Sediment Management at Superfund Megasites
A variety of subjects including environmental engineering, toxicology, environmental monitoring, human and environmental risk assessment, and risk management are relevant to evaluating remediation at contaminated sediment Superfund sites. In this chapter, a number of issues are briefly introduced to provide background for later discussions. Topics include the Superfund process and information available on contaminated sediment Superfund sites; evaluating and managing risks posed by contaminated sediments; and techniques for managing and remediating contaminated sediment with a focus on dredging technologies and their performance capabilities and limitations. The chapter is intended to provide a cursory overview of the topics while emphasizing other sources containing more detailed discussions.
OVERVIEW OF SUPERFUND AND SEDIMENT MEGASITES
Superfund and Environmental Remediation
In 1980, Congress enacted the Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA, 42 U.S.C. 9601-9675),
which authorized the establishment of the Superfund program. The goal of the program is to reduce current and future risks to human health and the environment at sites contaminated with hazardous substances. CERCLA established a wide-ranging liability system that makes those responsible for the contamination at sites liable for cleanup costs (see Probst et al. 1995 for greater detail). It also created the “Superfund,” a trust fund stocked primarily by a dedicated tax on oil and chemical companies, to fund cleanup activities where there was no financially viable responsible party. Since the taxing authority expired in 1995, the trust fund is largely depleted, and Congress now funds the program from general revenues through annual appropriations (Fletcher et al. 2006).1 The U.S. Environmental Protection Agency (EPA) implements the program through the National Oil and Hazardous Substances Pollution Contingency Plan (40 CFR § 300), commonly referred to as the NCP or the national contingency plan.
Most of the Superfund program’s efforts are aimed at cleaning up sites on the National Priorities List (NPL). Typically, a site is proposed for inclusion on the NPL after being evaluated with a hazard-ranking system, which assesses the potential for hazardous-substance releases at a site to harm human health or the environment (40 CFR § 300 Appendix A). The Superfund process progresses from an initial site assessment through cleanup and eventually deletion of the site from the NPL. Site activities can be paid for by EPA (known as “fund-led” cleanups),2 by parties connected to the site (referred to as responsible parties), or by some combination of the two.
Selection of a remedy begins with a remedial investigation and feasibility study (RI/FS). The RI is intended to determine the nature and extent of contamination and estimate the associated risk to people and the environment. The FS analyzes and compares remedial alternatives according to the nine NCP criteria (Box 2-1). The criteria require that the remedy, above all, be protective of human health and the environment and comply with all applicable or relevant and appropriate requirements
Evaluation Criteria for Superfund Remedial Alternatives
Before a remedial strategy is selected for a Superfund site, the options are evaluated on the basis of nine criteria (see below). The first two, overall protection of human health and the environment and compliance with applicable or relevant and appropriate requirements (ARARs), are termed threshold criteria, and a potential remedy must meet them to be selected as a final remedy.3 The next five criteria are termed balancing criteria and are used in weighing the advantages and disadvantages of potential remedies. The final two criteria are modifying criteria, and the agency is supposed to take them into consideration as part of the selection process.
Source: Adapted from EPA 2005a.
(ARARs).4 Remedies are also compared on whether they are technically feasible and cost-effective, provide long-term (permanent) effectiveness, and minimize deleterious effects and health risks during implementa-
tion. There is a preference for remedies that can reduce the toxicity, mobility, and volume of contaminants. Finally, there is a preference for remedies that have state and community support.
EPA uses the FS to identify each alternative’s strengths and weaknesses and the trade-offs that must be balanced for the site in question (EPA 1988). The agency then selects a remedy and describes it in a record of decision (ROD). Additional studies may be conducted to support the design of the remedy. Once constructed and implemented, the remedy is maintained and monitored to ensure that it achieves its long-term goals. EPA may delete a site from the NPL when a remedy has been implemented, the cleanup goals have been achieved, and the site is deemed protective of human health and the environment (EPA 2000).
If, after implementation of a remedy, contamination exists that could limit potential uses of the site, the site is subject to 5-year reviews even if it has been deleted from the NPL (EPA 2001). The reviews are intended to evaluate the performance of the remedy in protecting human health and the environment and are to be based on site-specific data and observations. However, monitoring is not limited to sites where 5-year reviews are required. EPA guidance states that “most sites where contaminated sediment has been removed also should be monitored for some period to ensure that cleanup levels and RAOs [remedial action objectives] are met and will continue to be met” (EPA 2005a, p. 2-17). Post-remediation monitoring (required in conjunction with 5-year reviews or otherwise) is the basis for evaluating remedy effectiveness and adapting remedial strategies and risk management to achieve remedial action objectives (for further discussion, see Chapter 5).
Sediment Contamination at Superfund Sites
Contaminated sediment is a widespread problem in the United States (EPA 1994, 1997, 1998, 2004a, 2005a). Its wide distribution results from the propensity of many contaminants discharged to surface waters to accumulate in sediment or in suspended solids that later settle. Contaminants can persist in sediment over long periods if they do not degrade (for example, metals) or if they degrade very slowly (for example, polychlorinated biphenyls [PCBs] or polycyclic aromatic hydrocarbons
[PAHs]). Historically contaminated sediment can become buried or, if it is resuspended, can settle out eventually and lie on the sediment surface.
At the national level, the geographic extent of areas with contaminated sediment is not fully defined. In the 2004 Contaminated Sediment Report to Congress (EPA 2004b), EPA reported on sediment sampling at 19,398 sampling stations nationwide, located in about 9% of the water-body segments in the United States. Of that nonrandom sample of sediment sampling stations, EPA classified 43% as having probable adverse effects, 30% having possible adverse effects, and 27% as having no indications of adverse effects. The 2005 Contaminated Sediment Remediation Guidance for Hazardous Waste Sites (EPA 2005a) cites EPA fish advisories covering all five Great Lakes, 35% of the nation’s other lakes, and 24% of total river miles as due partly to sediment contamination (EPA 2005b).
EPA does not maintain a current list of NPL sites with contaminated sediments, nor does it compile a list of contaminated sediment areas that are potential Superfund sites. It also does not maintain a list of contaminated sediment sites that are being (or have been) remediated under another authority. EPA did report that “as of September 2005, Superfund has selected a remedy at over 150 sediment sites” (EPA 2006a). In addition, the EPA Office of Superfund Remediation and Technology Innovation is tracking progress at 66 sites, termed tier 1 sites, where the sediment-cleanup remedy involves more than 10,000 cubic yards (cy) of sediment to be dredged or excavated or more than 5 acres to be capped or monitored for natural recovery (EPA 2006b).5 Of the aforementioned 150 NPL sites where remedies have been selected, EPA considers 11 to be sediment megasites, defined as sites where the sediment portion of the remedy is expected to cost $50 million or more.6 Of these 11 sites, 10 were proposed for inclusion on the NPL in the very early years of the Superfund program (in 1982-1985), and one (Onondaga Lake) was proposed for inclusion in 1993. Thus, the overwhelming ma-
jority of the megasites have been on the NPL for over 20 years. Only one of the 11, Marathon Battery, has been formally deleted from the NPL. In addition to the 11 megasites on the NPL, EPA lists two megasites that have been proposed for the NPL but are not final (GE Housatonic River, MA and Fox River, WI) and one that has not been proposed (Manistique River/Harbor area, MI). The 14 sites are listed in Table 2-1. The status of remediation at the sites varies. At some, such as Bayou Bonfouca and Marathon Battery, remediation has been completed; at others, such as Commencement Bay and Sheboygan Harbor, remedial activities are going on; and at still others, such as Hudson River and Onondaga Lake, remedial activities have not begun. Megasites are described only in terms of remediation cost (at least $50 million), so the size and volume of contaminated materials at the sites can vary greatly (see Box 2-2).
One might ask, Why all this attention to contaminated sediment megasites if there are only 14 nationwide? There are two reasons. First, at 13 of the megasites mentioned above (no cost information was provided on the Triana/Tennessee River site), total remedial costs are estimated to be about $3 billion, a huge amount of money even by Superfund standards.7 Second, the 14 sites probably constitute only a subset of the contaminated sediment sites that will entail expensive remedies and will be cleaned up under the Superfund program. For example, the EPA list of contaminated sediment megasites does not include some well-known sites, such as the Bunker Hill Mining and Metallurgical Complex, ID, and Love Canal, NY. Both those tier 1 sites are megasites by the conventional definition (total remediation cost of at least $50 million), but the sediment portion alone is not expected to be $50 million.
When comparing EPA’s list of tier 1 sites (EPA 2006b) with a somewhat dated list of megasites8 (that does not include federal facilities), one can find 11 “conventional” megasites on the tier 1 list.
Alcoa–Point Comfort/Lavaca Bay, TX
Allied Paper Inc./Portage Creek/Kalamazoo River, MI
Bunker Hill Mining and Metallurgical Complex, ID
TABLE 2-1 Sediment Megasites (Sites at Which Remediation of the Sediment Component Is Expected To Be at Least $50 million)
Site Name, State
New Bedford Harbor, MA
Hudson River PCBs, NY
Marathon Battery Corp., NY
Onondaga Lake, NY
Triana/Tennessee River, AL
Sheboygan Harbor and River, WI
Velsicol Chemical, MI
Bayou Bonfouca, LA
Milltown Reservoir Sediments, MT
Silver Bow Creek/Butte Area, MT
Commencement Bay, WA
GE Housatonic River, MA
Fox River, WI
Manistique River/Harbor area, MI
Source: EPA, unpublished data, “$50M Cost query_091306.xls,” Sept. 18, 2006.
Eagle Mine, CO
EI duPont–Newport landfill, DE
GM–Central Foundry Division (Massena), NY
Lipari landfill, NJ
Love Canal, NY
Nyanza chemical waste dump, MA
McCormick and Baxter Creosoting Co., CA
Wyckoff Co.–Eagle Harbor, WA
Furthermore, as described below, large and expensive sediment remediations are conducted under authorities other than Superfund.
A crucial question is how many additional major contaminated sediment sites are likely to be listed on the NPL. EPA does not designate “likely future megasites” in its tier 1 list of sites or NPL sites for which RODs have not been issued. According to EPA, the most likely future sediment megasites are the “tier 2” contaminated sediment sites (S. Ells,
How Large Is a Megasite?
Contaminated sediment megasites are among the most challenging and expensive sites on the NPL. Megasites are conventionally defined as those with remedial activities costing at least $50 million, but there are large differences in the magnitudes and scales of these sites. A few megasites, such as Bayou Bonfouca and Marathon Battery, are relatively small, with dredging activities covering tens of acres and operations occurring over a few years. Other dredging projects—such as those in the Fox River, New Bedford Harbor, and Commencement Bay—are components of broader activities at large-scale megasites where remedial activities are going on and will take years or decades to complete. The $50-million distinction for a megasite is not readily translatable into volume of materials removed. For example, sediment remediation (including design, mobilization, marine demolition, dredging, water management, transportation and disposal, construction oversight and EPA oversight, without the upland-based removal costs) at the Head of Hylebos Waterway in Commencement Bay, WA, removed 404,000 cy at a cost of $58.8 million (about $145/cy) (P. Fuglevand, personal commun., Dalton, Olmsted & Fuglevand, Inc., May 11, 2007). In Manistique Harbor, MI, dredging operations removed 187,000 cy at a cost of $48.2 million (about $260/cy) (Weston 2002). Dredging operations in Bayou Bonfouca, LA, removed 170,000 cy at a cost of $90 million (about $530/cy) (EPA, unpublished information, “$50M Cost query_091306.xls,” Sept. 18, 2006).
EPA, personal commun., Oct. 12, 2006). Tier 2 sites are designated for review by the Contaminated Sediments Technical Advisory Group because they are large, complex, or controversial contaminated sediment Superfund sites.9 There are 12 tier 2 sites. Three are on the earlier two lists provided, but nine are not. Of the nine, four are NPL sites (Ashland/Northern States Power, WI, Portland Harbor, OR, Lower Duwamish Waterway, WA, and the Pearl Harbor Naval Complex, HI), and five are not (Palos Verdes, CA, Kanawah River/Nitro, WV, Centredale Manor Restoration Sites, RI, Anniston PCB site, AL, and Upper Columbia River, WA). EPA also indicates that the Passaic River, NJ, Berry’s Creek at Ventron/Velsicol, NJ, and Tar Creek, OK, are likely future con-
Although, it should be noted that EPA indicates that “No quantifiable criteria were used to develop this list.” The list of sites is available at http://www.epa.gov/superfund/resources/sediment/cstag_sites.htm.
taminated sediment megasites, although they have not been designated as tier 2 sites (S. Ells, EPA, personal commun., Sept. 18, 2006).
Because predicting future NPL listings is more an art than a science, in some ways, it is not surprising that there is no official list of likely future contaminated sediment megasites. That said, the committee was surprised that there is so little effort devoted to tracking and understanding likely future sediment megasites at the national level. Apparently, fewer than two full-time employees are assigned to contaminated sediment issues at Superfund headquarters. It appears that EPA has not allocated the resources needed to identify the scope of the problem and to develop a strategy to address issues related to contaminated sediments. To develop an effective long-term contaminated sediment strategy it is critical to know how much work remains to be done. To address that question, one needs to have three pieces of information:
How much work remains at sites already categorized as contaminated sediment megasites.
How many contaminated sediment sites already on the NPL are likely to be determined to be megasites.
How many new such sites are likely to be added to the NPL in the coming years.
None of that information is readily available from EPA. Clearly, EPA should not stop and wait until this information is collected. However, it is important that EPA obtain this information and update it regularly in order to be able to forecast likely future costs and needed resources, as well as to assess what kinds of research and monitoring improvements are likely to have the largest benefit to the program.
Cleanup Under Authorities Other Than Superfund
Remediation of contaminated sediments is also conducted under authorities other than Superfund and can be led by various parties, such as state or federal agencies or private entities, in combination or individually. For example, a 5-mile reach of the Grand Calumet River, a highly industrialized tributary to Lake Michigan in northwest Indiana, was dredged by U.S. Steel Corporation pursuant to a Clean Water Act
consent decree and a Resource Conservation and Recovery Act corrective-action consent order (Menozzi et al. 2003). This project, described as “the largest environmental dredging project to be undertaken in North America,” removed 786,000 cy of sediment from the Grand Calumet River (U.S. Steel 2004).
State programs conduct and oversee sediment remediation under a variety of authorities. For example, the State of Washington Department of Ecology is charged with cleaning up and restoring contaminated sites under authority of the Model Toxics Control Act (MTCA) and Sediment Management Standards (SMS) (Washington Department of Ecology 2005). In 2005, 142 sediment cleanup sites were identified in Washington: 41 were being cleaned up under federal authorities, 48 were using state authority alone, 11 were under federal and state authorities, and the remaining 42 were either voluntary (conducted by the responsible party) or the authority had not been assigned (Washington Department of Ecology 2005).
Contaminated sediments in many harbors and rivers of the Great Lakes are addressed in the Great Lakes Water Quality Agreement between the United States and Canada, which established 43 areas of concern (AOCs) in U.S. and Canadian waters. The U.S. EPA Great Lakes National Program Office administers funds from the Great Lakes Legacy Act of 2002 for the remediation of contaminated sediment at AOCs (EPA 2004c). The first Legacy Act cleanup was in 2005 at the Black Lagoon in the Detroit River AOC near Trenton, MI. At that site, 115,000 cy of contaminated material was dredged, and the area was capped. Hog Island, near Superior, WI, in the St. Louis River AOC of Lake Superior, was remediated with dry excavation (see Sediment Management Techniques in this chapter for a description of remedial methods). In 2006, two projects were under way with Great Lake Legacy Act funds. The Ruddiman Creek remedial action in Muskegon, MI, contains an excavation and dredging component and is expected to remove around 80,000 cy. Dredging will also occur at the Ashtabula River, near Cleveland, OH, where it is expected that about 600,000 cy of contaminated sediment will be removed from the lower portion of the river.
Another program, the Urban Rivers Restoration Initiative, is a collaboration between EPA and the U.S. Army Corps of Engineers for urban-river cleanup and restoration (EPA 2003a). Eight demonstration pilot projects, including a dredging project in the Passaic River in New
Jersey, have been developed to coordinate the planning and implementation of projects to promote clean water and sediment among multiple jurisdictions and federal authorities.
This section is by no means a comprehensive listing of sediment projects or efforts outside of Superfund; rather, the intent is to convey that there are many sediment remediation projects outside of Superfund that are conducted by multiple groups and under several authorities. To the extent that other environmental dredging activities are conducted to address risk from contaminated sediments, many of the discussions and conclusions presented in the latter chapters of this report will be applicable.
EVALUATING RISK REDUCTION AT CONTAMINATED SEDIMENT SITES
Risks Posed by Contaminated Sediment
As briefly described in Chapter 1, contaminants in sediment can pose risks to human health and the environment. Apart from direct exposure to contaminated sediment during, for example, recreational activities, humans typically are exposed to contaminants through the ingestion of fish or wildlife that have accumulated contaminants from the sediment. Fish and wildlife are exposed to contaminants in sediments through a number of pathways, including absorption from pore water or sediments, incidental ingestion of contaminated sediments, and consumption of contaminated organisms. Several of those processes are presented graphically in Figure 2-1. Predicting effects of exposures can be complex. Variations in the sediment environments will alter the bioavailability of contaminants, and this can markedly affect their accumulation and effects on organisms (NRC 2003). For instance, the presence of sulfide will greatly decrease the bioavailability of many metals, and organic carbon can decrease the bioavailability of organic pollutants, such as PCBs (EPA 2003b, 2005c).
Accessibility is a primary factor in exposure to and effects of contaminated sediments. A common problem in assessing risks posed by contaminated sediments is that the contaminants (or the highest contam-
inant concentrations) can be buried beneath relatively clean sediment that has deposited over time (see Figure 2-2).
Because sediment contaminants typically are strongly associated with the sediment particles, contaminants buried below the biologically active zone are neither accessible nor available to sediment- or water-dwelling organisms. In such cases, a relatively small continuing source may pose a greater risk of exposure and associated injury than a large buried inventory of sediment associated contaminants. Risk due to sediments is usually limited to contaminants that are present in or can migrate into the biologically active zone, the upper layers of sediment where organisms live or interact. That layer typically ranges from a few centimeters to 10-15 cm deep, although some organisms (including aquatic plants) may penetrate more deeply (Thoms et al. 1995; NRC 2001).
Sediment-associated contaminants tend to collect in relatively stable depositional zones in water bodies. In such environments, buried contaminants (that is, those below 10-15 cm) may never be exposed to the biologically active zone. However, water bodies are dynamic systems and even in generally depositional and stable environments, high flow events and changes in hydrologic conditions may lead to short term
erosion, exposure, and transport of these contaminants to the biologic active zone. In environments subject to such conditions, removal of contaminant mass may be an effective remedial response to the risk posed by them. If contaminants buried below the biologically active zone are likely to remain buried, the potential exposure and risk may be so small that remediation of any kind is unwarranted. Remediation of deeply buried contaminated sediments that do not contribute to the exposure of aquatic systems now or under future conditions will not achieve risk reduction goals. In such cases, other contaminant sources, for example inadequately controlled surface discharges or atmospheric deposition, may control exposure and risk. A fair amount of effort in recent years has gone into developing approaches for assessing sediment column stability and refining hydrodynamic models and linking them with fate and transport models to estimate contaminant transport under various condi-
tions (e.g., Bohlen and Erickson 2006). Output from these approaches and models are important in estimating the risks associated with remedial alternatives.
Decision-Making in a Risk-Based Framework
Principles for understanding and comparing risk reduction from various sediment remediation techniques are discussed briefly below, however, it should be noted that it is not the mandate or intent of the report to develop specific recommendations and procedures for performing comparative risk analyses of remedial alternatives in selection of a sediment remedy. While important, that type of detailed assessment was not requested or undertaken. The brief discussion provided here on risk-based remedy selection is intended to provide background for later discussions on improving decision making.
The process of managing risk at contaminated sediment sites was evaluated extensively in the 2001 National Research Council report A Risk Management Strategy for PCB-Contaminated Sediments (NRC 2001). Perhaps its most relevant conclusion is that all decisions regarding the management of PCB-contaminated sediments should be made in a risk-based framework. The report further suggests that the framework developed by the Presidential/Congressional Commission on Risk Assessment and Risk Management provides a good foundation for assessing the risks and the management options for a site (see Box 2-3). The general framework exhibits several key features that make it appropriate for the management of contaminated sediment sites. It recognizes that risk-reduction should be the foundation of any decision-making process and the importance of the participation of interested and affected stakeholders in the decision-making process. It also provides a systematic and structured process for identifying and assessing risks, evaluating and implementing options, and monitoring the success of the overall process. The 2001 Research Council report also recommends that risk assessments and risk management decisions be site specific and concluded that current management options can reduce risks but cannot eliminate PCBs and PCB exposure from contaminated sediment sites. Because all remedial options will leave residual PCBs, the short- and long-term risks that
they pose should be considered in evaluating management strategies. Those ideas also apply to other sediment contaminants.
Presidential/Congressional Commission on Risk Assessment and Risk Management
The Presidential/Congressional Commission on Risk Assessment and Risk Management was formed in response to the 1990 Clean Air Act Amendments in which Congress mandated that a Commission on Risk Assessment and Risk Management be formed to “make a full investigation of the policy implications and appropriate uses of risk assessment and risk management in regulatory programs under various Federal laws to prevent cancer and other chronic human health effects which may result from exposure to hazardous substances” (PCCRARM 1997, p. i).
The commission ultimately developed a report that introduced a risk management framework “to guide investments of valuable public sector and private sector resources in researching, assessing, characterizing, and reducing risk” (PCCRARM 1997, p. i). The commission proposed a six-stage process:
This process should be conducted
The proposed process, depicted in the schematic below, is a systematic method to manage risks that the commission defined as “the process of identifying, evaluating, selecting, and implementing actions to reduce risk to human health and to ecosystems. The goal of risk management is scientifically sound, cost-effective, integrated actions that reduce or prevent risks while taking into account social, cultural, ethical, political, and legal considerations” (PCCRARM 1997, p. 2).
Remedy selection is a complex process with many considerations (see Box 2-1). In some cases, removal will be the best option for risk reduction and satisfying the NCP criteria, in others, capping or monitored natural recovery will be preferable. An analysis of alternative remedies typically includes a comparison of both the short- and long-term risks to human and environmental receptors associated with a particular site. For example, the risks from dredging can include exposure to contaminants during dredging, rehandling, and transport, and contaminants that remain after operations are completed. Those risks would be compared to other alternatives, including risks from unconfined contaminated sediment and potential future resuspension and transport during storm and non-storm events. Risks beyond those related directly to exposure to contaminants are also considered in this process (see Net Risk Reduction below).
Technical and policy guidance for making remedial decisions using a risk based framework at contaminated sediment sites was recently issued (EPA 2005a). The document provides a useful evaluation of the various sediment management approaches and their advantages and limitations in attaining risk reduction. It discusses in detail aspects of the
Superfund decision-making process (site characterization, feasibility study, and remedy selection) particular to contaminated sediment, and it offers recommendations for implementing an effective monitoring plan. The guidance concludes that “The focus of remedy selection should be on selecting the alternative best representing the overall risk reduction strategy for the site according to the NCP nine remedy selection criteria…. EPA’s policy has been and continues to be that there is no presumptive remedy for any contaminated sediment site, regardless of the contaminant or level of risk” (EPA 2005a, p. 7-16).
Estimating the degree of risk reduction is central in considering the potential effectiveness of a remedial action. Risk posed by chemical contamination is a function of the duration and intensity of exposure and the ability of the chemical or chemical mixture to exert adverse effects. There is not a direct measure of risk, so surrogate metrics are used to estimate risk. Environmental analyses have to use metrics that, in practice, can be employed relatively easily, are not time- and cost-prohibitive, and have sufficient accuracy and precision to be reliable. Estimates of risk reduction at contaminated sites have often centered on measuring changes in the mass, volume, and concentration of contaminated sediments. Those measures are related to the potential for exposure in the aquatic environment but do not provide information on effects. Therefore, although they are the most prevalent, they are not fully adequate to describe risk or to chart risk reduction. Toxicity testing, biologic community indexes, and tissue-residue analyses provide a fuller picture of effects, although they too have limitations in their ability to describe risk. (See Chapter 5 for further discussion on the metrics and their advantages and disadvantages in estimating risk to aquatic biota and humans.)
In characterizing risk and evaluating risk reduction, it is necessary to consider the duration of time over which exposure and effects occur. After remediation, risk is usually predicted to decline over time rather than reach a protective level immediately on completion of the remedy,
so remedy selection involves a comparison of time profiles of predicted risks, often on scales of decades. Such a comparison of risk profiles over time is how the long-term effectiveness of dredging is evaluated relative to alternative technologies for the largest and most complex sites. Factors dictating the time to reduce risk are site specific and include the time required to design and fully implement the remedy, the time required to cleanse the food chain of existing contaminant body burdens, and the time for natural recovery processes to attenuate any residual surface sediment concentrations after implementation is complete.
Net Risk Reduction
The 2001 National Research Council report indicated that the paramount consideration for contaminated sites should be the management of overall or net risks to humans and the environment in addition to specific risks. The report concludes that the evaluation of sediment management and remediation options should take into account all costs and potential changes in risks for the entire sequence of activities and technologies that constitute each management option. (For example, managing risks from contaminated sediments in aquatic environments might result in the creation of additional risks in both aquatic and terrestrial environments.) The report also suggests that a broader array of risks—including societal, cultural, and economic risks—should be evaluated comprehensively. The concept of net risk reduction has been embraced by EPA in its Contaminated Sediment Remediation Guidance for Hazardous Waste Sites (EPA 2005a); it states that “Project managers are encouraged to use the concept of comparing net risk reduction between alternatives as part of their decision-making process for contaminated sediment sites, within the overall framework of the NCP remedy selection criteria. Consideration should be given not only to risk reduction associated with reduced human and ecologic exposure to contaminants, but also to risks introduced by implementing the alternatives. The magnitude of implementation risks associated with each alternative generally is extremely site specific, as is the time frame over which these risks may apply to the site. Evaluation of both implementation risk and residual risk are existing important parts of the NCP remedy selection process” (EPA 2005a, pp. 7-13, 7-14).
Risk-Based Objectives and Cleanup Levels
Each site has its own set of contaminants with different concentrations and distributions in its own particular geologic, geochemical, geographic, social, ecologic, and economic setting. Therefore, management decisions based on the above framework are expected to differ among sites.
At Superfund sites, the overall goal of sediment management is reduction of risk to human health and the environment. That goal takes the form of remedial action objectives, which are used in developing and comparing alternatives for a site, and typically describe the desired effect of the remediation on risk (for example, reduction to acceptable levels of the risks to people ingesting contaminated fish). Attainment of remedial action objectives can be difficult to quantify or might occur in a time frame or encompass a spatial scale that makes it difficult to link to remedial actions. Under such circumstances, cleanup levels, such as achievement of a sediment concentration or removal of a given mass of contaminant or sediment, which can be more easily used to evaluate remedial actions, are often adopted (EPA 2005a). Ideally, clean-up levels are tied to effects-based risk thresholds, and take into account effects of combinations of contaminants.
That the application of a good risk management strategy is likely to result in significantly different cleanup levels at different sites makes it difficult for the committee to draw conclusions about the expected effectiveness of dredging. At some sites, cleanup levels are far less stringent than at others, and thus all other things being equal, a site with less stringent cleanup goals is more likely to be “successful” than a site with more stringent goals. Geologic and site-specific conditions also differ, so even if the cleanup goals are similar at different sites, the technical ability to reach the goals may differ. Thus, one needs to be highly cautious in suggesting that success with a remedial option at one site necessarily means that the same success is likely at another.
To help to ensure that remedial actions achieve their desired objectives, 11 principles have been developed by EPA to guide sediment remediation (Box 2-4). The principles were developed partially in response to the recommendations of the National Research Council Committee on Remediation of PCB-Contaminated Sediments (NRC 2001).
Eleven Principles of Contaminated Sediment Management
In February 2002, the EPA Office of Solid Waste and Emergency Response promulgated 11 principles of contaminated sediment management (OSWER Directive 9285.6-08):
The principles were designed to help EPA site managers to make scientifically sound and nationally consistent risk-management decisions at contaminated sediment sites. The principles are consistent with the recommendations of the National Research Council A Risk Management Strategy for PCB-Contaminated Sediments and the Presidential/Congressional Commission on Risk Assessment and Risk Management. They were incorporated into the Contaminated Sediment Remediation Guidance for Hazardous Waste Sites (EPA 2005a).
The Conceptual Site Model: A Working Understanding of Processes Leading to Risk from Sediment Contamination
The development of remedial action objectives and cleanup levels to reduce risk is based on a conceptual understanding of cause-effect
relationships among contaminant sources, transport mechanisms, exposure pathways, human receptors, and ecologic receptors at each affected level of the food chain. That understanding of causal relationships is known as a conceptual site model (CSM) (EPA 2005a) and is typically developed for each site on the basis of site-specific conditions. The link between risk and the inventory of contaminants in sediments is not always obvious. An accurate CSM is critical for identifying the processes and pathways that might lead to risk and appropriate means of intervening to reduce risk. For example, evaluation of the stability or potential instability of buried deposits and their potential for exposure in the bioavailable zone is a key component of a CSM because a CSM must be able to differentiate between important and unimportant routes of exposure.
In addition to linking site contaminant sources to exposures and risks, the CSM must account for background conditions, including contaminant distribution from offsite sources. Ecosystems may be highly stressed because of multiple watershed and atmospheric effects on conventional water quality measures, such as nutrients, suspended solids, acidity, dissolved oxygen, and temperature. The contribution of background stressors or other background sources to site effects should be evaluated, including assessment of their importance relative to the contaminants of concern, and recognized as potentially complicating factors in ecosystem restoration.
The CSM should guide site investigation, and its hypotheses and assumptions should be tested and refined as site data are acquired. When the CSM has been accepted with a high degree of confidence, it is used to define remedial action objectives. Basing remedial action objectives on the best scientific understanding of the mechanisms that lead to site-specific risk maximizes the likelihood that remedial actions will meet the objectives.
For most sites, but especially for the largest and most complex, a quantitative dimension must be added to the CSM to support development and selection of a remedy. The result is a mathematical model or a set of models of the various component processes. Mathematical models are used to quantify the same cause-effect relationships that are embodied in CSMs so that magnitudes of predicted outcomes can be associated with specific causes or actions, such as contaminant loads, environmental conditions, and remedies. To be most accurate, the models
should be supported by and calibrated to site-specific data on the environmental media (such as sediments, pore water, and water), receptors (such as benthic organisms, fish, and humans), and processes (such as toxicity, bioavailabity, and bioaccumulation) that are being examined.
Mathematical models range from simple to complex, including analytic equations representing established scientific relationships between independent and dependent variables, statistical cause-effect relationships between site variables, and systems of differential equations representing multiple fate and transport processes. With a mathematical model, quantitative versions of hypotheses can be tested and refined on the basis of site data, including data from field surveys of site conditions and pilot studies of remedial technologies, and then the relative effectiveness of alternative remedies in reducing exposures can be estimated, including the sensitivity of exposure to the remedies and the time needed to reduce exposure. The measures of predicted effectiveness are used to support remedy selection.
Models are subject to uncertainty because of the uncertainty in parameters and process representations. Model testing and refinement does not end with the selection of a remedy through a record of decision. It is important that the conceptual and mathematical site models also be used in designing monitoring of conditions during implementation and post-remedy phases and that the monitoring data be used to validate the models’ predictions. When risk reduction deviates significantly from a model’s predictions, the model should be modified or recalibrated to improve its accuracy so that more reliable predictions can be available to guide midcourse adjustments in the remedy (EPA 2005a).
CONTAMINATED SEDIMENT MANAGEMENT TECHNIQUES
Contaminated sediment is managed with various techniques, including source control, natural recovery, capping, and removal (dry excavation and dredging). Removal necessitates management of the removed material, which normally includes dewatering, transport, and disposal. Treatment of dredged material to remove or destroy contaminants is an option, but cost and other factors usually lead to disposal in upland landfills or in near-shore confined disposal facilities. In some
cases, dredged material can be returned to the aquatic environment through containment in confined aquatic disposal facilities (EPA 2005a).
The National Research Council Committee on Contaminated Marine Sediments (NRC 1997) and Committee on Remediation of PCB-Contaminated Sediments (NRC 2001) have reviewed and reported on a number of sediment management techniques. The committees stated that source control is advisable in all contaminated sediment management projects, notwithstanding the difficulties of identifying some sources of contamination. Beyond source control, interim controls (temporary measures to address exposures immediately) and long-term controls (such as in situ management technologies, sediment removal and transport, and ex situ management) may be needed to address sediment contamination.
More recently, EPA (2005a) lists both in situ and ex situ remedial strategies for managing risks posed by contaminated sediment. The in situ strategies include monitored natural recovery (MNR), in situ capping, hybrid (thin-layer placement) approaches, institutional controls, and in situ treatment; the ex situ strategies include dredging and dry excavation (following dewatering or water diversion). See Box 2-5 for an explanation of these approaches. The present committee’s focus is on environmental dredging, which is conducted specifically to remove contaminated sediments, as opposed to navigational dredging, which typically is intended to maintain depth in waterways for navigation or other purposes.
HISTORICAL PERSPECTIVE ON THE USE OF REMOVAL TECHNOLOGIES TO REDUCE RISK
Although there has never been a presumptive remedy for sediments, the historical preference for removal is evident in the large percentage of sites whose remedy was based entirely or in part on dredging. In an overview of Superfund sediment remediation, EPA presented information from 60 tier 1 sites for which a remedy had been selected. Of the 60, 57% had only removal as the remedial action, 15% capping with removal, 13% removal with MNR, 5% only capping, 2% only MNR, and 8% all three remedies (Southerland 2006). The historical preference for
Remedial Approaches to Contaminated Sediment In Situ Approaches
In Situ Approaches
Ex Situ Approaches
Source: Adapted from EPA 2005a.
removal is probably based on the perception (in both agencies and the public) of the permanence of the remedy. Dredging and excavation remove the mass of contaminants from the aquatic environment, and this has historically been viewed as key to reducing human health and environmental risks.
Technologies for removing sediment were already well established in the early years of sediment cleanup, in part as an extension of remediation technologies applied at upland sites. Most of the initial technologies for managing sediment came from the U.S. Army Corps of Engineers experience with navigational dredging and disposal. Other remedies were typically viewed as less certain by regulators and the public with respect to long-term effectiveness or permanence. Leaving contamination in place under a capping or MNR remedy was often considered more uncertain because of the residual risk posed by contaminants left in place.
The dynamic nature of aquatic environments has often led to the selection of removal as the preferred alternative in many areas of the country. Contaminated sediment is often associated with industrial, urban harbors where operational and navigational constraints are viewed as limiting the feasibility of capping or natural recovery. Those environments are often subject to disturbances, such as those caused by prop wash, seasonal flooding, ice scour, and storm surges, which were viewed as creating substantial risk if contaminants were left in place.
Removal of contaminated sediment has brought unique challenges that were initially not well recognized. Navigational dredging techniques adopted for environmental dredging are designed to achieve a specific bottom elevation or the removal of a specific volume, often in the shortest possible time, whereas environmental dredging typically must achieve a specific final concentration while minimizing contaminant releases during dredging, handling, and disposal. As dredging remedies have been implemented at various sites, the effects of resuspension and transport of contaminated material off site and residual contamination in a remediated area have become apparent (Bridges et al. in press). The risks associated with the implementation of environmental dredging have received a great deal of attention in the last few years (EPA 2005a; Wenning et al. 2006).
Greater experience with capping remedies has been gained over the last decade; cap performance can be better predicted and quantified, and this has led to greater acceptance among agencies. In addition, capping typically has been less expensive and can be implemented more quickly, so it is often preferred by responsible parties (Palermo et al. 1998). In response to the increasing experience with remedial technologies, recent guidance from EPA has called for a more equitable evaluation of all remedies with careful analysis of the short-term and long-term risks associated with any remedy and thorough consideration of site-specific conditions (EPA 2005a).
OVERVIEW OF ENVIRONMENTAL DREDGING
Dredging refers to the removal of sediment from an underwater environment. It involves dislodging and removing material on the bottom of a waterway. Dredges are normally classified according to the basic
operation by which sediment is removed, such as mechanical or hydraulic10 (EPA 1994). For purposes of this report, excavation in the dry using conventional equipment operating within dewatered containments such as sheet-pile enclosures or cofferdams is not covered. The term environmental dredging is more generally associated with removal of sediment from under water. Environmental dredging can be accompanied by backfilling of the dredged areas. Placement of clean material covers and mixes with dredging residuals and further reduces risk from contamination that remains after dredging. Unlike capping, permanent confinement of underlying material is usually not the goal.
Typical objectives of environmental dredging are shown in Box 2-6. Because the purpose of navigational dredging is to restore navigable depth to a waterway, the selection of equipment and operational approaches considers economics, effectiveness, and environmental protection (USACE/EPA 1992) in that order. Conversely, environmental dredging has remediation as its stated purpose. The distinction results in reversing the order of importance of the selection factors for equipment and operational approaches; that is, one needs to consider environmental protection and effectiveness first before considering economics (Palermo et al. 2006).
Objectives of Environmental Dredging
Source: Palermo et al. 2006.
Types of Environmental Dredges
Selection of dredging equipment is sediment specific, site specific, and operations specific. Many textbooks and manuals describe the science and engineering principles of dredges, their selection, and their operation (Bray 1979; USACE 1983; Herbich 2000). This section provides basic definitions of dredging methods and equipment types normally considered for environmental dredging. There is no attempt to list all the possible types of dredge equipment that may be applicable to environmental dredging. Box 2-7 lists the equipment most commonly used for environmental dredging according to type (category) and definition (Palermo et al. 2004). Figure 2-3 shows the basic dredge types. More detailed descriptions of environmental-dredging equipment are available elsewhere (Averett et al. 1990; EPA 1994; EPA 2005a).
Other dredge types—such as hopper dredges, dustpan dredges, and bucket-ladder dredges—are not included in Box 2-7, because they are used primarily for navigational dredging. In addition, within dredge types, specific designs may differ and may have varied capability. In general, the dredge types listed above represent equipment that is readily available and used for environmental dredging projects in the United States.
A number of newer dredges, including some specifically designed for environmental dredging, are available. They have been termed specialty dredges and are intended to provide benefits by reducing sediment resuspension and contaminant releases. Other advantages may include operational efficiency for removal of sediment and transportation, depending on the sediment and project conditions and the performance standards. Most specialty dredge designs originated outside the United States, but several U.S. companies have now formed partnerships that allow use of specialty equipment from various countries. Field experience with specialty dredges in the United States is limited (Palermo et al. 2003). The dredges have been proposed for use at contaminated sediment sites, but little information is available about their sediment-extraction efficiency or about the claimed improvements in innovations, such as improved solids capture and reduced resuspension.
The equipment used for environmental dredging is usually smaller than that commonly used for navigation dredging because removal
Equipment Commonly Used in Environmental Dredging
Specialty Dredges and Diver-Assisted Dredges
Source: Adapted from Palermo et al. 2004.
volumes and rates tend to be lower and water to be shallower. Mechanical-bucket sizes range from 2 to 8 m3 (about 3 to 10 cy), and hydraulic-pump sizes range from 15 to 30 cm (about 6 to 12 in.) (Palermo et al. 2006). Obviously, larger dredges are available for both mechanical and hydraulic equipment and can be used for environmental dredging if needed.
Dredging—One Part of the Overall Process Train
Physically removing sediments by dredging is only one component of the overall remediation process. The key processing steps shown in Figure 2-4 include (EPA 2005a):
Mobilization and setup of equipment.
Site preparation including debris removal and protection of structures.
Removal (environmental dredging).
Staging, transport, and storage (rehandling).
Treatment (pretreatment, solidification and stabilization of solids, treatment of decant water and/or dewatering effluents and sediment, and potentially separate handling and treatment of materials with and without special requirements under the Toxic Substances Control Act [TSCA]).
Disposal (liquids and solids).
Environmental dredging must be compatible with all later steps in the process train. For example, the production rate of a dredge (either mechanical or hydraulic) depends heavily on the mode of transportation and the ability to rehandle or directly manage the dredged material on the other end of the process. Compatibility must be considered with respect to the type of pretreatment, treatment, and disposal being planned, especially the availability, size, and capacity of disposal sites, the distance from dredging site to treatment or disposal sites, and constraints associated with production rates for transport, storage, rehandling, treatment, or disposal. Inefficiencies in remedial dredging projects can result from constraints associated with components of the remedy other than dredging, such as dewatering capacity, water-treatment effectiveness, and disposal location and capacity (Palermo et al. 2006).
Dredging accounts for only part of the overall cost of an environmental-dredging project. In a complex project, large costs may be associated with the transport, dewatering, and ultimate disposition of the dredged material. Recent data, described below, support the premise that dredging accounts for 10-20% of the total cost of an environmental dredging project. For example, EPA Region 1 (EPA 2005d) reported on the costs of the 2004 New Bedford Harbor Superfund dredging project,
as shown in Figure 2-5; dredging itself represents only 17% of the total yearly construction and operations cost. Similarly, in the Head of Hylebos remediation (Figure 2-6) dredging operations conducted from 2003 to 2006 (Dalton, Olmsted & Fuglevand, Inc. 2006), dredging represents 17% of the total cost. Dredging cost varies widely, depending on many factors, including site conditions, the nature of the sediments and contaminants, the type and size of dredge selected, production rates, and seasonal construction windows. However, when dredging is selected as a remedy, all the other components of the process train will probably be required and will account for most of the overall cost. Only in those cases where transportation and disposal of sediments are relatively inexpensive (for example, where there is an existing in-water or upland disposal site, both suitable for the long-term containment of contaminated sediment and in close proximity to the dredging) will the dredging be a major cost element.
TECHNICAL ISSUES ASSOCIATED WITH DREDGING
Environmental dredging typically strives to achieve contaminant-specific cleanup levels set at each site. A number of technical issues can limit ability or efficiency in achieving those levels. This section describes several of the issues, many of which are revisited in the context of site evaluations in Chapter 4.
Accuracy of Dredging vs. Accuracy of Sediment Characterization
The benefits of being able to position a dredge cut accurately may be achieved only if a corresponding degree of accuracy is reflected in the site and sediment-characterization data. The ability to map the precise location of chemical concentrations accurately both horizontally and vertically depends on the data density (grid density), accessibility of deeper sediments, and other aspects of site characterization (Palermo et al. 2006). In some cases, the ability to locate the dredge cut accurately exceeds the accuracy of the knowledge of the location of the contaminated sediments (Palermo et al. 2006).
In the context of this discussion, vertical operating accuracy is the ability to position the dredgehead at a desired depth or elevation for the cut, whereas vertical precision is the ability to maintain or repeat the vertical position during dredging. The key to the success of an environmental-dredging project is the removal of the target layer, which is de-
lineated by the cut line, without unnecessary removal of clean material (Palermo et al. 2006).
The ability to dredge to a specified cut line in the sediment has been greatly improved by the advent of electronic positioning technologies, such as differential global positioning systems (DGPSs) and kinematic differential global positioning systems (KDGPSs). Depending on site conditions, dredge operator ability, size and type of dredge, and positioning instrumentation and software, the dredgehead and cut elevation may be locatable with vertical accuracy of less than 30 cm. Vertical accuracies of 10 cm for fixed-arm dredgeheads should be consistently attainable, whereas vertical accuracies of 15 cm should be attainable with proper operator training in the use of wire-supported buckets (Palermo et al. 2006).
Notwithstanding the previous statements regarding accuracies and positioning of dredging equipment, there are numerous challenges in attaining them. For example,
Effective use of sophisticated dredge positioning systems requires sophisticated operators and contractors in order to achieve the stated accuracies.
In order to get effective positioning with any of the software packages, the operators must be specifically trained and capable of system operation, and the systems must be properly operated and calibrated.
Experience has shown that some systems are more difficult to operate than others, and some systems may experience difficulties maintaining calibration. Simply using an electronic positioning system on a dredge does not guarantee that the stated accuracy will be achieved.
Resuspension, Residuals, and Release of Contamination
All dredging equipment disturbs sediment and resuspends some fraction of it in the water column. Resuspended sediment and the associated contaminants can settle back to the bottom in the dredge cut; finer-grained materials can remain in the water column and be transported to other locations. Those materials are deposited as residuals and result from dredging. Dissolved contaminants may also be released to the wa-
ter column during dredging from resuspended or exposed contaminated sediment. Figure 2-7 is a conceptual illustration of environmental dredging and those processes.
Dredged sediment resuspension, release, and residual and the resulting risk (the “4 Rs”) were the focus of a recent workshop held at the U.S. Army Engineer Research and Development Center in Vicksburg, MS (Bridges et al. in press). Effective remediation by dredging requires minimizing the 4 Rs while maximizing the fifth R, removal—either the dredging production rate or the volume removed (Francingues and Thompson 2006). The type and amount of sediment resuspension, contaminant release, and residuals during a dredging operation depend on many site-specific project factors, as shown in Box 2-8.
Resuspension is the process by which dredging and associated operations result in the dislodgement of embedded sediment particles, which disperse into the water column. Resuspended particles may settle in the dredging area or be transported downstream. Recent EPA guidance for sediment remediation states that
When evaluating resuspension due to dredging, it generally is important to compare the degree of resuspension to the natural sediment resuspension that would continue to occur if the contaminated sediment was not dredged, and the length of time over which increased dredging-related suspension would occur.… Some contaminant release and transport during dredging is inevitable and should be factored into the alternatives evaluation and planned for in the remedy design.… Generally, the project manager should assess all causes of resuspension and realistically predict likely contaminant releases during a dredging operation (EPA 2005a, pp. 6-21, 6-22).
Resuspension concerns related to dredging include the physical effects of turbidity and burial that can result in seasonal restrictions on dredging operations (dredging windows). Sediment resuspension can
result in chemical releases to the water column (for example, from pore water displaced from the dredged sediment or by desorption from resuspended sediment particles) and residual contamination on the bottom after dredging. Resuspension can be caused not only by dredging equipment but by propwash of tenders (push boats or tugs used to move equipment) and during rehandling and transport operations, such as filling and overflowing of barges and leaky pipelines. Estimates of resuspension from environmental-dredging projects range up to 10% of the mass of sediment dredged (Patmont 2006). Rates of resuspension depend on equipment, material, operator, and other site-specific factors.
Residuals are contaminated sediment that remains after dredging. There are two general types of residuals: generated residuals, contaminated sediment that is dislodged or suspended during dredging and later redeposited within or adjacent to the dredging footprint; and undisturbed residuals, contaminated sediments found at the post-dredge sedi-
Site-Specific Factors Affecting Resuspension, Release, and Residuals Sediment Physical and Chemical Properties
Sediment Physical and Chemical Properties
Source: Adapted from Palermo et al. 2006.
ment surface that have been uncovered but not fully removed as a result of the dredging operation (Bridges et al. in press).
Residuals may result from incomplete characterization, inaccuracies of dredging, mixing of targeted material with underlying materials during dredging, fallback (dislodged sediment not picked up), and resettlement of resuspended sediments (Palermo et al. 2006). Also contributing to residual contamination are such processes as sloughing of sediment into the dredging cut and sloughing induced by bank or slope
failures. Site-specific factors, such as debris or limitation of dredging by bedrock or hardpan can influence the amount of residuals. Box 2-9 describes specific processes during dredging that contribute to residual formation.
The residual contaminant mass is typically limited to the upper few inches of sediment, which is populated and actively processed by sediment-dwelling organisms (although in the case of undisturbed residuals the depth can be substantially greater). That upper layer is subject to erosion and other physical and chemical processes that may promote release into the overlying water because of the entrainment of water into the dredged sediment, which causes physical (decreased consolidation) and chemical (redox) changes in the residuals. Residual contamination may also be attributable to sediment that was not dredged, because of the dredger’s failure to meet dredge cutlines (either depth or areal targets) or errors or incompleteness in site characterization that failed to identify appropriate depth and areal extent of contaminated sediment.
Patmont (2006) compiled data on residuals from 12 environmental-dredging projects. Final generated residuals ranged from approximately 2 to 9% (average = 5%)11 of the mass of contaminant dredged during the last production cut. There is little research on the amount of generated residuals transported outside the dredge prism, but their presence has been documented analytically (EcoChem Inc 2005) and visually with sediment-profile imagery (Baron et al. 2005).
Release is the process by which the dredging operation results in the transfer of contaminants from sediment pore water and sediment particles into the water column or air. Contaminants sorbed to resuspended particles may partition to the water column and be transported downstream in dissolved form along with contaminants in the released
Specific Processes Contributing to the Residual Layer During Dredging
For mechanical dredging, processes that contribute the residual layer are
For hydraulic dredging, processes that contribute the residual layer are
Dredging, either mechanical or hydraulic, can result in the formation of a residual layer through a variety of mechanisms including
Sources: Adapted from Dalton, Olmsted & Fuglevand, Inc. 2006; Fuglevand and Webb 2006, 2007; Hartman 2006.
pore water. Contaminants in the generated or undisturbed residuals may be released to the water column by densification, diffusion and bioturbation (Bridges et al. in press).
Releases of contaminants from the aforementioned sources and processes are considered to be up to about 5% of the contaminant mass in the sediment dredged, but larger or smaller releases may be observed, depending on site-specific factors and the type and operating characteristics of the dredge (Sanchez 2001; Sanchez et al 2002). The degree of contaminant release to the air and water is directly related to the degree of sediment resuspension (and pore water release) and chemical properties affecting the mass transfer of contaminants. Therefore, control of resuspension should have high priority at many dredging project sites that involve contaminated sediment. Contamination can also be released from sediment beds to the water column in soluble form without particle resuspension (Thibodeaux and Bierman 2003; Erickson et al. 2005). That suggests that the residual layer is also a contributor of contaminant release after dredging. Control of solids is important but is not always sufficient to prevent contaminant losses.
Impact on Risk
Risk can result from contaminant exposures driven by resuspension, production of residuals, and contaminant release. Those processes are important because they can alter the accessibility bioavailability of contaminants, create additional contaminant exposure pathways that
potentially affect the risk resulting from dredging, and may continue to influence risk after remedial operations cease. Surface-water concentrations and surface-sediment concentrations may increase during and after dredging and can result in adverse effects and accumulation of contaminants in organisms. The potential for volatile compounds to be released into the air may be an additional concern in connection with highly contaminated sites (EPA 2005a).
Release, resuspension, and production of residuals will affect risk over different spatial scales and time frames depending on the site characteristics and nature of the dredging operation. As described by Bridges et al. (in press), “Characterizing how dredging will influence direct risks includes considering how the processes contributing to risk change with time, which elements or receptors in the ecosystem are affected by these changes, the spatial scales over which effects would be expected to occur, and the uncertainties associated with the predicted changes and risk reduction.” As will be discussed in much greater detail in Chapter 4, resuspension and release occur in a shorter time frame during dredging operations. Residuals will remain following dredging, however, their distribution, longevity, and effects are poorly understood. To the extent that release, resuspension, and production of residuals are present and contribute risk at a site, they detract from the overall or net risk reduction resulting from the remedial activity. As such, they are an important consideration in evaluating the effectiveness of a remediation. As noted in the 4 Rs workshop (Bridges et al. in press) and recent EPA sediment guidance (EPA 2005a), there is increasing recognition of the importance of these processes and of factors that influence their control.
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