and petroleum products and depend on many environmental factors, such
as temperature, nutrient concentration, and oxygen content. The heterogeneity of oil distribution on
shorelines or wetland sediments makes kinetics studies even more difficult. To reduce the variability
associated with heterogeneous oil distribution, Venosa etal. (1996) utilized hopane normalization in
studying the kinetics of oil biodegradation and developed first-order biodegradation rate constants
for resolvable alkanes and important two- and three-ring PAH groups present in a light crude oil on a
sandy beach in the Delaware field study. The first order relationship was expressed as:
(1.1)
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 analyte, A.
Tablel.2 Summary of first order biodegradation rate constants from field studies
Field
Study
Location
Delaware
Quebec
Canada
Texas
Louisiana
Nova
Scotia
Canada
Shoreline
Type
Sandy
beach
Tidal
freshwater
wetland
Brackish
wetland
Salt marsh
Salt marsh
Oil Type
Bonny light
crude oil
Mesa light crude
oil
Phase II:
Arabian light
crude oil
Phase III:
Arabian medium
Crude oil
South Louisiana
crude oil
Mesa light crude
oil
Treatment
Control
Nutrient
Inoculum
Control
Nutrient
Control
Nutrient
Control
Nutrient
Inoculum
Control
Nutrient
Control
Nutrient
First order biodegradation rate
day'1
Alkanes
0.026
0.056
0.045
0.0028
0.0023-0.0034
0.019
0.042-0.061
0.020
0.024
0.019-0.030
0.005
0.005
0.020
0.026-0.039
PAHs
0.021
0.031
0.026
0.0028
0.0016-
0.0041
0.017
0.018-0.027
0.015
0.013
0.016-0.017
N/A
0.010
0.011-0.013
Reference
Venosa et
al., 1996
Venosa et
al, 2002
Simon et al.,
1999
Shin et al.,
1999
Unpublished
data
Since the Delaware study, several field trials conducted in other types of environments, including
salt marshes and a freshwater wetland, have reported first-order oil biodegradation rate constants
obtained using the same approach (Shin et al., 1999; Simon et al., 1999; Venosa et al., 2002). The
results of these kinetic studies in the field are summarized in Table 1.2. It can be seen that except for
15
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Phase II of the San Jacinto field study (Mills etal., 1997; Simon etal., 1999), nutrient amendments
did not show significant biodegradation rate enhancements in most of these wetland environments.
Nonetheless, it is encouraging to see that crude oils can be biodegraded intrinsically on marine
wetlands at similar rates as on sandy beaches (i.e., the Delaware site). The higher intrinsic oil
biodegradation rates reported for salt marshes when compared to freshwater wetlands may be
attributed to the generally greater oxygen limitation in freshwater wetland environments (Mitsch and
Gosselink, 2000). It should be noted here that in the St. Lawrence River freshwater wetland study
(Venosa et al., 2002), the oil was manually raked into the sediment, causing penetration of oil into
the anaerobic zone. It is not known whether higher biodegradation rates would have ensued had this
raking not been done.
1.3.5 Monitoring biological responses to quantify the efficacy of remediation treatment
In addition to the demonstration that remedial treatments reduce the concentration of residual oil, it
is necessary to demonstrate that they do not produce any undesired environmental and ecological
effects. As discussed in the sandy shoreline and freshwater wetland guidance document (Zhu etal.,
2001), two complimentary approaches are available: (1) bioassessments, which typically monitor
changes in populations and communities of flora and fauna (Herri cks and Schaeffer, 1984); and (2)
bioassays, which include toxicity tests and bioaccumulation studies (Chapman, 1989).
1.3.5.1 Bioassessment
The monitoring of alterations in benthic community structure is frequently used to assess the
potential impacts of residual oil within sediments. For example, in a follow up of the Exxon Valdez
oil spill clean up, Driskell etal. (1996) noted negative effects including reductions in size, biomass,
fecundity, and increased mortality as a result of hot water washing. 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 Persian Gulf oil spill (Watt et al., 1993). The
possibility of adverse ecological effects such as algal blooms and invertebrate mortality from
excessive nutrient amendments associated with bioremediation treatments is also a concern (Lee,
2000a; Lee et al, 2001a, 2001b).
To date, sediment bioassessments have been largely based on the tracking of changes in
macroinvertebrate community structure. For a holistic approach, it is recommended that
consideration should also be given to the bioassessment offish and other non-benthic community
organisms (e.g. bacteria, phytoplankton, cladocera, and amphibians). Furthermore, with recent
advances in biotechnology, micro-scale bioassays are now available to monitor alterations at the
subcellular or multicellular level of biological organization (Lee etal., 1998; Wells etal., 1998). In
wetland environments, quantification of potential impacts on vegetative growth can be used to
document the efficacy of bioremediation strategies. For example, in a tidal freshwater wetland
experiment, the predominant plant species (Scirpuspungens) was reported to be tolerant to the oil,
and its growth was significantly enhanced above that of the unoiled control by the addition of
nutrients (Lee et al., 200la). Monitoring of recolonization within impacted areas should be
considered as an endpoint in bioassessments, as it provides integrated information on the impact of
contaminants on processes such as immigration, emigration, competition, and predation.
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1.3.5.2 Bioassays
As discussed in the sandy shoreline and freshwater wetland guidance document (Zhu et a/., 2001),
acute and/or chronic bioassays can be performed on whole sediment (e.g., solid-phase), suspended
sediment, sediment liquid phases (pore water, interstitial water), or sediment extracts (elutriates,
solvent extracts). Since various forms of biota differ in their sensitivity to toxicants, it is highly
recommended that a test battery approach with species from different trophic levels be utilized in
environmental assessments to ensure ecological relevance.
Sediment bioassays have been used extensively to diagnose the effect of oil spills (Teal etal., 1992;
Gilfillanefor/., 1995; Neff and Stubblefield, 1995; Randolph etal., 1998) and their counter-measures
such as bioremediation (Lee et al., 1995b; Mearns et al., 1995; Mueller et al., 1999). Criteria to
consider for the selection of bioassays include: (1) sensitivity to test material, (2) ecological and/or
economic relevance, and (3) the availability of regional expertise for the analysis and interpretation
of results.
1.3.5.2.1 Numerous bioassays can be used to document the impact of oil spills in coastal
environments. Benthic invertebrates such as amphipods and shellfish have been found to be highly
sensitive to residual hydrocarbons following oil spill incidents (Teal et al. 1992; Gilfillan et al.,
1995; Mueller at al., 1999; Wolfe et al., 1996). They have been used in both field and laboratory
studies to monitor the impact and effectiveness of oil spill countermeasures such as bioremediation
(Mearns et al., 1997, 1995; Lee et al., 2001a,b). In terms of quantifying a microbial response, the
Microtox test is based on the measurement of changes in light emission by a nonpathogenic,
bioluminescent marine bacterium (Vibriofisherf) upon exposure to test samples. This commercial
assay has been used by regulatory agencies for toxicity screening of chemicals, effluents, water and
sediment, and for contamination surveys and environmental risk assessment, and its application for
monitoring the efficacy of oil spill remediation methods has been proven (Lee et al., 1995b, 1997;
Mueller etal, 1999).
1.3.5.2.2 Due to their economic, recreational, and aesthetic value, fish have been historically
selected as a primary bioassay organism. Biochemical and physiological alterations induced by
toxicant exposures can result in: 1) anatomical changes, 2) structural alterations in organelles, cells,
tissue, and organs, and 3) alteration of metabolic processes. For example, the observation of
neoplasms in fish was one of the first histopathological indices used in ecotoxicology. Biomarkers,
defined as biochemical, physiological, or pathological responses measured in individual organisms
on exposure to environmental contaminants, such as mixed function oxidase (MFO) reactions (Ortiz
de Montellano, 1986) are also used. MFO reactions induced by PAHs and a variety of halogenated
hydrocarbons are highly sensitive to contaminants. In the tidal freshwater wetland study with early
life stages offish, Hodson et al. (2001) noted that oil alone, oil mixed with sediments in the lab, and
oiled sediments from the experimental plots all caused induction of MFO (CYP1 A) enzyme activity
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 treatments.
To date, detrimental effects from nutrient enrichment have not been observed following full-scale
17
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field operations (Prince, 1993; Mearns etal., 1997). However, field experiments have suggested that
the possibility of detrimental effects from bioremediation treatments cannot be fully discounted
(Mueller etal., 1999). For 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 a/., 1995a). 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 a/., 1997). Bioassays used to document the effectiveness of bioremediation
treatments in sandy intertidal shoreline sediments oiled with a weathered light crude oil showed an
inhibitory effect on the hatching of grass shrimp due to the addition of nutrients (Mearns et a/.,
1995). Furthermore, most recently, in the tidal freshwater study described in the sandy shoreline and
freshwater wetland guidance document (Zhu et al. , 200 1), it was noted that amphipod toxicity levels
became elevated during the study due to excessive nutrient enrichment (Lee et a/., 2001a). It is
recommended that future operational guidelines include ecotoxicological-monitoring protocols.
1.3.6 Bioremediation options on salt marshes
Major bioremediation options have been described in the sandy shoreline and freshwater wetland
guidance document (Zhu et al., 2001). This document provides a summary of the most current
information on restorative techniques pertaining to salt marshes.
1.3.6.1 Nutrient Amendment
As stated in previous sections, biostimulation has been ineffective in accelerating the disappearance
of oil on certain oil-contaminated salt marshes (Garcia-Blanco and Suidan, 2001; Shin etal., 1999)
due to either the presence of high background nutrient concentrations or oxygen limitation.
However, a few field studies did show enhanced oil biodegradation through nutrient addition (Lee
and Levy, 1991; Mills et a/., 1997); therefore, nutrient amendment may still be a viable option for
removing hydrocarbons from an oil-contaminated wetland when nutrients are limiting. Nutrients
used for biostimulation can be classified as water-soluble, slow-release, oleophilic, and organic.
• Water-soluble nutrients — Commonly used water-soluble nutrient products include mineral
nutrient salts (e.g. KNO3, NaNO3, NH3NO3, K2HPO4, MgNH4PO4), 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. 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. The main disadvantage is that they are more likely to be washed away by tidal
and wave action. However, this washout effect is of lesser concern in salt marshes, since
they generally represent low-energy environments that are subj ect to little turbulent mixing.
A field study on nutrient hydrodynamics showed that water-soluble nutrients could remain in
contact with oiled sediments for weeks on low energy shorelines before being washed out
^a/., 1997a; Harris et al., 1999).
• Slow-release fertilizers — Slow release fertilizers are normally available in solid forms that
consist of inorganic nutrients coated with hydrophobic materials like paraffin or vegetable
18
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oils or organic nutrients encapsulated by semi-permeable or controlled-rate degradable
surface coatings. They are designed to overcome the washout problems and provide a
continuous supply of nutrients to oil contaminated areas. This approach may also cost less
than adding water-soluble nutrients due to less frequent applications (Lee et al., 1993). The
Gladstone field trial has shown promise for the application of slow-release fertilizers in
coastal wetlands (Burns etal, 2000). In this study, the degradation of a Gippsland crude oil
in salt marsh plots was stimulated by the addition of Osmocote™, a slow release fertilizer
consisting of a mixture of inorganic nutrients coated with an organic resin.
• Oleophilic nutrients - Another approach to overcome the problem of water-soluble nutrient
washout is to utilize oleophilic organic nutrients. The rationale for this strategy is that oil
biodegradation mainly occurs at the oil-water interface, and 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.
Results have been mixed. Some studies have suggested that oleophilic fertilizers might be
more suitable for use in high-energy, coarse-grained environments due to poor penetration of
fine sediments by oleophilic fertilizers (Sveum et al., 1994; Sveum and Ladousse, 1989).
Bioremediation agents containing organic substrates such as meat and fish-bone meal and
yeast extracts may have the capacity to provide essential micro-nutrients and organic growth
substrates that may be limiting. However, the large amount of organic carbon within this
type of amendment may also cause problems. For example, the organic carbon in the product
may be biodegraded by microorganisms preferentially over petroleum hydrocarbons, thus
contributing to oxygen depletion and resulting in undesirable anoxic conditions (Lee and
Levy, 1987,1989; Lee etal, 1995a,b; Swannelle^a/., 1996). Considering their high cost and
lack of demonstrated effectiveness, oleophilic fertilizers are unlikely to be the choice
biostimulation agent for oil cleanup in coastal wetlands.
1.3.6.2 Microbial amendments
Addition of oil-degrading microorganisms (bioaugmentation) has been proposed as another type of
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 (Leahy and Colwell, 1990). Although many vendors of microbial agents
claim that their product aids the oil biodegradation process based on laboratory tests, the
effectiveness of microbial amendments has not been convincingly demonstrated in the field (Zhu et
al., 2001). Actually, results from most field studies indicate that bioaugmentation is not effective in
enhancing oil biodegradation on marine shores. Field studies conducted on sandy beaches have
shown that nutrient addition or biostimulation alone had a greater effect on oil biodegradation than
microbial seeding (Lee and Levy, 1987; Lee et al., 1997, Venosa et al., 1996). The San Jacinto
study, the only reported field trial on oil bioaugmentation in a coastal wetland environment, also
revealed that addition of microbial products did not significantly enhance oil biodegradation rates
(Simon et al., 1999). This is because hydrocarbon-degrading microorganisms are ubiquitous in the
environment, and their density can increase by many orders of magnitude after exposure to crude oil,
as evidenced in recent studies of coastal wetlands (Garcia-Blanco and Suidan, 2001; Ramsay etal.,
2000; Townsend etal., 1999). Also, added bacteria may not be able to compete with the indigenous,
19
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well-adapted population (Lee and Levy, 1989; Venosa etal., 1992). The mass of the hydrocarbon-
degrading bacterial population on coastal wetlands is also limited 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 tidal activity. Therefore, it is unlikely that exogenous
microorganisms would persist in contaminated wetlands even when they are added in large numbers.
As a result, microbial amendments will not have any long-term or short term beneficial effects
in shoreline cleanup operations.
1.3.6.3 Oxygen amendment
Because wetland soils are inundated with water, the diffusion rates of oxygen through the soils are
very slow, and oxygen in the interstitial water is quickly depleted by aerobic metabolism of detritus
that is abundant in wetlands. A few centimeters, and often only a few millimeters below the
sediment surface, the wetland sediments are anaerobic. Therefore, oxygen is likely a limiting factor
for oil biodegradation in marine wetlands. However, an appropriate technology for 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. Many of the
oxygen amendment technologies developed in terrestrial environments (e.g. tilling, forced aeration,
and the addition of chemical oxidants), are currently not considered viable options for use in coastal
wetlands. There is concern that their deployment is expensive and environmentally intrusive.
Furthermore, their effectiveness in enhancing oil biodegradation in wetland environments is
unproven.
The Gladstone field trial showed that a forced aeration strategy was only able to increase the depth
of the aerobic layer of the wetland sediments from 1 mm to 2 mm, and could not significantly
stimulate oil biodegradation in the anaerobic mangrove environment (Burns etal., 2000). Strategies
involving the mixing of surface sediments, such as tilling or disking, have also been proven
ineffective in recent field studies (Garcia-Blanco and Suidan, 2001; Garcia-Blanco et a/., 2001b;
Venosa et a/., 2002). Not only does this approach cause severe ecological damage to wetlands, it
also enhances oil penetration deep into the anaerobic sediments, resulting in slower oil
biodegradation. As for adding alternative electron acceptors, there is no strong evidence yet to
suggest that the addition of nitrate as an electron acceptor can enhance oil biodegradation when
oxygen is limiting (Garcia-Blanco and Suidan, 2001; Townsend et a/., 1999). The high oil
degradation rates under sulfate-reducing conditions found in some laboratories (Caldwell et a/.,
1998; Coates etal., 1997) have not been convincingly demonstrated in the field. Therefore, further
research is still required to explore cost-effective oxygen amendment techniques for the
bioremediation of coastal wetlands.
1.3.6.4 Plant amendment (phytoremediation)
Phytoremediation, the stimulation of contaminant degradation by the growth of plants and their
associated microorganisms, is emerging as a potentially cost-effective option for cleanup of
petroleum hydrocarbons in terrestrial environments (Banks et a/., 2000; Frick et a/., 1999a).
Mechanisms responsible for oil phytoremediation may include degradation, containment, and the
transfer of contaminants from soil to the atmosphere (Cunningham etal., 1996). Frick etal. (1999b)
indicated that the primary loss mechanism for petroleum hydrocarbons is the degradation of these
20
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compounds by microorganisms in the rhizosphere of plants. Phytoremediation was hypothesized to
be particularly effective when used together with nutrient enrichment because hydrocarbon
contamination may result in nutrient deficiencies in contaminated soil. Added fertilizers could
increase the rate of oil degradation by indigenous microorganisms in the rhizosphere and
simultaneously stimulate plant biomass production, thereby increasing the effectiveness of
phytoremediation and accelerating the recovery of the affected wetland plant ecosystem.
Extensive studies have been conducted on the phytoremediation of petroleum hydrocarbons in
terrestrial environments (Frick etal. 1999a,b). Researchers at University of Saskatchewan, Canada,
recently developed a catalogue of plants with the potential to phytoremediate hydrocarbon
contaminated soils following a review of information in the literature and the conduct of field
surveys (Godwin etal., 1999; and Frick etal., 1999c). Nevertheless, only limited studies have been
carried out on the effectiveness of phytoremediation in enhancing oil degradation in coastal wetland
environments. Lin and Mendelssohn (1998) found in a greenhouse study that application of
fertilizers in conjunction with the presence of salt marsh and brackish marsh transplants significantly
enhanced oil degradation. In another mesocosm study, Dowty etal. (2001) evaluated the effects of
soil organic matter content, plant species, soil oxygen status and nutrient content on oil degradation
and plant growth response in fresh marsh environments. The study found that the amount of oil
remaining after 18 months was lowest in aerated and fertilized mesocosms containing either/1.
hemitomon or S. lancifolia and a substrate of low organic matter content. Field studies, however,
have not demonstrated such significant effects as in the mesocosm studies. A recent Nova Scotia
field trial showed that addition of nutrients did not result in significant enhancement of
biodegradation of crude oil, whether or not plants were left intact or removed (Garcia-Blanco and
Suidan, 2001). Similar results were also found in the St. Lawrence River freshwater wetland field
study (Garcia-Blanco etal., 200 Ib; Venosa etal., 2002). On the other hand, the results of these field
trials did suggest that although application of fertilizers in conjunction with the presence of wetland
plants may not significantly enhance oil degradation, it could accelerate habitat recovery. There is
evidence that nutrient amendments could stimulate vigorous vegetative growth, reduce sediment
toxicity and oil bioavailability (Lee etal, 2001a).
In summary, on the basis of field trials conducted to date, the effectiveness of phytoremediation in
enhancing oil degradation in coastal wetlands is highly site-specific and does not promise to be an
effective oil cleanup technique per se. However, it does show promise in accelerating the recovery
and restoration of wetland environments contaminated with oil and oil products, which is the
ultimate goal of the treatment.
1.3.6.5 Monitored natural attenuation
Natural attenuation has been defined as the reliance on natural processes to achieve site-specific
remedial objectives (USEPA, 1999). When used as a clean up method, a monitoring program is still
required to assess the performance of natural attenuation. This approach is increasingly viewed as
the most cost-effective, although the least cosmetically appealing, option for the cleanup of oil spills
in coastal wetland environments since it causes the least adverse ecological impacts often associated
with cleanup activities (Baker, 1999; Owens etal, 1999; Sell etal, 1995). Sell etal. (1995); Mills
et al, 2003) compared the rates of recovery between treated and untreated wetlands based on 20
case studies of heavily oiled salt marshes. They concluded that most traditional cleanup methods did
21
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not promote significant long-term ecosystem recovery.
Recent field studies on oil bioremediation have demonstrated that the availability of oxygen, not
nutrients, is often the limiting factor for oil biodegradation in coastal wetlands. However, as
discussed earlier, no feasible technique is currently available for increasing the availability of
oxygen in such an environment. Fortunately, these field studies also showed that the natural
biodegradation of alkanes and PAHs could occur to a very high extent and at similar rates in coastal
wetlands as in sandy beaches (See section 1.3.4). Therefore, in consideration of the potential impacts
associated with physical clean-up procedures in wetlands (i.e. trampling), natural attenuation should
be given more preference in decision making for oil spill cleanup in coastal wetlands when the oil
concentration is not high enough to destroy the ecosystem.
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2 RECOMMENDED APPROACHES TO BIOREMEDIATION IN SALT MARSHES
Existing studies have demonstrated that oil biodegradation on marine wetlands is often limited by
oxygen, not nutrient availability. Natural attenuation is increasingly becoming a promising and even
a preferred strategy for the restoration of oil-contaminated wetlands. However, field studies also
showed that on some coastal wetlands, nutrients might still be a limiting factor for oil
biodegradation, particularly if the oil does not penetrate deeply into the anoxic zone of the wetland
sediment (Lee and Levy, 1991; Mills etal., 1997; Venosa etal., 2002). Therefore, biostimulation
with nutrient amendment can still be an appropriate countermeasure treatment under some
circumstances. General guidelines for the bioremediation of oil-contaminated marine shorelines,
which are mostly derived from studies and practices on sandy beaches, have been presented in the
sandy shoreline and freshwater wetland guidance document (Zhu et a/., 2001). Although the general
principles for achieving successful oil bioremediation for all types of marine shorelines are the same,
a simple transfer of response strategies may not be necessarily the most appropriate since salt marsh
habitats are significantly different from other marine situations. Therefore, guidelines and special
considerations for oil bioremediation in coastal wetland environments are presented here based on
current understandings and field studies, particularly the findings of the Nova Scotia field study
(Garcia-Blanco and Suidan, 2001).
Similar to the general protocol presented in the sandy shoreline and freshwater wetland guidance
document (Zhu et al., 2001), a general procedure or plan for the selection and application of
bioremediation technology in salt marshes is illustrated in Figure 2.1. 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 biodegradability of the spilled oil, the depth of oil penetration
and oxygen availability, concentrations of background nutrients, the presence of
hydrocarbon-degrading microorganisms, the type of shoreline substrate, and other logistic
and environmental factors (pH, temperature, 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
step 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, lexicological, and ecological analysis.
This document will focus on the operational guidelines for decision-making and planning of oil
bioremediation in salt marshes. Guidelines with respect to the assessment of field results and
establishment of appropriate treatment endpoints can be found in the Chapter 6 of the sandy
shoreline and freshwater wetland guidance document (Zhu et al., 2001).
23
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Step 1:
Pretreatment Assessment
Oil
biodegradability
Oil penetration
and oxygen
availability
^
\
>
Background
nutrient content
>
Climate, prior oil
exposure and other
site characteristics
T
Ifbioremediation
is selected
Step 2:
Bioremediation Planning
Nutrient products
Nutrient application
strategy
Sampling and
monitoring plan
Step 3:
Assessment and Termination
Analysis of oil biodegradation
and physical loss
Toxicological and ecological
analysis
Figure 2.1 Procedures for the selection and application of oil spill biore mediation in salt marshes
24
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2.1 Pre-treatment Assessment
Major considerations in the assessment of the need for biostimulation in salt marshes include the
evaluation of 1) oil types and concentrations, 2) oil penetration and oxygen availability 3)
background nutrient content, and 4) other environmental factors such as the prevalent climate and
prior oil exposures. Among these factors, the assessments of oil penetration, oxygen availability and
background nutrient content are of particular importance for bioremediation of salt marshes and will
be discussed in following sections. Detailed discussion on the assessments of oil types and
concentrations, oil biodegradability, climate and other environmental factors can be found in the
sandy shoreline and freshwater wetland guidance document (Zhu et a/., 2001).
2.1.1 Oil penetration and oxygen availability
Unlike other types of marine shorelines (e.g. sandy beaches), the most important limitation for
cleanup of an oil-contaminated marine wetland is oxygen availability. Wetland sediments become
anoxic often below a few millimeters to centimeters of the soil surface. When substantial penetration
of spilled oil into anoxic sediments has taken place, available evidence suggests that biostimulation
with nutrient addition has limited potential for enhancing oil biodegradation, and it would likely be
best simply to leave it alone and not risk further damage to the environment by trampling and the
associated bioremediation activities. Therefore, the evaluation of oil penetration and oxygen
availability is probably the most important pre-treatment assessment for determining whether
bioremediation is a viable option.
The thickness of the oxidized layer within wetland sediments varies from a few millimeters to
several centimeters, depending on the population of oxygen utilizers, the rate of photosynthetic
oxygen production by algae, the soil chemical composition, and the rate of oxygen transport into the
wetland sediments (Mitsch and Gosselink, 2000; Shin etal, 2000). For example, soil organic matter
is a major oxygen sink in salt marshes and, therefore, oxygen deficiency is more likely to occur in
wetland soil with high organic matter content. Oxygen limitation will be less severe in the area
where the wetland surface is exposed to the atmosphere or is subjected to strong surface mixing by
convection currents and wave action.
The depth of the aerobic layer can be identified through both visual observation and measurements
of DO and redox potential. The wetland surface or the aerobic layer is often a brown or brownish-
red color due to the presence of ferric ions. The anoxic zone in wetland sediments is either bluish
gray due to the presence of ferrous ions or, more often in salt marshes, black along with a foul odor
associated with the production of hydrogen sulfide under sulfate reducing conditions. Anoxic
conditions can also be determined by measuring dissolved oxygen in pore water. Oxygen will
become a limiting factor when DO concentration in pore water approaches zero. When using DO
probe, a reading of 0.1 or 0.2 mg/L indicates the depletion of dissolved oxygen. Redox potential is a
more sensitive measurement of the degree of reduction of wetland sediments. For example,
denitrification occurs at a redox potential of approximately 250 mV and sulfates are reduced to
sulfides at a redox potential between -100 and -200 mV (Mitsch and Gosselink, 2000).
The depth of oil penetration also depends on many factors, such as oil type, concentration and
shoreline substrate. In general, fresh crude oils and heavy oils tend to adhere to the marsh surface
25
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sediment or pool on the sediment surface. Light oils and oil components can generally penetrate the
top few centimeters of wetland sediment. However, penetration can be much deeper into burrows
and cracks extending up to one meter (NOAA, 1992). A microcosm study on the penetration of
weathered light Arabian crude oil in freshwater wetland sediments showed that the oil was able to
penetrate about 2.5 cm in 16 weeks for both a flooded condition and a saturated but non-flooded
condition (Purandare, 1999). However, the amount of the oil able to penetrate into the sediment was
much less for flooded sediments, where most of the oil floated on the surface of the water. The depth
of oil penetration also increases with the increase of oil concentration and therefore affects the
potential of oil biodegradation. In the field trial reported by Lee and Levy (1991), the rates of oil
degradation in the salt marsh were not stimulated by nutrient amendments at the higher test
concentration (3.0% v oil/v sediment), where oil penetrated to anaerobic layers of sediment.
However, bioremediation was effective at the lower test concentration (0.3 % v oil/v sediment),
where the oil did not penetrate beyond the aerobic sediment surface layers. In the same field study,
Lee and Levy (1991) also examined the effect of oil concentration on a sandy beach and found that
oil biodegradation rates were enhanced by the nutrient amendment at the higher oil concentrations
(3%), where oxygen was not a limiting factor. The result suggested that the favorable concentrations
for using bioremediation would be much lower in salt mashes than on sandy beaches.
The type of shoreline substrate is another important factor affecting the oil penetration and the
feasibility of using bioremediation. Shoreline substrate can affect oil penetration from the
perspectives of both the sediment texture and the soil chemical composition. Generally, oil
penetrates coarse sediments more readily than fine sediments. However, because the texture of all
wetland sediments is normally very fine, the substrate chemical composition plays a more important
role in oil penetration in salt marsh environments. Studies have shown that the rates of oil
penetration and biodegradation are strongly related to the soil organic matter content (Dowty etal.,
2001; Lin and Mendelssohn, 1996). Oil is more likely to penetrate into sediments with higher
organic content since it associates more readily with organic matter than with mineral particles. In a
greenhouse study, Lin and Mendelssohn (1996) investigated the performance of oil biodegradation
in three types of coastal wetlands - salt, brackish, and freshwater marsh. They found that the rates of
oil degradation were highest in the salt marshes and lowest in the freshwater marshes. The difference
in oil residue was mainly attributed to the difference in the soil organic content, which was lowest in
the salt marsh sediments and highest in the freshwater marsh sediments. The study also measured the
concentrations of the oil that penetrated the soils in digested (to remove the associated organic
matter) and undigested marsh soil and found that the oil concentrations were 41 -279 times higher in
the undigested soil than the digested one. Similar results were observed by Dowty etal. (2001) in a
mesocosm study conducted in fresh marsh environments. They found that the rates of oil
degradation were significantly higher in the inorganic sediments than the organic ones under
different oil concentrations and nutrient levels. These results are consistent with the notion that
oxygen demand is higher and oil is more readily able to penetrate into organic sediments. Therefore,
oil bioremediation seems more likely to be successful when applied in a wetland with lower organic
matter content.
2.1.2 Background nutrient content
To determine whether nutrient amendment is a viable option, it is necessary to assess the
background nutrient levels in the contaminated site, particularly the nutrient concentrations within
26
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the interstitial water in that environment. There is no need to add nutrients if natural nutrient
concentrations are high enough to sustain rapid intrinsic rates of oil biodegradation. However,
because oxygen is often the determining factor in oil degradation on coastal wetlands, the
assessment of background nutrient concentration is important and needed only after the assessments
of oil penetration and oxygen availability conclude that oxygen limitation is not a serious
impediment. In other words, when substantial oil penetration into the anoxic zone of the wetland
sediments occurs, nutrient amendment is not likely to be effective even if nutrient deficiency exists.
As shown in the St. Lawrence River field trial (Venosa etal., 2002), the average pore water nitrogen
concentration in natural attenuation plots was only about 0.74 mg N/L, well below the levels needed
for maximum hydrocarbon biodegradation (Venosa etal., 1996). However, the dramatic increase in
nutrient levels in the biostimulation plots did not enhance oil biodegradation above that achievable
in the natural attenuation plots due to the oxygen limitation within that freshwater wetland sediment.
However, when oxygen availability is not a limiting factor, the decision to use nutrient amendments
should be based on how high the natural levels are relative to the optimal or threshold nutrient
concentrations. It has been recommended in the sandy shoreline and freshwater wetland guidance
document (Zhu et al., 2001) that the threshold concentration for optimal hydrocarbon biodegradation
on marine shorelines is in the range of 2 to 10 mg N/L based on the field experiences on sandy
beaches (Bragg et al., 1994; Venosa et al., 1996) as well as in an estuarine environment (Oudet et
al., 1998). Although no such threshold concentration has been experimentally identified in salt
marsh environments, recent field experiences did provide some insights. The Nova Scotia study
found that the average background nitrogen concentration in pore water was about 10 mg N/L at the
experimental site. Thus, nitrogen limitation was not an important factor (Garcia-Blanco and Suidan,
2001). The ineffectiveness of nutrient amendments in enhancing oil biodegradation under this high
background nutrient level suggested that the nitrogen threshold concentration should be lower than
10 mg N/L. However, the San Jacinto River study suggested that the threshold nitrogen
concentration may be higher than 2 mg N/L on coastal wetlands. During Phase II of the study,
nutrient addition apparently enhanced oil degradation even when the background nitrogen
concentration was about 5 mgN/L (Harris etal, 1999; Mills etal., 1997). This study, however, was
inconclusive because the same enhancement was not observed when the treatment was repeated
during the following year (Simon etal., 1999). Although further research is still needed, it appears
from existing evidence that the threshold nitrogen concentration for optimal oil biodegradation in
salt marshes is likely similar to that obtained in other shoreline types (e.g. 2-10 mg N /L).
The investigation of background nutrients should also determine whether the present nutrient
concentrations are typical of the area or sporadic (i.e., determine the impact of chronic runoff from
nearby agricultural practices and local industrial and domestic effluents). As described in Part I,
coastal marshes are generally considered high-nutrient wetlands. However, inorganic bioavailable
nutrient concentrations in salt marsh sediments may exhibit a strong seasonal pattern with a
concentration peak usually during the summer months probably due to a high mineralization rate at a
higher temperature (Cartaxanae^ al., 1999; Nixon et al., 1980). The available nutrient levels can also
be elevated as a result of runoff, fire and death of plants. If these events are sporadic, biostimulation
may still be appropriate when the nutrient levels fall below threshold concentrations.
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2.1.3 Summary of pretreatment assessment
Based on the current understandings discussed in the previous sections as well as in the sandy
shoreline and freshwater wetland guidance document (Zhu et al., 2001), the following pretreatment
assessments should be conducted to determine whether bioremediation is a viable option in response
to a spill incident in salt marsh environments:
• 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
alkanes are also 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 (of any weight) may also inhibit biodegradation. For details, see
Zhu etal. (2001).
• Determine whether oxygen is a factor limiting oil biodegradation by measuring the depth of
oxidized sediment layer and the extent of oil penetration - When a substantial portion of the
spilled oil has penetrated into anoxic sediments, biostimulation with nutrient addition has
limited potential for enhancing oil biodegradation. Oxygen limitation is less likely to occur
in wetland sediments with lower organic matter and/or contaminated with oil at moderate
concentrations.
• Determine whether the nutrient content at the impacted area is likely to be a limiting factor
by measuring the background nutrient concentrations within the interstitial water in that
environment - If oxygen is not the limiting factor, 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., > 5 mg N/L). 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 climatic or seasonal conditions are favorable for using bioremediation -
Bioremediation may be more effective during warmer seasons, particularly in cold
environments, since oil biodegradation rates are higher during these seasons. However, this
does not necessarily mean that summer is the most favorable season. Because inorganic
nutrient levels in salt marsh sediments often peak during the summer, biostimulation will not
be effective if the nutrient content is no longer the limiting factor during warmer seasons.
Prior exposure to oil will also be a favorable but not a solely determinative condition for
selecting bioremediation.
2.2 Treatment Selection and Design
If biostimulation by nutrient addition is determined to be a potentially effective cleanup option based
on the pretreatment assessments, further evaluation and planning are needed before its application.
This step involves selection of the rate-limiting nutrients, determination of optimal nutrient
concentrations and application strategies, and design of sampling and monitoring plans.
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2.2.1 Nutrient selection
One of the first tasks during the stage of treatment selection and design would be to select
appropriate nutrient products. The laboratory treatability tests, especially well-designed microcosm
or mesocosms tests, are most commonly used approaches to determine the type and level of
amendments. However, responders will likely not have time or resources to conduct a treatability
study. This section, therefore, serves to support a reasonable approach to deciding which type of
formulation to use.
Screening and treatability tests that have been reported in the literature involve the determination of
rate limiting nutrients as well as optimal forms of nutrient species. Nitrogen, phosphorus, or both can
limit oil degradation in salt marshes. In a microcosm study, Jackson and Pardue (1999) found that
nitrogen but not phosphorus was the rate-limiting nutrient for oil degradation in sediment from a
Louisiana salt marsh. Wright et al. (1996, 1997) reported the opposite result for a mesocosm study
where oil degradation was mainly limited by the concentration of phosphorus in sediment from a
Texas salt marsh.
The molecular form of nutrients is also important. For example, although both ammonium and
nitrate are capable of enhancing oil degradation when nitrogen is a limiting factor, their
effectiveness may differ depending on the type of oil and the properties of shoreline substrate.
Jackson and Pardue (1999) found that addition of ammonium appeared to stimulate degradation of
crude oil more effectively than nitrate in salt marsh soils in a microcosm study. The ammonium
requirement was only 20% of the concentration of nitrate to achieve the same increase in
degradation. The authors concluded that ammonium was less likely to be lost from the microcosms
by washout due to its higher adsorptive capacity to sediment organic matter. A recently completed
study at a salt marsh in Nova Scotia also showed that the ammonium spikes after nutrient addition
were always substantially higher than the nitrate spikes, even though the only exogenous source of
nitrogen was NH4NO3 (Table 1.1). The lower pore water nitrate concentrations can be attributed to
the higher washout rate for nitrate and its loss through denitrification within the anoxic sediments.
Under such circumstances, ammonium based nutrients may be superior to nitrate based nutrients
because the nutrient dosage will be much lower when using ammonium than nitrate to achieve the
same pore nitrogen concentration. However, this may not always be the case. Actually, the St.
Lawrence River field study showed that the pore water nitrate concentrations were always higher
than the ammonium concentrations after NELtNOs was added (Venosa etal., 2002). This finding was
attributed to the adsorption of NH4+ onto the negatively charged soil particles and its uptake by the
root systems of the wetland plants. This result also suggests that the effects of nitrate washout and
denitrification were less important in this fresh water marsh.
Nutrient selection might also be influenced by temperature conditions. In a field study, Lee et al.
(1993) investigated the efficacy of water-soluble inorganic fertilizers (ammonium nitrate and triple
super phosphate) 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 suppress nutrient release rates. Water-soluble fertilizers
29
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such as ammonium nitrate were then recommended under these temperature conditions. Based on
the above discussion, it is recommended that, if temperature conditions allow use of slow-release
fertilizer (i.e., temperatures in excess of 15°C), then that would be the preferred fertilizer to use in a
salt marsh. If the temperature were lower, then ammonium nitrate would be appropriate. In either
case, the amount of fertilizer to use should be based on maintenance of a minimal amount that would
not limit biodegradation (i.e., something greater than about 5 mg N/L and 0.5 mg P/L).
In addition to demonstrating the efficacy of nutrient products in enhancing oil degradation, it is also
critical to demonstrate that bioremediation products have low toxicity and do not produce any
undesired environmental and ecological effects, especially when applied to such sensitive
ecosystems as salt marshes. Various toxicity test protocols have been discussed in part I as well as in
the sandy shoreline and freshwater wetland guidance document (Zhu etal, 2001). A case study on
the assessment of bioremediation treatment through monitoring biological responses in an oil-
contaminated salt marsh will also be presented later in this document.
2.2.2 Concentrations 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 etal., 1994; Venosa etal., 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 provide no added benefit but may lead to potentially
detrimental ecological and toxicological impacts.
Only a few studies have been reported on the optimal nutrient concentration in salt marsh
environments. In a microcosm study using salt marsh sediment slurry, Jackson and Pardue (1999)
found that oil degradation rates could be increased with increasing concentrations of ammonia in the
range of 10 - 670 mg N/L, with most of the consistent rate increases occurring between 100 -670 mg
N/L. They further proposed a critical nitrogen concentration range of 10-20 mg N/L. Harris et al.
(1999) examined the nutrient dynamics during natural recovery of an oil-contaminated brackish
marsh and found that there was an interdependency between the natural nutrient levels and the extent
of oil degradation when the background nitrogen concentration in pore water declined from 40 mg
N/L to 5 mg N/L. Evidence from bioremediation field studies also suggested that concentrations of
approximately 5 to 10 mg/L of available nitrogen in the interstitial pore water is sufficient to meet
the minimum nutrient requirement of the oil degrading microorganisms (Garcia-Blanco and Suidan,
2001; Mills et al., 1997; See Section 2.1.2). As mentioned earlier, the threshold concentration range
for optimal hydrocarbon biodegradation on marine shorelines is around 2 to 10 mg N/L based on
field experiences on sandy beaches (Bragg et a/., 1994; Venosa et a/., 1996) and in an estuarine
environment (Oudet et al., 1998). The apparent higher threshold nitrogen concentrations in salt
marshes are mainly due to the lack of information with respect to oil biodegradation under lower
nitrogen concentrations, since all the existing field studies were conducted in salt marshes with
background nitrogen concentrations of at least 5 mg N/L (Garcia-Blanco and Suidan, 2001; Harris et
al., 1999; Mills etal., 1997; Shin etal., 1999). Therefore, it is reasonable to recommend, as for other
types of shorelines, that biostimulation of oil impacted salt marshes 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,
30
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and practicality. In practice, a safety factor should be used to achieve target concentrations, which
will depend on anticipated nutrient washout rates, selected nutrient types, and application methods.
The safety factor used in salt marsh environments may generally be smaller than that used in higher
energy beaches due to the reduced degree of nutrient washout expected in salt marshes. One needs
always to keep in mind, however, that nutrient toxicity might exist if too much nutrient is applied to
a coastal wetland (Mueller et al., 1999). The factors that lead to higher nutrient losses in wetland
environments may also be important, such as sediment adsorption, plant uptake, and denitrification
(if applicable).
2.2.3 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.
2.2.3.1 Frequency of nutrient addition
The frequency of nutrient addition to maintain the optimal nutrient concentration in the interstitial
pore water mainly depends on shoreline nutrient loss rates. A tracer study conducted on a low-
energy beach and a high-energy beach in Maine demonstrated the influence of shoreline types on
nutrient washout rates (Wrenn et a/., 1997a; Zhu et a/., 2001)). The study shows 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 during this period. Salt marshes are low-energy systems and nutrient
washout rates in such environments should be similar to the observations made on the low energy
beach in the Maine study. In salt marshes, the washout rates may be further reduced when using
ammonium-based nutrients due to their higher affinity to adsorb onto the sediment as compared to
nitrate-based fertilizers (Jackson and Pardue, 1999). Therefore, weekly to monthly additions may be
sufficient for biostimulation of salt marshes when the nutrients are applied during neap tide. It is
even possible that only one nutrient dose is required for the bioremediation of some coastal marshes.
A study on the nutrient dynamics in an oil contaminated brackish marsh showed that it took more
than one year for nutrient concentrations to decrease to background levels after being naturally
elevated by flooding and perturbations due to the spill (Harris etal., 1999). However, this may not
be truly indicative of nutrient application dynamics, since exogenous nutrients were not added in this
case. Nutrient sampling, particularly in sediment pore water, must 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.
2.2.3.2 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. In many
intertidal environments, particular high-energy shorelines, the primary consideration in developing
and selecting a nutrient application method has been how to overcome the washout problems. Many
attempts have been made in this regard, including the use of slow release and oleophilic fertilizers
-------
(Prince, 1993) and the subsurface application of nutrients (Wise et a/., 1994). However, since
nutrient washout in coastal wetland environments is relatively slow, the more important
considerations in such cases should be on the use of less expensive and less environmentally
intrusive application methods. As discussed in the sandy shoreline and freshwater wetland guidance
document (Zhu et a/., 2001), current experience indicates that surface application of dry granular
fertilizer (either slow-release or water-soluble) to the impact zone at low tide is probably the most
cost-effective and less environmentally intrusive way to control nutrient concentrations.
2.2.4 Sampling and Monitoring Plan for Bioremediation Operations
2.2.4.1 Important variables and recommended measurements
Important variables to be monitored in an oil bioremediation project include the environmental
factors that limit oil biodegradation rates (e.g., temperature, interstitial nutrient and oxygen
concentrations), evidence of oil biodegradation (e.g., concentrations of oil and its components),
microbial activity (e.g., bacterial numbers and activity), and toxicological effects. Primary variables
recommended for monitoring of bioremediation field programs in coastal wetland environments are
listed in Table 2.1.
If pretreatment assessments determine that oil biodegradation in the field is likely to be limited by
nutrient rather than oxygen availability, pore water nutrient analysis becomes 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. This is to make certain that (1) the treatment is reaching and penetrating the impact
zone, (2) target concentrations of nutrients are being achieved, and (3) toxic nutrient levels are not
being reached. The location from which nutrient samples are collected is also important. Recent
research on solute transport in the intertidal zone has shown that nutrients may remain in the beach
subsurface for much longer periods than in the bioremediation zone (Wrenn etal., 1997b). Nutrient
concentration profiles along the depth of the oil-contaminated region may be monitored by using
multi-port sample wells or by the extraction of sediment samples collected from the oil-
contaminated region (Venosa et a/., 2002; Lee et a/., 2001a). The sampling depth should be
established from the results of site surveys to determine the maximum depth of oil penetration. To
counter the inherent heterogeneity observed in field studies, a positive "margin of error" should be
added to ensure that the samples will encompass the entire oiled depth throughout the project. The
sampling depth must 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 of the sandy shoreline and
freshwater wetland guidance document (Zhu et a/., 2001), 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 occasionally using chromatographic techniques
(e.g., GC/MS) and are reported relative to a conservative biomarker such as hopane. However, 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 protocols for the
measurement of TPH, especially those using infrared absorption of Freon-extracts, are generally not
32
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reliable and have limited biological significance. Using GC/FID and integrating the area under the
chromatogram is better. TLC-FID appears to be a promising screening tool for monitoring oil
biodegradation (Stephens et a/., 1999), although not enough experience is available to make any firm
recommendations on its use at this time.
As suggested in the foregoing paragraph, GC/MS operated in the selected ion monitoring mode
(SIM) is the preferred method to use to assess the progress of biodegradation. One sampling per
month of composited samples from the site analyzed by GC/MS should suffice to provide evidence
that hydrocarbons are being biodegraded. To this end, normalization of biodegradable constituents in
the oil to hopanes, steranes, and/or other potential biomarkers (e.g., highly substituted 4- or 5-ring
PAHs like C4-chrysene) is essential to ensure that the disappearance observed is due to the
bioremediation action rather than physical washout. Samples are normally extracted with
dichloromethane and cleaned up using column chromatography prior to conducting the GC/MS.
However, due to the expense and expertise involved with GC/MS analysis, more frequent analysis of
TPH is appropriate to follow the temporal progress of treatment. It is suggested that at least one TPH
sampling event per week be conducted at the spill site. Either the gravimetric or GC/FID method of
TPH analysis should be used. Interpretation of chromatographic methods may be confounded by the
presence of plant lipids and other biogenic compounds present in the environment; thus, care should
be exercised in interpreting results. For example, plant lipids normally give rise to peaks in the
chromatograms at retention times that coincide with odd-numbered higher molecular weight alkanes
in the range of €25, C2?, C29, Csi, and €33. Thus, it is essential that the chromatogram of the spilled
oil be known to compare to actual samples analyzed.
In addition to monitoring treatment efficacy, 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 also been summarized in the sandy shoreline and freshwater
wetland guidance document (Zhu et a/., 2001). These micro-scale bioassays may provide an
operational endpoint indicator for bioremediation activities on the basis of toxicity reduction (Lee et
a/., 1995b). A summary of criteria for selecting an appropriate bioremediation endpoint based on
both oil degradation and toxicity reduction has been presented in the sandy shoreline and freshwater
wetland guidance document (Zhu etal., 2001). Examples for a salt-marsh study are presented in the
following sections.
33
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Table 2.1 Monitoring plan for an oil bioremediation project in a coastal wetland environment
Analysis
*Dissolved nitrogen
Dissolved phosphorus
*Residual oil
constituents
* Total petroleum
hydrocarbons
Redox and sulfide
*Dissolved oxygen of
pore water
pH of pore water
Microbial populations
Microbial activity
*Toxicity of residual
oil
Shoreline profile
Matrix
Sediment
(interstitial
pore water)
Sediment
(interstitial
pore water)
Sediment
Sediments
Sediment
(interstitial
pore water)
Aqueous
Aqueous
Sediment
Sediment
Sediment,
pore water
Contaminated
site
Recommended Methods
Extract in acidified 0.1% NaCl.
4500-NH3 H (Automated
Phenate Method) and
4500-NO3" F (automated Cd-
reduction)
Extract in acidified 0.1% NaCl.
4500-P E (ascorbic acid method)
Extract into dichloromethane
(DCM).
Analyze components by GC/MS-
SIM
Gravimetric analysis
(dichloromethane extraction) or
GC/FID analysis of DCM
extracts.
Redox and sulfide electrodes
Hach high range assay
Potentiometric with combination
electrode
MPN for alkane and PAH
degraders
Genetic biomarkers
Uptake/respiration of
radiolabelled substrates.
In-situ respiration
Biotests (e.g. Microtox Test,
Amphipod survival test, MFO
induction, etc.)
Intertidal/supratidal zone surveys
using fixed benchmarks at the
study site. (e.g., wells, plot
boundary markers).
References
Eaton etal., 1995
Page etal, 1986
Eaton etal, 1995
Page etal, 1986
Venosaetal, 1996
NET AC, 1993
ATI Orion,! 99 la,b
Hach Company,
Loveland, CO
Page etal, 1986
Wrenn and Venosa,
1996
Macnaughton et al.,
1999
Lee and Levy, 1989
Prince et al, 1999
See Section 1.3.5
and 2.2.4.2
Wrenn et al, 1997a,b
* Critical measurements
The sandy shoreline and freshwater wetland guidance document (Zhu et al., 2001) presents other
important variables in a comprehensive monitoring plan, including site background conditions (e.g.,
oxygen, redox, pH, sediment grain size, and temperature) and shoreline profiles. Oxygen availability
34
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is crucial for rapid bioremediation since hydrocarbon biodegradation is primarily an aerobic process.
Although the pretreatment assessments may have determined that oxygen availability might not be a
serious concern for the on-going project, oxygen limitation is always a potential problem in a
wetland environment. 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 a/., 1995b; Sveum and Ramstad, 1995; See
Section 4.1.3), to insure that anoxic conditions do not result. When available oxygen does become
limiting, the nutrient dosage and application frequency should be adjusted accordingly. Monitoring
oil penetration and analyzing redox potential and sulfide concentrations with depth of wetland
sediments will assist in determining whether oil has penetrated into the anoxic zone during the
process of bioremediation. This assessment can also be used as a criterion in determining treatment
endpoints.
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. For
accurate interpretation of field data, analysis of sediment grain size should be conducted to verify
study site homogeneity.
2.2.5 Environmental assessment of an oil-contaminated salt marsh: a case study
Since environmental assessment is a relatively new approach in evaluating the effectiveness of oil
bioremediation treatments, a case study outline is presented as a means to provide operational
guidance to spill responders. This example is based on a controlled oil spill field trial recently
conducted in a salt marsh at Conrod's Beach, Nova Scotia, Canada, to determine if bioremediation
by nutrient enrichment or phytoremediation by enhanced plant growth would accelerate the rates of
residual oil loss and habitat recovery. The experimental design and bioremediation performance with
respect to oil biodegradation has been reviewed (Section 1.3.3). Standard bioassessment and biotest
procedures (Section 1.3.5) were used to quantify the rates of habitat recovery and to identify
detrimental treatment effects (e.g., toxicity of the bioremediation agent or oil degradation by-
products). The overall success of the remedial operations was based on the integration of results
from a suite of assays, which were chosen on the basis of ecological relevance to the site of concern,
cost considerations, and the availability of technical expertise (Venosa and Lee, 2002).
35
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2.2.5.1 Bioassessments.
2.2.5.1.1 Recovery of vegetation. Growth (biomass) of the predominant plant species (Spartina
alterniflord) within the salt marsh was significantly suppressed by oiling, and recovery was not
observed during the first growing season. While initial results showed some recovery in all oiled
plots in the following spring, there was also evidence of changes in species composition within the
treated plots. Another opportunistic plant species that was more tolerant to the altered site conditions
increased its percentage of cover. By the end of the second growing season, the treated plots showed
substantial evidence of recovery in the sections of the plots that were not removed by sampling.
2.2.5.1.2 Microbial responses.
• Oil degradation potential. Potential hydrocarbon degradation rates of representative alkane
and PAH components within sediment samples were determined by quantifying the
respiration rate of added 14C-labelled hexadecane and 14C-labelled phenanthrene as
representatives of w-alkane and PAH class components within the test oil (Lee and Levy,
1989, Caparello and LaRock 1975, Walker and Colwell, 1976). Time-series changes (Week:
4, 7, 9, 12, 16, 20) in the turnover time of these specific tracers were calculated with the
actual concentrations of residual hexadecane and phenanthrene in each sample determined
by GC/MS to account for dilution by the unlabelled fraction of the specific substrates under
study.
Results of the added 14C-labeled hexadecane studies clearly illustrated the stimulation of
indigenous organisms with the potential to degrade alkanes within the first 10-weeks after
the application of oil (Figure 2.2). The lower the turnover time, the greater is the stimulatory
effect (lower turnover times mean higher biodegradation rates). A stimulatory effect on
potential hexadecane degradation rates by the addition of nutrients to unoiled sediments was
also observed. However, within the oiled sediments, remedial treatments based on nutrient
additions did not appear to cause a stimulatory effect that could be adequately resolved by
measurement of hexadecane respiration rates. Natural attenuation (Treatment C: oil without
nutrients in the figure) appeared to be relatively effective. These radiotracer studies are in
agreement with detailed chemical analysis that showed that 87% of the target w-alkanes were
degraded in the test sediments within 20 weeks. Similar observations were made for the
biodegradation of PAHs (represented by phenanthrene, a 3-ringed poly cyclic aromatic
hydrocarbon) with the exception that nutrient amendments to the unoiled control sediments
had no stimulatory effect (Figure 2.3) as contrasted with the hexadecane results. These
observations are in full agreement with the corresponding field studies on microbial growth
by MPN analysis. It was noted that besides oil, the addition of nutrients to unoiled plots also
resulted in an increase in the number of potential w-alkane degraders by two orders of
magnitude with respect to background levels. Only oiled plots showed an increase in the
number of PAH degraders.
• Denitrification activity. Denitrification is a primary process that regulates the nitrogen cycle
in wetland sediments (Figure 1.1). Microbial denitrification activity was monitored on each
sampling occasion by placing a gas chamber on each plot and taking headspace samples over
a 30-minute period, which were subsequently analyzed for nitrous oxide, an intermediate in
36
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the denitrification of nitrate. The seasonal average denitrification activity, plotted against
treatment type showed that in all cases where nutrients were applied (Treatments D = oil +
nutrients, E = oil + nutrients + cut plants, F = oil + nutrients + disking), there was a net
positive denitrification potential (Fig. 2.4). Natural attenuation (Treatment C: oil without
nutrients) and Treatment A (unoiled control) showed a net negative denitrification potential.
These results indicate that nutrient application resulted in increased denitrification activity in
the sediments.
C7UU ~
.— v 80° •
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liniinilllll nnfifrrirT linilfrnnrr llllnrTFinpT
479 12162062 479 12162062 479 12162062 479 12162062 479 12162062 479 12162062
Control Control Oil Oil Oil Oil
Nutrients Nutrients Nutrients Nutrients
Cut Disking
Figure 2.2. Average turnover time of hexadecane. Error bars = 1 standard error.
37
-------
ouuuu -
20000 -
'(fl 10000 -
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Q^ 400 -
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Control Control
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Oil Oil Oil Oil
Nutrients Nutrients Nutrients
Cut Disking
Figure 2.3. Average turnover time of phenanthrene. Error bars = 1 standard error.
38
-------
ouu
^ 600
\ 40°
0 200
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X U
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i , i , i , i , i ,
Oil
No oil
Nutrients
Figure 2.4. Seasonal average in denitrification activity with different treatments from Conrod's
Beach sediments from June to November, 2000. Letter designations are the same as in Figure 2.2.
• Structural deformity ofForaminifera. For ecological relevance in the monitoring of potential effects
in contaminated environments, it is preferred to use native (indigenous) species as indicators. In this
study, Foraminifera (forams), single-celled microorganisms (protozoa) that construct a shell from
available mineral particles or secrete one of calcium carbonate or of silica, were found to be a unique
indicator species, due to their sensitivity to residual oil. This is attributed to the fact that the process
of forming their shell has been reported to be highly susceptible to certain types of environmental
pollution resulting in deformities. Foram skeletons are also resistant to decay, and many are found as
fossils. Having these properties, foram tests and deformities can be used to monitor the ecological
effects of oil spills and treatments.
39
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The forams under observation at the study site were typically between 63-500 |im. They occurred in
the sediment at an abundance of 400-4000 species per sample. Sediment samples (1 cm depth) were
taken with a metal 10-cc core bi-weekly for the first two months and monthly for the last three
months until the end of the study. The samples were sieved, processed, and analyzed under a
stereomicroscope to determine the types of species, the number of living vs. dead, and normal vs.
deformed populations. Preliminary results from the first study season suggest that the oil impacted at
least one particular species offorams.,Miliaminajusca, resulting in a high percentage of structural
deformity in comparison to non-oiled specimens. Time-series studies can prove an estimate of the
time required for natural attenuation or remedial treatments to reverse this biological effect.
2.2.5.1.3 Bioassays. Establishing the actual exposure level of biota to residual oil is difficult. While
chemical measures of oil in sediments, water, and tissues are routine, there is no guarantee that all
biological organisms accumulate oil or its components equally or in proportion to environmental
concentrations. Further, many of the components of oil such as alkanes and PAHs are metabolized,
so that chemical analyses of tissue may not represent the true dose or dose rate. The key to sediment
assessment is bioavailability since elevated concentrations of toxic compounds may not necessarily
result in adverse effects to the organisms living within the sediments. 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
to determine 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.
• Microtox solid phase test. In the Microtox® Solid Phase Test (AZUR Environmental, 1999; Lee et
a/., 1995b; Microbics Corporation, 1992), the bacterium, Vibriofisheri, is exposed to test sediments.
A significant decrease in bioluminescence relative to water-only controls is indicative of sediment
toxicity. Toxicity levels are calculated as the concentration of sample that would result in a 50%
reduction in luminescence ('effective concentration,' ECso). To account for interference 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 2.5a). If one sets an
arbitrary ECso toxicity threshold at 1,000 mg/L [which Environment Canada uses in its regulations
(Tay et a/., 1997)], then even though there was a detrimental response observed in the control
sediments treated with nutrients only, all unoiled sediment samples would be deemed non-toxic
according to this guideline, while toxicity was identified following oil treatments (Fig. 2.5a). There
is no implied suggestion that the 1,000 mg/L threshold is being or should be adopted by EPA. The
threshold was reported as an example to demonstrate how one may utilize toxicity data in decision-
making. On comparison of results, it appears that natural attenuation (the bars labeled Oil in the
figure) could account for most recovery. By week 9, all treatments were non-toxic. The significance
of natural attenuation was also illustrated by a comparison of the relative recovery of the plots using
ECso's for each treatment and sampling time normalized to the unoiled control (Fig. 2.5b).
40
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O)
O
ill
13000
12000 -
11000 -
10000 -
9000 -
8000 -
7000 -
6000 -
5000 -
4000 -
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Control Nf°ntro'
Nutrients
Oil
Oil
Nutrients
Oil Oil
Nutrients Nutrients
Cut Disking
Figure 2.5. Sediment toxicity for sediment samples from Conrod' s Beach, Nova Scotia, at weeks 0
to 62, as reflected by (a) EC50 for experimental treatments and (b) EC50 normalized to the mean
control value at each sampling time. Error bars = 1 standard deviation.
41
-------
Amphipod survival test. The Amphipod Test measured the effects of sediment samples on survival
of sediment-dwelling Eohaustorius estuarius (Environment Canada, 1992). 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. The results are reported as percent
mortality (Figure 2.6). Mortality was high in all of the oiled treatments, but it began to decrease by
Week 12 largely as the result of natural processes. Addition of nutrients accompanied by disking
appeared to cause the most rapid rates of detoxification (recovery) within the oiled plots as measured
by this test. However, the results of chemical analyses (GC/MS) indicated that this observation
could also be attributed to the physical removal of oil (enhanced dispersion with tides) mediated by
disking operations. By Week 62, the difference between the disked plots and the natural attenuation
plots was highly significant (32% vs. 5% mortality).
^
0^
s
re
o
E
100
90
80
70
60
50
40
30
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Figure 2.
Survival
1
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1
1
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I
OCMI^CMOCM OCMI^CM
• CM O CM
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Control
Nutrients
Oil
6.
Test
Oil Oil Oil
Nutrients Nutrients Nutrients
Cut Disking
Changes in sediment toxicity from 0 to 62 weeks as quantified by the Amphipod
Error bars = 1 standard deviation
Gastropod survival. Although many organisms have been used as sentinels or bio-monitors 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. Mollusks are
42
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abundant and widely distributed, and their use as in situ bio-monitors has been on the rise (Lagadic
and Caquet, 1998). Saltwater marshlands present unique restoration challenges following oil spills
due to the sediment's high capacity for oil absorption, low oxygen content, fluctuating salinity, and
tidal flow. The mud-snail, Ilyanassa obsoletci, an abundant detritovore inhabiting these marshlands,
was evaluated for its suitability as a bio-monitor to assess the impacts or efficacy of the
bioremediation treatments. It was selected for use as an in situ bio-monitor as it feeds on sediment
detritus, algae and decaying organic matter within the wetland. Snails (n = 50/treatment/sampling
time) were caged in 20 x 20 x 22 cm open mesh polypropylene baskets moored to the sediment
surface of experimental plots. At the end of the second year, cages were recovered after being
exposed to the experimental plots for 30,60 and 90 days to evaluate effects on survival at the end of
the second field season (Week 62). Healthy snails were also exposed for a 30-day period under
laboratory conditions to test sediments recovered from the plots, and to determine survival rates. The
mud snails did not survive long in captivity in the field, and laboratory exposures were erratic.
Mortality in the field was likely due to environmental factors as mortality was high even in control
cages. It is unlikely that anoxia due to crowding and/or eutrophic conditions was a factor since these
snails are tolerant of anoxic conditions and can grow in dense aggregates. Mortality after 5 d
exposure was generally higher for snails caged within the experimental plots amended with
nutrients. This toxic response was attributed directly to the use of fertilizers.
Acute and chronic effects on fish. Fish biotests were performed with salt marsh sediments recovered
from the experimental site at Conrod's Beach, Dartmouth, Nova Scotia using euryhaline rainbow
trout. Bioavailability was assessed by quantifying the extent of CYP1A (MFO enzyme, see Section
1.3.5.2) induction (Guiney et a/., 1997) in fingerling trout following 96 h exposures to test
sediments. S-9 fractions from liver homogenates were prepared for the measurement of
ethoxyresorufin-o-deethylase (EROD, CYP1A enzyme) activity (Hodson et a/., 1996). Each
bioassay included negative controls (water only), un-oiled sediment controls, and positive water
controls (fish exposed to the compound B-naphthoflavone, which is a model inducer).
All data were analyzed after log transformation and the extent of induction calculated by
normalizing to control activity, i.e. induction equaled activity of treated fish divided by activity of
control fish, and hence had no units. The Mesa crude oil contained sufficient PAH to cause high
levels of EROD induction in trout, as shown by a preliminary bioassay of clean sediments spiked
with oil in the lab (10 mL oil/L of sediment). By diluting the spiked sediment with clean sediment, a
clear exposure-dependent gradient of induction was found (Figure 2.7). The plateau of activity at the
highest oil concentrations suggests acute toxicity (Hodson etal. 1996), likely due to the combined
narcotic effects of all the components of oil. A similar effect was observed when oil was simply
added to water (data not shown). The threshold concentration causing induction was about 0.1 mL
oil/L of reference sediment (Figure 2.7).
43
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100
I
o
&
O
o
a:
LU
10 -
0.1
0.1 1
Oil Concentration (mL/L sediment)
10
Figure 2.7. EROD activity of rainbow trout exposed to reference sediments spiked with oil. Error
bars are 95% confidence limits while the shaded zone represents the 95% confidence limits of
control activity. Numbers represent sample sizes.
Study results showed that PAHs were bioavailable from Conrod's Beach oiled sediments. While
EROD induction was evident for fish exposed to sediments sampled 1 d after oiling, induction
actually increased in July (50 d later) before decreasing somewhat in October (140 d; Figure 2.8).
The initial low extent of induction may have been caused by oil toxicity. With higher oil
concentrations, it is likely that EROD induction was inhibited, as was evident from the leveling-off
of the exposure-response curve in the test of sediments spiked with oil in the lab (Figure 2.7). There
did not appear to be a major effect of bioremediation treatments, with the possible exception of
disking. Plots disked to aerate the sediment showed 45% lower EROD induction potency (p<0.05)
than plots with or without nutrients (Figure 2.9). The plots with plants cut showed 44% lower
induction potency, but the difference was just below the level of statistical significance. Disking may
have reduced induction potency by facilitating the transport of oiled surface sediments into deeper
underlying sediments, enhancing the dispersion of disturbed surface sediments by tides, and
stimulating microbial activity by improving oxygen availability within the wetland sediments (Lee
2000b).
44
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100 H
c
o
D
"D
_C
Q
O
tE
HI
10 -
0 100 200
Days fro m oiling
Figure 2.8. MFO induction in fish exposed to oiled sediments. Each point is the average of data
pooled across all treatments. Error bars = 95% confidence limits. Numbers = sample size.
100i
c
o
is
D
•o
Q
O
10-
1 -
0.1
* *
JJ
Control Control Oil Oil Oil Oil
Nutrients Nutrients Nutrients Nutrients
Cut Disking
Figure 2.9. Effect of experimental treatments on EROD induction of trout exposed to oiled
sediments. Asterisk indicates induction was significantly lower than the highest activity. Error bars =
95% confidence limits. N = 15/treatment.
In summary, because there was a strong link between concentrations of PAHs in beach sediments
and the extent of induction in exposed trout, the induction bioassay successfully tracked the changes
over time in the concentrations and bioavailability of PAH and of the crude oil itself. Over the long
term, we would expect that the relationship between induction and uptake of non-PAH hydrocarbons
would weaken due to the differential rates of degradation and weathering of different components of
oil. However, within the time frame of this study, it appears that both the bioremediation and
phytoremediation treatments did not markedly affect the rate of PAH degradation. While there
appeared to be a significantly greater loss of PAH from aerated sediments, the overall enhancement
was less than two-fold, which is at the limit of detection of the induction bioassay for the sample
sizes tested.
45
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Preliminary studies with eyed-eggs (about 15 d post fertilization) of trout also suggested that
symptoms of Blue sac disease (BSD - characterized by yolk sac and pericardial edema,
hemorrhaging, deformities, and induction of mixed function oxygenase enzymes) were more
frequent in fish exposed to oiled than to un-oiled sediments (Zambon etal., 2000); indicating a risk
to early life stages of species that spawn on tidal beaches.
These laboratory bioassays with fish represent a 'worst-case' scenario as the test organisms could
not avoid exposure to sediments that have been mixed, thereby destroying surface layers that might
be depleted of oil due to weathering. The ratio of water to sediment was also fixed, which is very
different from the situation in well-flushed tidal beaches. Finally, the test species is a useful model,
but is not a beach spawner. Nevertheless, the model fish do provide a useful surrogate for other
species, such as smelt, capelin, and herring that spawn on both freshwater and marine beaches or in
marine littoral zones. In many estuaries, contaminated water is often not well mixed, but moves back
and forth with the tide, thereby causing prolonged exposure of fish entrained in the water mass
(Elson et al. 1972). As illustrated by the Exxon Valdez spill, eggs deposited in beach sediments
cannot move and are also subject to continuous exposure. The utility of freshwater species as a
surrogate for marine species might also be questioned. However, it is worth noting that eggs of pink
salmon, a species closely related to rainbow trout, were exposed to oil from the Exxon Valdez
because they were spawned in river mouth shoals (Marty et al. 1997). The eggs were bathed
alternately in fresh and salt water as the tide rose and fell, so that an exposure of trout to oiled
sediments in saline water is not entirely unrealistic. To resolve the uncertainties associated with
assessing exposure and effects, it is clear that the next step is to refine and adapt bioassays for
application in situ, using species endemic to the test sites.
2.2.5.2 Risk assessment
In this case study, overall sediment quality was determined from the integration of results from
analysis of sediment chemistry, community structure, alteration of primary metabolic processes, and
sediment toxicity. The results of detailed chemical analysis, bioassessments, and bioassay tests
suggested that in the Conrod Beach study, natural attenuation was the primary process that reduced
residual oil concentrations and toxic effects. The biotest results showed that the remediation
strategies under evaluation, stimulation of bioremediation and phytoremediation activity by nutrient
amendments and physical mixing (disking), were not highly effective in regards to restoration. It
was also evident in the results of some biotests (e.g. Amphipod Survival Test, Microtox Test) that
possible detrimental effects may be linked to the addition of fertilizers.
It is imperative that one fully understands the various processes that may affect biotest endpoints.
Failure of the Gastropod Survival Test to resolve differences in experimental treatments could be
attributed to adverse environmental conditions that caused high levels of inherent variability within
the test matrix. The Biotox Solid-Phase Flash Assay (Lappalainen et al, 1999) is currently being
considered as a relevant adjunct (or alternative) test to the Microtox solid-phase assay, since it
allows the evaluation of large numbers of environmental samples at a more reasonable cost using the
same test organism. This assay was used with success to provide evidence of toxicity reduction by
remedial activities in an oil-impacted freshwater environment (Blaise et al., 2002). However, in this
case study, all sediments collected during the first 2 sampling events (Week 0,2) show marked and
46
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similar levels of toxicity. It is hypothesized that the assay was unable to discriminate between
toxicity of oiled and unoiled sediments, as it was responding to the presence of natural contaminants
present in the anaerobic sediments (e.g., NH3 and/or H2S) and possibly also to their degree of oil
contamination (in the case of the oiled sediments). Similarly, the Algal Toxicity Test with
Selenastrum capricornutum (Blaise and Menard, 1998) that readily identified the inhibitory effects
of residual oil in sediments on esterase enzyme activity in the previous freshwater marsh case study
was ineffective in this test case. This was attributed to interferences associated with benthic diatom
growth. This has necessitated the development of a new algal toxicity test using the marine macro-
alga Champiaparvula to assess archived sediments from this study.
With further refinement, guidelines for selection of bioassessment and bioassay test suites will be
provided to oil spill response managers that are tasked to implement and verify the success of
countermeasure operations including the extent of habitat recovery. For this case study in a marine
salt marsh environment, the results of the ecological risk assessment with all available quantitative
chemical and biological data suggest that natural attenuation may be the most environmentally sound
and cost-effective treatment option. Although there was some evidence of changes in microbial
community structure and activity, no significant differences were observed among treatments in oil
degradation rates or toxicity reduction. Active remedial treatment is not supported by cost-to-benefit
analyses.
2.3 Summary and Recommendations
Most of the information presented in this guidance document was based on only a few field studies
of oil bioremediation. Not many studies have been done in a definitive manner. The Conrod Beach
experiment in Dartmouth, Nova Scotia, demonstrated that biodegradation of the alkane fraction and
some of the PAH fraction was stimulated following the application of inorganic fertilizers directly to
the plots. Disking (or tilling) caused substantial damage to the rhizosphere, and such drastic
measures cannot be recommended as a means of increasing oxygen content in the root zone. Not
much can be done in that regard. Thus, if significant penetration has taken place into the subsurface,
then not much hope of acceleration in hydrocarbon disappearance is possible since anaerobic
conditions rapidly set in at greater depths. If, however, penetration is limited to the top several mm,
then sufficient oxygen might be available to permit biostimulation to accelerate greater hydrocarbon
disappearance than via natural attenuation. So, of major importance in the event of an oil spill in a
salt marsh (or any wetland oil spill) is to assess the degree to which penetration has taken place
below the surface. If it is minor, then biostimulation could be considered as a viable strategy for
cleanup. If it is more than a few mm penetration, then biostimulation will have diminished
effectiveness due to the increased likelihood of limiting oxygen concentration in the oil impact zone.
Salt marshes are among the most sensitive ecosystems and, therefore, the most difficult to clean.
Applications of some traditional oil spill cleanup techniques in wetland habitats have caused more
damage than the oil itself. Several long-term field studies have been carried out in coastal wetlands
to evaluate the potential of oil bioremediation, one of the least intrusive technologies. The studies
have shown that oil biodegradation on coastal wetlands is often limited by oxygen, not nutrient
availability. Natural attenuation is increasingly becoming the preferred strategy for the restoration of
oil-contaminated wetlands. However, field studies also indicate that nutrient amendments may still
be a viable option for removing hydrocarbons from an oil-contaminated wetland if the oil does not
47
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penetrate deeply into the anoxic zone of wetland sediments and when nutrients are limiting. When
biostimulation is selected, it is recommended that nitrogen concentrations of at least 2 to as much as
10 mg N/L should be maintained in the pore water to achieved optimal oil biodegradation, with the
decision on higher concentrations to be based on a broader analysis of cost, environmental impact,
and practicality. The overall success of the remedial operations should be not only based on the
efficiency of oil degradation but also the integration of results from a suite of assays, which are
chosen on the basis of ecological relevance to the site of concern, cost considerations, and the
availability of technical expertise.
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