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Chronic Sublethal Effects of San Francisco Bay Sediments on Nereis 

(Neanthes) arenaceodenrutu; Nontreatment Factors 

I. AUTHOR(S) 

David W. Moore, Thomas M. Dillon 

. PERFORMING ORGANIZATION NAME(S) AND ADDRESS 

US Army Engineer Waterways Experiment Station 

Environmental Laboratory 

3909 Halls Ferry Road, Vicksburg, MS 39180-6199 

8. PERFORMING ORGANIZATION 

REPORT NUMBER 

Miscellaneous Paper D-92-4 

. SPONSORING/MONITORING AGENCY NAME(S) AND ADDRESS 10. SPONSDRING/MONlTORlNG 

AGENCY REPORT NUMBER 

US Army Corps of Engineers, Washington, DC 203 14-1000; 

USAE District, San Francisco, San Francisco, CA 941051905 

1. SUPPLEMENTARY NOTES 

Available from National Technical Information Service, 5285 Port Royal Road, Springfield, VA 22161. 

2a. DISTRIBUTION! AVAILABILITY STATEMENT 12b. DISTRIBUTION CODE 

Approved for public release; distribution is unlimited. 

3. ABSTRACT (Maximum 200 words) 

Initial development of a chronic sublethal sediment bioassay with the polychaete Nereis (Neanthes) 

arenaceodentatu is described. The test was initiated with 2- to 3-week-old pbstemergent juvenile worms and 

terminated after 21 days. The sublethal test end point was estimated individual somatic growth rate (milligrams dry weight/day). The potential bias due to selected nontreatment factors on polychaete survival and 

growth was evaluated. For example, grain size had no effect, while the number of worms added to each expo 

sure vessel was critical. Direct transfer from 30 ppt to salinities <15 ppt had a highly significant and adverse 

effect on survival and growth. Both survival and growth of juvenile worms may be adversely affected if test 

conditions involve exposures to 220 mg/L ammonia or 25.5 mg/L hydrogen sulfide. 

4. SUBJECT TERMS 15. NUMBER OF PAGES 

Chronic Neunthes 

Growth Sediment 

Sublethal 

41 

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OF REPORT 

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298.102 
Contents 

Preface . . , . . . , . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

l-Introduction . . . . . . . . . . . . . . . , . . . . . . . . . . . . . . . . 

Background ................................ 

Regulatory History of Dredged Material Management 

in San Francisco Bay ......................... 

Overview of Sediment Toxicity Test Development in the 

United States .............................. 

Scope ................................... 

2-Material and Methods .......................... 

Test Species. ............................... 

Laboratory Cultures ........................... 

Nontreatment Factors .......................... 

Statistic Analysis ............................. 

3-Results ................................... 

Effect of Intraspecific Densities on Juvenile Worms ........ 

Effect of Sediment Grain Size on Juvenile Worms ......... 

Effect of Salinity on Juvenile Worms ................. 

Effect of Ammonia on Juvenile Worms ................ 

Resistance of Juvenile Worms to Hypoxia and 

Hydrogen Sulfide ........................... 

4-Discussion ................................. 

5-Conclusions ................................ 

6-Recommendations . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

Figures l-9 

Tables l-5 

iv 

1 

1 

2 

4 

6 

9 

9 

10 

10 

12 

13 

13 

13 

13 

14 

14 

15 

19 

20 

21 

. . . 

III 
Preface 

The work reported herein was conducted by the US Army Engineer 

Waterways Experiment Station (WES) for the Headquarters, US Army 

Corps of Engineers (HQUSACE), and the US Army Engineer District 

(USAED), San Francisco. Financial support was provided by the USAED, 

San Francisco, through an Intra-Army Order for Reimbursable Services. 

Additional funding was provided by the HQUSACE through the Long-Term 

Effects of Dredging Operations (LEDO) Program, Work Unit 374-9, 

“Chronic Sublethal Effects.” The LED0 Program is managed through the 

Environmental Effects of Dredging Programs, Dr. R. M. Engler, Manager. 

Technical Monitors were Mr. David B. Mathis and Dr. William L. Klesch, 

HQUSACE. 

This report was prepared by Drs. David W. Moore and Thomas M. Dillon 

of the Ecosystem Research and Simulation Division (ERSD), Environmental Laboratory (EL), WES. The authors gratefully acknowledge the support provided by personnel of the USAED, San Francisco, Mr. Thomas 

Chase, Mr. Duke Roberts, Dr. Thomas Wakeman, Ms. Sandra Lemlich 

(currently with the Coastal Engineering Research Center, WES), and 

Mr. Brian Walls (currently with Battelle-Duxbury, MA). 

The work was performed under the general supervision of Dr. Bobby L. 

Folsom, Acting Chief of the Contaminant Mobility and Regulatory Criteria 

Group. Mr. Donald L. Robey was Chief, ERSD, and Dr. John Harrison 

was Chief, EL. 

At the time of publication of this report, Director of WES was Dr. Robert W. Whalin. Commander and Deputy Director was COL Leonard G. 

Hassell, EN. 

This report should be cited as follows: 

Moore, David W., and Dillon, Thomas M. 1992. Chronic 

sublethal effects of San Francisco Bay sediments on Nereis 

{Neanthes) arenaceodentata; Nontreatment factors. 

Miscellaneous Paper D-92-4. Vicksburg, MS: US Army 

Engineer Waterways Experiment Station. 

iv 
Background 

San Francisco Bay is a highly altered estuary. Two of the major reasons 

for this condition are the diversion of freshwater inflow from the SacramentoSan Joaquin River systems and the loss of wetlands. By 1980, the amount 

of fresh water flowing into San Francisco Bay had been reduced by 60 percent. This reduction is projected to increase an additional 10 percent by the 

year 2000. About 95 percent of all freshwater/estuarine marshlands had been 

lost to land reclamation before 1850. It is not surprising, therefore, that the 

estuary has experienced general decline in health and viability. One of the 

more noticeable symptoms of this decline has been the gradual loss of biological resources such as the striped bass and Pacific herring fisheries (Nichols 

et al. 1986). 

An increase in the input of environmental contaminants has accompanied 

the physical alterations of San Francisco Bay. Major pollutant sources include the freshwater inflow from the Sacramento-San Joaquin River systems, 

over 50 waste treatment plants, and about 200 industries that are permitted 

to discharge directly into the Bay (Luoma and Phillips 1988). Environmental contaminants discharged into aqueous systems tend to associate 

with particulate material in the water column and with bedded sediments. 

Periodically, bedded sediments must be removed to maintain navigable waterways. There is a concern that the relocation of these dredged materials 

may be causing unacceptable adverse impacts on aquatic biota within the 

San Francisco Bay. 

A large amount of sediment is dredged each year in San Francisco Bay. 

Approximately 5.5 million cubic meters (mcm) of sediment from Federal 

projects and permit actions is relocated annually. This value approximates 

the estimated average annual sediment inflow from natural sources, 6 to 

8 mcm (US Army Corps of Engineers (USACE) 1979). It has been estimated that 3 to 4 mcm of material leaves the Bay annually, while Central 

and North Bays experience a combined net accumulation of 4.2 mcm 

(USACE 1979). South Bay shows a net loss of nearly 0.8 mcm per year 

(Krone 1979). Despite these large numbers, the greatest yearly source of 

suspended sediment in San Francisco Bay is the resuspension of existing 

Chapter 1 Introduction 
bottom material. Approximately 120 to 130 mcm of sediment is resuspended each year by wind waves and currents (USACE 1979). The effect 

of these resuspended sediments on fish and aquatic invertebrates is unknown. 

To examine whether San Francisco Bay dredged material was causing 

adverse biological effects, the Planning and Engineering Division of the 

US Army Engineer District, San Francisco, contracted with the Environmental Laboratory of the Waterways Experiment Station to develop and 

conduct a series of chronic sublethal sediment bioassays using material 

from selected sites within the Bay. 

Regulatory History of Dredged Material 

Management in San Francisco Bay 

To help define what is known regarding the potential toxicity of San 

Francisco Bay sediments, it is useful to first examine how dredged material 

has been regulated in the past. Important milestones in that process are 

shown in Table 1. It was recognized very early that San Francisco Bay is 

a physically dynamic system and that most dredged material disposal sites 

were dispersive. Consequently, initial management concerns were mostly 

operational. That is, efforts were directed toward optimizing dredging 

and disposal operations to minimize transportation costs and redredging. 

Passage of the National Environmental Policy Act in 1970 outlined the 

Federal Government’s policy toward the environment and signaled an increasing awareness of environmental protection in this country. That 

same year, the San Francisco District initiated the Dredge Disposal Study 

(DDS) (USACE 1977). The DDS was a multifaceted interdisciplinary study 

designed, in part, to address some of the environmental concerns regarding 

potential impacts of dredge disposal operations. Although sediment toxicity was not examined directly, the physical impacts on biota (USACE 

1975b) and the bioaccumulation of contaminants from dredged material 

were evaluated in laboratory and field studies (USACE 1975c, 1975d). 

These studies demonstrated that: 

a. Estuarine animals can survive suspended sediment loads in excess of 

those normally encountered during dredging and disposal. 

b. In laboratory exposures to San Francisco Bay sediments, estuarine 

animals can bioaccumulate trace contaminants. 

c. In field studies, contaminant tissue concentrations in animals near 

the disposal operations were not different from those far removed. 

The one exception was slightly elevated p,p’-DDE concentrations in 

mussels, Mytilus edulis, during disposal. These differences were 

not detected 1 month postdisposal. 

2 

Chapter 1 Introduction 
In 1972, the California Regional Water Quality Control Board (RWQCB) 

adopted the Jensen criteria (Bowden 1977). These numerical criteria were 

developed by the US Environmental Protection Agency (USEPA) for freshwater sediment in the Great Lakes, and classified sediment as highly polluted, moderately polluted, or slightly polluted based on bulk sediment 

chemistry. As research on dredged material progressed, it became clear that 

these and other chemically based numerical criteria were technically inadequate because they did not assess either bioaccumulation potential or toxicity. Both assessments were evaluated in bioassay procedures contained 

in the USEPA/USACE Ocean Disposal Implementation Manual (USEPA/ 

USACE 1977). 

The San Francisco District adopted the use of bioassays for evaluating 

dredged material. Regulatory procedures were outlined in Public Notice 

(PN) 78-l. Elutriate procedures were emphasized since disposal sites in 

San Francisco Bay were generally dispersive. PN 78-l also reduced the 

number of disposal sites from 5 to 3, These were located in the Carquinez 

Strait, San Pablo Bay, and at Alcatraz Island. To facilitate net export out 

of the Bay, most dredged material was taken to the Alcatraz disposal site. 

In 1982, shoaling was noted at the Alcatraz site. As a result of this 

important development, the San Francisco District took several steps. 

They instituted a slurry policy to enhance dispersion during disposal. The 

District greatly reduced the amount of new dredged material taken to the 

Alcatraz site, and removed 30 tons (27,200 kg) of construction debris 

from the site. They monitored the physical configuration of the mound at 

Alcatraz and found it to be stable after two winter seasons. All of these 

actions led to the conclusion that the Alcatraz site could not be considered 

fully dispersive. Since the majority of dredged material in San Francisco 

Bay was taken to Alcatraz, a reduction in the capacity of that site represented a major impediment to maintenance dredging and to anticipated 

new work activities. The San Francisco District established the Disposal 

Management Program (DMP) in 1985 and charged it with finding solutions 

to the disposal problem. 

The Long-Term Management Strategy (LTMS) was initiated in 1989 to 

address increasing environmental concerns and to reflect San Francisco 

District’s commitment to a LTMS for dredged material. In 1991, the 

Ocean Disposal Implementation Manual was revised to reflect 14 years of 

regulatory experience and the many scientific advances that had occurred 

since 1977 (USEPA/USACE 1991). 

Chapter 1 Introduction 3 
Overview of Sediment Toxicity Test 

Development in the United States 

As indicated in the foregoing discussion, the regulation of dredged material in San Francisco Bay has taken advantage of scientific advancements 

that have taken place elsewhere in the United States. To address concerns 

specific to the potential toxicity of San Francisco Bay sediments it is important to have some general knowledge of advances in the field of sediment 

ecotoxicology. The following is not intended to be a comprehensive review per se; rather, it is meant to provide the reader a general sense of the advances that have occurred over the past 20 years. 

The first peer-reviewed journal article that reported an attempt to assess 

sediment toxicity was published in 1971 by Gannon and Beeton (Table 2). 

The laboratory procedures involved exposing amphipods to freshwater 

dredged material that had been placed in modified milk cartons. In 1973, 

recognizing the need for a strong technical base in its regulatory program, 

the USACE initiated the Dredged Material Research Program (DMRP). 

Included in the scope of this large program was the development of elutriate and solid phase bioassays to assess potential water column and benthic 

impacts, respectively (Saucier, Calhoun, and Engler 1978). The bioassays 

developed during the DMRP were subsequently incorporated into both the 

Ocean Disposal Implementation Manual (USEPAKJSACE 1977) and the 

interim guidance manual for discharge of dredged or fill material into navigable waters (i.e., the 404 Manual) (USACE 1976). These sediment bioassays represented a balance between the state of the art and what could be 

routinely conducted in a regulatory program. 

Prior to the mid-1970s, the scientific community expressed relatively 

little interest in sediment toxicity. Most of their energies were focused on 

the fate and effects of environmental contaminants dissolved in aqueous 

solutions. After the Priority Pollutant List was published in 1976, emphasis shifted for two reasons. One, it was discovered that many chemicals on 

the Priority Pollutant List were not very water soluble. Second, as more 

field data were gathered, it became apparent that concentrations of many 

contaminants on the Priority Pollutant List were much higher in the sediment than in the overlying water. Those findings led to initial speculation 

that sediments might be extremely toxic. However, subsequent research 

showed that the same forces causing chemicals to partition into the sediments also restricted their bioavailability to aquatic organisms. 

A major milestone marking these scientific advances was the 6th Pellston Conference held in 1984 (Dickson, Maki, and Brungs 1984). This 

was the first time leaders in the scientific community formally met to discuss the fate and effects of sediment-associated contaminants. Bioassay 

procedures contained in the 1977 USEPA/USACE Ocean Disposal Implementation Manual formed the basis for initial discussion. The researchers reached consensus regarding sediment toxicity (Anderson et al. 1984). 

They recognized that species sensitivity was related, in part, to the degree 

4 Chapter 1 Introduction 
of contact between sediment and organism. They recommended amphipods 

and mysid shrimp for lethal tests, and polychaetes, bivalves, oligochaetes, 

and fish for behavioral or sublethal tests. There was also a strong endorsement of the Tiered Testing Approach for evaluating contaminated sediments 

(USEPA/USACE 199 1). This approach eliminates unnecessary testing 

and directs limited resources to solving more urgent problems. 

Another important milestone in the evolution of sediment toxicity methods occurred in 1987. Members of the American Society for Testing and 

Materials (ASTM) created a new Subcommittee, E47.01 Sediment Toxicology. This Subcommittee was charged with identifying technically sound 

procedures for evaluating sediment toxicity and drafting appropriate standardized guideline documents. Guidelines, which are in various states of 

preparation, include: 

a. Solid Phase Toxicity Tests with Freshwater Invertebrates. 

h. Solid Phase Toxicity Tests with Marine Amphipods. 

c. Solid Phase Toxicity Tests with Marine Polychaetes. 

d. Solid Phase Bioaccumulation Tests with Invertebrates. 

e. Solid Phase Bioaccumulation Tests with Fish. 

J Guidance for Designing Sediment Toxicity Tests. 

g. Guidance for Collection, Storage, Characterization, and Manipulation 

of Sediment Prior to Toxicity Testing. 

When the USEPAKJSACE Ocean Disposal Implementation Manual was 

first published in 1977, the procedures it contained represented a balance 

between the state of the art and what could be practically achieved in a 

routine regulatory testing environment. It was realized at that time that 

revisions would have to be made to reflect anticipated advances in the scientific community as well as experience gained in regulatory testing programs. 

The Manual was revised in 1991. Significant improvements to the current 

Manual, as related to sediment toxicity evaluation, include the following: 

a. Formalizing the tiered testing approach. 

h. Refinements to the species selection process. 

c. Provisions for evaluating chronic sublethal effects. 

The assessment of chronic sublethal effects is treated as a Tier IV 

assessmenit and would be carried out only if there is a reason to believe 

chronic impacts may be occurring and if technically sound test protocols 

are available. 

Chapter 1 Introduction 5 
Scope 

Test procedures for evaluating potential chronic sublethal effects of 

dredged material on aquatic biota have not been fully developed. Most 

suggested protocols have been either water column tests ill-adapted for 

sediment or tests that utilize biological end points with little or no ecological relevance. Before the chronic sublethal effects of San Francisco Bay 

area sediments can be evaluated in a technically sound manner, a number 

of issues must be resolved, including (a) identification of appropriate test 

end points, (b) selection of a test organism, (c) development of test protocol, and (d) development of interpretative guidance. 

In acute toxicity tests, only one end point is measured, percent survival. 

In contrast, a plethora of end points exist for sublethal tests. These end 

points may be categorized according to the level of biological organization 

they represent. In order of increasing complexity, these levels are: 

molecular, cellular, tissue, organismic (whole animal), population, and 

community. When a sublethal effect occurs at any level of biological organization, mechanistic explanations may generally be found at lower levels 

while ecological consequences are found at higher levels of complexity. 

In the aquatic environment, the ultimate focus of environmental protection is the preservation of viable populations of organisms. Forecasting 

the potential impact at this level of biological complexity is difficult if not 

impossible. Bioassessments at lower levels of complexity (molecular-tissue) 

are possible, but their ecological relevance is uncertain. For these reasons, a 

surrogate toxicological bioassay approach is desirable. This approach, 

which examines whole animal (organismic) responses, represents a propitious balance between response sensitivity in the sublethal end point and 

ecological relevance of the results (Figure 1). Two of the most desirable end 

points for use in the surrogate toxicological bioassay approach are growth 

and reproduction. If reproductive success is impaired for a sufficient period of time, the viability of a population may be at risk. In addition, 

somatic growth and reproductive or gametic growth represent competing 

energy demands on the bioenergetics of aquatic animals. Therefore, if 

sediments are shown to reduce somatic growth, reproductive success may 

also be adversely affected. 

Both end points, growth and reproduction, are widely accepted in the 

scientific community as ecologically relevant. The California RWQCB, 

for example, has identified growth as a highly desirable sublethal end 

point. The Board utilizes growth bioassays in its regulatory program for 

effluent applicants. Test results involving growth and reproduction have 

the additional benefit of being generally understood and appreciated by a 

wider nontechnical audience. This latter characteristic is a very important 

consideration since data for large and/or controversial dredging projects 

will be carefully scrutinized by the public, and perhaps, the courts. 

6 

Chapter 1 Introduction 
Selection of an appropriate animal model is another important step in 

developing a chronic sublethal sediment bioassay. The benthic infaunal 

polychaete worm Nereis (Neanthes) arenaceodentuta will be used to 

evaluate chronic sublethal effects of San Francisco Bay sediments. Several 

features make this species particularly well suited for use in sediment 

toxicity tests. First, it maintains intimate contact with the sediment 

throughout its entire life cycle. Second, unlike many test organisms, N. 

arenaceodentuta can be used to evaluate both solid phase and suspended 

phase material. This allows direct comparisons to be made between the 

two phases. Third, N. arenaceodentutu is a sediment ingester. In both 

solid phase and suspended phase exposures, it readily ingests sediments 

while foraging for food and tube-building material. Fourth, it is well 

suited for monitoring of reproductive end points because, unlike most 

nereid polychaetes, it has no planktonic trochophore larvae. Instead, development is via metatrochophore larvae, which are easier to observe and 

manipulate from an experimental standpoint. Finally, because the whole 

life cycle can be completed in the laboratory, cultures producing test organisms of known age and background are possible. This is an attractive 

logistical characteristic from the perspective of regulatory testing. 

Test protocols for a chronic sublethal sediment bioassay with N. 

arenaceodentutu have already been developed for the Corps’ Seattle District, in cooperation with the State of Washington and Region X of the 

USEPA. In addition, a “Guide for Conducting Acute and Chronic Sediment 

Toxicity Test with Polychaetous Annelids” is currently under consideration 

by ASTM. Both of these tests are 20-day juvenile growth assays initiated 

with 3-week-old N. arenaceodentata. 

To have regulatory utility, a sediment bioassay must be able to assess 

the effects of anthropogenic contaminants in sediment without undue influence from nontreatment factors such as sediment grain size, intraspecific 

density, and ammonia toxicity. While it is important to evaluate these nontreatment effects in acute toxicity sediment bioassays, it is critical to do 

so for chronic tests. This is because the biological response following 

chronic exposure is likely to be more subtle (i.e., sublethal) and thus more 

susceptible to bias from nontreatment influences. For these reasons this 

report will address important nontreatment effects on survival and growth 

in the growth assay with juvenile N. arenaceodentata. Included will be a 

discussion of the effects of sediment grain size, intraspecific density, salinity, ammonia toxicity, and tolerance to hypoxia and hydrogen sulfide. The 

report will conclude with a series of recommendations for conducting 

chronic sublethal bioassays with the marine polychaete Neanthes. 

Future reports will focus on interpretative guidance for growth and 

reproductive end points, bioaccumulation and chronic sublethal effects of 

bedded and suspended sediments from selected sites in the San Francisco 

Bay area, direct effects of Bay area sediments on fecundity and larval development, effects of food ration on test end points, effect of storage on 

sediment toxicity, and a discussion of quality assurance/quality control 

procedures for chronic sublethal sediment bioassays. 

Chapter 1 Introduction 7 
2 Material and Methods 

Test Species 

Nereis (Neanthes) arenaceodentata is a benthic infaunal polychaete 

widely distributed in shallow marine and estuarine benthic habitats of Europe, all three coasts of North America, and the Pacific (Reish 1957, 1963; 

Sanders et al. 1962; Pettibone 1963; Reish and Alosi 1968; Day 1973; Gardiner 1975; Whitlatch 1977; Taylor 1984). This subsurface deposit-feeder 

constructs one or more mucoid tubes in the upper 2 to 3 cm of sediment 

and ingests sediment particles up to 70 p with a preference for particles 

around 12 p (Whitlatch 1980). Nereis (Neanthes) arenaceodentata has 

been accepted by the regulatory community as an appropriate test species 

for evaluating sediment (Johns, Gutjahr-Gobell, and Schauer 1985; 

USEPA/USACE 1991). A considerable amount of toxicological information on a wide variety of environmental contaminants already exists for 

this species (Reish 1985, Jenkins and Mason 1988, Anderson et al. 1990). 

Taxonomists are still debating the appropriate status of this species. 

Pettibone (1963), who suggested the name Nereis (Neanthes) arenaceodentata, 

lists five other names for this species: Spio caudatus, Nereis (Neanthes) 

caudata, Nereis arenaceodentata, Neanthes cricognatha, and Neanthes 

caudata. Day (1973) dismissed arenaceodentata in favor of acuminata, 

which was subsequently used by Gardiner (1975), Taylor (1984), and 

Weinberg et al. (1990). Neanthes arenaceodentata is most commonly 

used in the toxicological literature. Recent evidence suggests that Atlantic 

and Pacific populations are genetically dissimilar, reproductively isolated, 

and probably different species (Weinberg et al. 1990). Until the taxonomic 

status of this species is resolved, we will use the name most familiar to 

toxicologists and report the original source of worms. 

A life cycle of N. arenaceodentata is well documented, as are culture 

methods (Reish 1980). As worms approach sexual maturity, males and 

females establish pairs and occupy a common tube. Eggs are deposited 

by the female within the tube. The male presumably fertilizes the eggs at 

this time. The spent female either exits the tube and dies within 1 to 2 days 

or is eaten by the male. The male remains in the tube to incubate and 

8 Chapter 2 Material and Methods 
guard the developing eggs. He creates a current of water via rhythmic undulations to remove metabolic wastes and prevent hypoxic conditions. 

Larval development is direct via a nonplanktonic metatrochophore 

larva and occurs entirely within the parental tube. Emergent juveniles 

(EJs) exit the parental tube about 3 weeks after the egg deposition. They 

begin to feed and establish tubes of their own. Juvenile worms grow, and 

eggs become visible in the coelom of females about 6 weeks postemergence. 

Egg deposition follows 3 to 7 weeks later. The entire life cycle can be 

completed in the laboratory in 12 to 16 weeks at 20 to 22 ‘C. The nonplanktonic benthic larva and paternal care are unique among the Nereidae. 

This feature also facilitates laboratory culture and the experimental investigation of sublethal effects on growth and reproduction. 

Laboratory Cultures 

Stock populations of Nereis (Neanthes) arenaceodentata were obtained 

in March 1988 from Dr. D. J. Reish, California State University at Long 

Beach. Laboratory cultures were maintained using methods adapted from 

those described by Reish (1980) and Pesch and Schauer (1988). 

Briefly, the EJs were raised to sexual maturity in 38-L aquaria containing 30 L of 30-ppt seawater (Instant Ocean) maintained at a temperature 

of 20 ‘C. The photoperiod was 12 hr light. Animals were fed a combination of ground Tetramarin flakes (2 mg/worm) and alfalfa (1 mg/worm) 

twice weekly. This feeding regime was sufficient to maintain adequate 

water quality in a static-renewal system and has been found to produce 

survival and reproduction consistent with that reported for other laboratory populations of Neanthes (i.e., survival >80 percent; fecundity, ca. 

lOO-1,000 eggs/brood; EJ production, ca. 50 to 500 EJs/brood) (Reish 

1980, Pesch et al. 1987, Anderson et al. 1990). 

Seawater was renewed (80 percent of volume) every 3 weeks. This renewal schedule, based on water-quality monitoring data, was sufficient to 

maintain good water quality. After 10 weeks, worms were paired using 

the fighting response (Reish and Alosi 1968) and the presence or absence 

of eggs in the coelom. Unpaired worms were discarded. Pairs were 

placed in 600-ml beakers with 500 ml of seawater. Gentle aeration was 

provided via Pasteur pipettes, and the beakers were covered with watch 

glasses to reduce evaporation. Water was carefully renewed weekly in 

such a manner as to avoid disturbing worm pairs. 

Beakers were monitored daily for the presence of eggs and EJs. When 

discovered, EJs are mixed with other broods and returned to the 37-L 

aquaria to complete the culture cycle. Culture conditions and feeding 

rations were used in all experiments described below unless otherwise 

noted. 

Chapter 2 Material and Methods 9 
Nontreatment Factors 

Intraspecific densities 

EJs were randomly assigned to 600-ml beakers containing 500 ml of 

seawater at densities of 4.0, 2.0, 1.0, and 0.5X. The value X (two worms 

per beaker) is equal to the optimal density reported by Pesch et al. ‘( 1987) 

scaled downward to the surface area of the bottom of a 600-ml beaker 

(ca. 20 cm*). Food rations were adjusted according to the number of worms 

per beaker. The experimental regime included five replicate beakers per 

treatment. Gentle aeration was provided, and seawater was renewed 

weekly. Survival and growth (established individual dry weights) were 

determined after 3 to 6 weeks. 

The presence of sediment in beakers essentially adds a third dimension 

(height) to the worms’ habitat. Presumably, the number of worms per 

beakers could be increased. To evaluate this possibility, the above beaker 

experiment was repeated with 3 cm of the Range Point Pond sediment 

(90 percent sand treatment, see following paragraph) layered on the bottom 

of each beaker. There were four replicates per treatment. 

Grain size 

Emergent juveniles were exposed for 6 weeks to a series of grain sizes 

created by diluting uncontaminated natural fine-grained sediment with 

quartz sand that has been cornbusted in a muffle furnace. Surficial sediment was collected by hand from Range Point Pond, near Pensacola, FL. 

This site is far removed from any known source of pollution and has been 

used historically as a control sediment by the USEPA Environmental Research Laboratory at Gulf Breeze, FL (personal communication, Dr. Jim 

Clarke). The sediment was gently sieved to remove the fine fraction 

(<0.5 mm). This fraction was proportionally diluted with quartz sand (combusted in a muffle furnace at 500 ‘C for 12 hr) to yield five grain size 

treatments, 5, 30, 60, 90, and 100 percent sand. Nominal grain size treatments were confirmed by particle size analysis (Plumb 1981) Sediment 

was layered in 4-L all-glass aquaria to a depth of about 3 cm. Three liters 

of seawater was added, and gentle aeration was provided. The following 

day, 12 EJs were randomly assigned to each aquarium with three replicate 

aquaria per treatment. Seawater was completely renewed every 3 weeks. 

Worms within each replicate were removed after 3 and 6 weeks, counted, 

briefly rinsed in reverse osmosis (RO) water, and placed on aluminum foil 

tares. The worms were dried at 60 ‘C to a constant weight (48 hr), and 

total dry weight biomass was determined to the nearest 0.01 mg on a Cahn 

electrobalance. Individual worm dry weights were estimated by dividing 

the total dry weight biomass by the number of worms in each sample. 

10 Chapter 2 Material and Methods 
Salinity 

Three-week-old juvenile worms (reared in 30-ppt salinity seawater) 

were randomly assigned to 600-ml beakers containing 500 ml of seawater 

at 30, 25, 20, 15, or 10 ppt. The experimental regime included two worms 

per beaker and five beakers per treatment. Salinity was determined daily 

with a hand-held refractometer. Gentle aeration was provided, and seawater was renewed weekly. After 3 weeks the worms were counted, and individual dry weights were determined. 

Ammonia 

The effects of ammonia (as ammonium chloride) were determined with 

the same experimental design used to evaluate salinity effects. Preliminary 

experiments indicated that nominal ammonia concentrations (0, 2.5, 5.0, 

10, 20, and 40 mg/L) were very stable under these conditions. During the 

toxicity test, water samples were taken just prior to each weekly renewal 

to confirm nominal concentrations. Samples were adjusted to a pH of 

2 with concentrated HCl and stored at 4 ‘C for no longer than 2 weeks. 

Total unionized ammonia was determined in this and all experiments with 

the Orion ammonia-specific electrode after adjusting to a pH of 12 with 

concentrated NaOH. 

Hypoxia and hydrogen sulfide 

Hydrogen sulfide occurs in aqueous solution only under hypoxic conditions. Thus, the two factors are inextricably linked. To evaluate the effects 

of hydrogen sulfide, one must first evaluate hypoxia. A preliminary rangefinding experiment indicated that all juvenile worms survived 96 hr at 

concentrations 22.0 mg O+. A definitive test was conducted at nominal 

oxygen concentrations of 0.25, 0.50, 1 .OO, 1.50, and 6.50 mg/L. Hypoxia 

conditions were created by metering high-purity nitrogen gas (99.99 percent) into the normal laboratory air supply. Supply lines were fed through 

stoppered (two-hole) 1-L Erlenmeyer flasks, each containing 800 ml of 

30-ppt seawater. The following day, each flask was briefly unstoppered, 

dissolved oxygen (DO) was determined (02 probe and YSI meter), and 

five 3-week-old worms were added. There were five replicate flasks per 

treatment. Temperature, salinity, pH, DO, and survival were recorded on 

a daily basis, and ammonia was determined at test termination (96 hr). 

Worms were not fed for the duration of the experiment. 

Resistance to hydrogen sulfide was determined using the same experimental protocol as described for the hypoxia experiment. All sulfide exposures were conducted at 1.50 mg O+, the lowest oxygen concentration 

shown to result in loo-percent worm survival. Each day a stock solution 

of hydrogen sulfide was prepared by dissolving 60 g of sodium sulfide 

(Na,S*9H,O) in 1 L of deoxygenated (co.5 mg O&) RO filtered water. 

Nominal exposure concentrations (2.5,5.0, 10, and 20 mg/L) were prepared 

Chapter 2 Material and Methods 11 
from this primary dosing stock, There were two negative controls: 

normoxia (6.5 mg OJL) and hypoxia (1.5 mg 02/L). Test concentrations 

were adjusted daily, by adding predetermined amounts of the primary dosing stock to each flask. Hydrogen sulfide concentrations were measured 

daily (both prior to and following adjustment with primary dosing stock) 

using a HACH HS-7 test kit. This kit makes use of the reaction between 

lead acetate and hydrogen sulfide that produces a color change in a lead 

acetate-impregnated filter pad as PbS is formed. This color change is then 

compared to a chart to derive a semiquantitative measurement. 

Statistical Analysis 

Statistical analysis was conducted using the SYSTAT statistical package 

(Wilkinson 1988). Homogeneity of variance was determined for each biological end point using Bartlett’s Test for Homogeneity (Sokal and Rohlf 

1981). Dry weights were log-transformed prior to analysis to achieve homogeneous variances. Treatment effects were analyzed using one-way 

analysis of variance with subsequent mean separation via Tukey’s HSD 

(Honestly Significant Difference) test (Sokal and Rohlf 1981). All tests 

for significance were conducted at a significance level of a = 0.05. 

12 Chapter 2 Material and Methods 
3 Results 

Effect of Intraspecific Densities 

on Juvenile Worms 

In the absence of sediment, survival was high (81 to 100 percent) and 

essentially unaffected at worm densities up to 12 worms/beaker after 3 and 

6 weeks exposure (Figure 2). The one exception was significantly lower 

survival (60 percent) in the highest density treatment after 6 weeks. Growth 

was unaffected after 3 weeks but was significantly diminished after 

6 weeks at worm densities 14/beaker (Figure 3). 

In the presence of sediment, survival was again high (81 to 100 percent) 

and unaffected at worms densities up to 12/beaker (Figure 4). Growth 

was unaffected after 3 weeks exposure, but significantly lower after 

6 weeks at worm densities >8/beaker (Figure 5). 

Effect of Sediment Grain Size 

on Juvenile Worms 

Juvenile worms can tolerate a wide range of grain sizes from 5 to 100 percent sand. Survival was high (89 to 100 percent) after 3 and 6 weeks, and 

treatments were not significantly different from one another (Figure 6). 

Likewise, there were no significant effects of grain size on estimated individual worm growth after 3 or 6 weeks (Figure 7). 

Effect of Salinity on Juvenile Worms 

There was a very sharp threshold response of juvenile worms to salinity. 

At 220 ppt, salinity had no effect on either survival or growth (Figure 8). No 

worms survived, however, at 15 and 10 ppt. 

Chapter 3 Results 13 
Effect of Ammonia on Juvenile Worms 

As with salinity, there was a very sharp threshold response to ammonia 

(Figure 9). Survival and growth were unaffected in juvenile worms exposed 

to 0, 5, and 10 mg/L total ammonia for 3 weeks. Survival was always 

100 percent, and mean dry weights ranged from 2.660 to 2.811 mg. Both 

survival and dry weights were slightly but not significantly diminished 

(80 percent and 1.840 mg) in worms from the 20-mg/L treatment. Survival 

was 0 percent in the higher ammonia exposures (40 and 60 mg/L). Ammonia concentrations in the exposure beakers were very stable during the 

3-week static-renewal exposures. Prior to weekly renewals, measured ammonia concentrations were very similar to nominal values (Table 3). 

Resistance of Juvenile Worms to 

Hypoxia and Hydrogen Sulfide 

There was also a very sharp gradient in the resistance of juvenile worms 

to short-term (96-hr) hypoxic conditions. Percent survival was 100, 68, 

and 0 percent in 1.50, 1 .OO, and 0.50 mg 0*/L, respectively (Table 4). 

Measured oxygen concentrations were very similar to nominal values. 

Ammonia was essentially the same in all treatments at the end of the test 

(ca. 0.20 NH, mg/L) except in the very hypoxic treatment 0.25 mg O+ 

(0.07 NH, mg/L) where loo-percent worm mortality occurred early during the exposure. 

Survival of juvenile worms appears to be unaffected at sulfide concentrations 15.0 mg/L and adversely affected at concentrations 210.0 mg/L 

(Table 5). Mean survival at sulfide concentrations of 2.5, 5.0, 10.0, and 

20.0 mg/L were 100, 100, 44, and 0 percent, respectively (Table 5). Measured hydrogen sulfide concentrations generally mirrored nominal values. 

Measured oxygen concentrations approximated the threshold level where survival of juvenile worms is most affected by hypoxia (0.50 to 1.50 mg 0*/L). 

This may have obscured the response to sulfide. Also, differences in pH, 

which varied positively with hydrogen sulfide treatment, may have affected 

survival in the sulfide treatments. Mean values were 8.52, 8.77, 8.84, 8.87, 

and 9.42 in the 0-, 2.5-, 5.0-, lo-, and 20-mg/L treatments, respectively. 

14 Chapter 3 Results 
4 Discussion 

High intraspecific densities can adversely affect survival, growth, and 

reproduction in N. arenaceodentata (Pesch et al. 1987). These adverse effects are believed to be the result of more frequent aggressive encounters 

among worms. This speculation is firmly based in the numerous reports 

of intraspecific aggression in nereid polychaetes, especially N. arenaceodentata (Clark 1959, Evans 1973, Starczak 1984, Kristensen 1988). We 

have shown that the growth of juvenile N. arenaceodentata is unaffected 

when held fdr 3 weeks at the same no-effect density reported by Pesch 

et al. (1987) for nonsediment systems scaled to the bottom surface area of 

a beaker. The presence of sediment permits slightly higher worm densities. 

This enhanced survival and growth is probably the result of two factors. 

Sediment essentially adds a third dimension to the system, depth. As a 

consequence, the frequency of aggressive encounters is probably diminished. Second, N. drenaceodentata is an infaunal burrowing organism 

with a strong thigmotactic requirement. The presence of sediments is 

likely to 

reduce the general level of biological stress by meeting this requirement 

even in the absence of intraspecific aggression. 

Benthic species often show a preference for specific grain sizes. For 

example, Rhepoxynius abronius, a marine amphipod widely used in sediment bioassays, occurs in well-sorted sand and does not tolerate finegrained material very well (Dewitt, Ditsworth, and Swartz 1988). Since 

fine-grained material often carries a higher contaminant load, mortalities 

in sediment bioassays due to physical impacts in addition to sedimentassociated chemicals confound test results and interpretation. 

The distribution of sediment-ingesting polychaetes is associated with 

differences in grain size (Gaston 1987, Kristensen 1988). This distribution 

reflects not only individual preferences and tolerances but the physical energy of systems as well as available food resources. Results here indicate 

that survival and growth of N. arenaceodentata is unaffected by up to 

6 weeks exposure to a wide range of grain sizes (5 to 100 percent sand). 

McFarland (1981) reported similar laboratory results following 12 days exposure to sediments with various grain sizes. These laboratory findings 

are consistent with field observations which show that N. arenaceodentata 

inhabits sediments ranging from 2.6 to 87.6 percent silt and clay 

Chapter 4 Discussion 
(Whitlatch 1977). Thus, for this species, difference in grain size does not 

appear to be an important bias in sediment bioassays. 

Ammonia is highly toxic to many fish and aquatic invertebrates (USEPA 

1985). Acutely lethal concentrations (48- to 96-hr LC,, values) range between 0.5 and 2.0 mg/L for many aquatic organisms while chronic effects 

are observed at concentrations generally an order of magnitude lower (Tables 1 and 2 of USEPA 1985). In sediments, ammonia is derived primarily 

from the hydrolysis of macromolecules and subsequent deamination of 

amino acids (Santschi et al. 1990). Near the sediment surface, in situ interstitial ammonia concentrations are about 1 mg/L and increase with 

depth to 20 to 50 mg/L (Ho and Lane 1973; Murray, Grundmanis, and 

Smethie 1978; Kristensen and Blackburn 1987; Ankley, Katko, and Arthur 

1990; Kemp et al. 1990; Lerat, Lasserre, and le Corre 1990). 

Jones and Lee (1988) suggested that ammonia may be an important 

cause of toxicity in marine sediment bioassays. More recently, Ankley, 

Katko, and Arthur (1990) demonstrated that this was true for freshwater 

sediments containing substantial amounts of anthropogenic chemicals. If 

ammonia is a major causative agent in sediment toxicity bioassays, then 

past (and future) interpretations regarding potential environmental impacts of sediments have been erroneous. 

Results reported here for N. arenaceodentata suggest that juvenile 

worms are adversely affected when exposed for 3 weeks to ammonia concentrations 110 to 20 mg/L. Fine-grained organic sediments may have interstitial ammonia concentrations that meet or exceed these levels. 

Another important consideration is the actual in situ exposures for N. 

arenaceodentatu. Aller and Yingst (1978) demonstrated that the burrow 

wall of the deposit-feeding polychaete Amphitrite ornata is the site of intense decomposition. Ammonia concentrations were higher at the burrow 

wall (10 to 20 mg/L) than in the surrounding interstitial waters (1 to 

8 mg/L). Irrigation of the worm tube may act to partially offset these elevated ammonia concentrations. 

Many benthic animals inhabiting low-oxygen environments have evolved 

physiologic, biochemical, and behavioral strategies to survive hypoxic 

conditions (Prosser 1973). While N. arenaceodentata can survive lowoxygen tensions, it does appear to have a very sharp threshold response to 

hypoxia. Davis and Reish (1975) exposed female worms to low DO concentrations for 56 days. Survival and fecundity were unaffected at concentrations as low as 3 mg/L. At 2 mg/L, fecundity was reduced by 50 percent 

and, at 1 mg/L, both survival and fecundity were severely impacted. 

16 

Results reported here indicate that survival of juvenile worms is unaffected under very hypoxic conditions (1.5 mg/L). However, at slightly 

lower DO concentrations (1 .O and 0.5 mg/L), 96-hr survival drops quickly 

to 68 and 0 percent, respectively. In sediments, IV. arenaceodentatu avoid 

hypoxic conditions by drawing water through their tubes via rhythmic 

undulations. These findings suggest that as long as overlying water is 

Chapter 4 Discussion 
aerated, hypoxia will probably not be an important nontreatment factor in 

sediment bioassays with N. arenaceodentata. 

Hydrogen sulfide is a metabolic poison that is lethal to many fish and 

invertebrate species at concentrations less that 1 mg/L (Smith, Oseid, and 

Olson 1976; USEPA 1976; Main and Nelson 1988). Hydrogen sulfide is 

almost always associated with hypoxic conditions (Theede 1973) since sulfide is rapidly oxidized to SO4 or elemental sulfur in the presence of oxygen. 

Thus, H,S and hypoxic conditions are inextricably linked. This renders 

laboratory assessments of sulfide toxicity problematic. For N. arenaceodenturu, this situation is exacerbated by the fact that this species has a very 

sharp threshold response to declining oxygen tensions. In these experiments, all worms’survived sulfide concentrations 13.4 mg/L. At slightly 

higher concentrations (5.5 mg/L), survival was reduced to 44 percent. 

However, DO concentration in the treatment (1.20 mg/L) approximated 

hypoxia threshold levels. Consequently, the cause of worm mortalities in 

this treatment cannot be attributed solely to either H,S or hypoxia. 

Assigning causality to H$- or hypoxia-induced lethality from these 

data becomes less important when one considers in situ sediment exposures and possible detoxification strategies for sulfides. Hydrogen sulfide 

is produced by the bacterial reduction of sulfates and the putrefaction of 

proteins. Total sulfides in interstitial water of marine sediments commonly 

range from 1 to 300 mg/L (Berner 1963; USACE 1975a; Murray, Grundmanis, and Smethie 1978; King et al. 1982; Howarth et al. 1983). In contrast, interstitial H,S concentrations to which benthic infaunal organisms 

may be exposed are quite low, from 1 to 30 pg/L in bioturbated sediments 

to 100 pg/L in undisturbed anoxic sediments (USACE 1975a; McLachlan 

1978; Emerson, Jahnke, and Heggie 1984). These relatively low concentrations of interstitial H,S result from its removal as insoluble sulfides, 

primarily FeS (Ponnamperuma 1972, Santschi et al. 1990). In addition, 

some infaunal invertebrates have evolved strategies for excluding or detoxifying H,S (Powell, Crenshaw, and Rieger 1979; Arp and Childress 1983; 

Powell and Somero 1983). 

It has been shown that some infaunal organisms may even choose sulfidic environments over other habitats (Powell, Crenshaw, and Rieger 

1979; Meyers, Powell, and Fossing 1988). Vismann (1990) recently demonstrated that the blood, intestinal wall, and intestinal fluid of two nereid 

polychaetes (Nereis (Hediste) diversicolor and Nereis (Neunthes) virens) 

are capable of detoxifying H,S via sulfide oxidation. 

(b) 

Based on (a) the low in situ interstitial water concentrations of H,S, 

the high survival of N. urenuceodentutu to parts/million-level H,S 

exposures, and (c) sulfide-detoxifying capabilities in nereid polychaetes, 

H,S does not appear to be a serious nontreatment factor in chronic sediment bioassays with N. arenuceodentuta. 

Chapter 4 Discussion 17 
A number of critical issues require additional research and resolution. 

These topics fall under two general categories: methodology and interpretation. An important methodological issue to be resolved is the influence 

of food ration on chronic sediment bioassays with N. arenaceodentata. For 

example, if adequate nutritious food is furnished during sediment exposures,‘juvenile worms may feed on this ration preferentially and ignore the 

sediment as a food source. If inadequate or no food is provided, then sediment ingestion, and thus contaminant exposure, will likely increase. Tests 

on the same sediment under these two scenarios would probably yield 

quite different results. Differences in nutritional quality among sediments 

may also be a significant factor. For example, worms exposed to a nutritionally rich sediment (i.e., high organic content) may assimilate more energy and grow larger than worms exposed to less contaminated (control or 

reference) sediment with less organic matter. This is not an unreasonable 

scenario since nutritionally rich fine-grained sediments also often have 

high contaminant loadings. Taghon and Greene (1990) showed that as the 

level of enzymatically available protein in sediments declined, the feeding 

increased in the polychaete Abarenicofa pacifica until a maximum feeding 

rate was achieved at 0.05 to 0.1 mg protein/g dry sediment. Below this 

level, feeding rates declined. However, despite this functional adaptation, 

growth decreased steadily with decreasing protein concentrations (Taghon 

and Greene 1990). Clearly, additional research on the influence of food is 

required. 

To have regulatory utility, any chronic sublethal sediment bioassay 

must be accompanied by technically sound interpretive guidance. For 

N. arenaceodentata, this guidance must be able to answer the question, What 

diminution in growth is biologically important to N. arenaceodentata? 

For example, if growth in Sediment A is statistically different from Sediment B by 15 percent, is that difference biologically important? What is 

the minimum required level of absolute growth (milligrams dry weight) 

or growth rate (milligrams dry weight day-l) for N. arenaceodentata? 

To provide answers to these questions, one must establish the relationship 

between growth and reproductive success. This relationship must be expressed quantitatively and have predictive value. 

18 Chapter 4 Discussion 
5 Conclusions 

Conclusions based on this study are summarized below. 

Intraspecific density can have a statistically significant effect on 

growth and survival in N. arenaceodentata. 

The presence of bedded sediment in a test chamber permits slightly 

higher intraspecific densities than possible in the absence of bedded 

sediment. 

Nereis (Neanthes) arenaceodentata is tolerant of a wide range of 

sediment grain sizes with no statistically significant effect on either 

growth or survival. 

Nereis (Neanthes) arenaceodentata exhibits a sharp threshold 

response to salinity with statistically significant effects on survival 

at salinities 115 ppt. 

Juvenile N. arenaceodentata can tolerate total ammonia 

concentrations 120 mg/L with no statistically significant effect on 

either growth or survival. 

Juvenile N. arenaceodentata exhibit a sharp threshold response to 

hypoxia with no statistically significant effect on survival after 

96 hr of exposure at DO concentrations 21.5 mg/L. Davis and Reish 

(1975) reported effects on fecundity at DO concentrations 

13.0 mg/L. 

Juvenile N. arenaceodentata tolerate hydrogen sulfide concentrations 

~3.4 mg/L (DO = 1.5 mg/L) with no statistically significant effect 

on survival after 96 hr of exposure. 

Chapter 5 Conclusions 19 
6 Recommendations 

Based on the results of this study, the following recommendations are 

made: 

l Tests with N. arenaceodentata should be conducted at intraspecific 

densities 225 cm*/worm (i.e., 2 warms/600-ml beaker) in the absence 

of bedded sediment, and 125 cm3/worm (i.e., 4 warms/600-ml 

beaker) with bedded sediment (at a depth of 2 cm). 

l Tests with N. arenaceodentata should be conducted at salinities 

120 ppt (overlying water). 

l Measured total ammonia levels in tests with N. arenaceodentata 

should be 110 mg/L, (overlying water). 

l Measured DO levels in tests with IV. arenaceodentata should be 

13.0 mg/L (overlying water). Note: This level corresponds to 

levels previously shown (Davis and Reish 19’75) to adversely affect 

reproduction. 

l Measured hydrogen sulfide concentrations in tests with N. 

arenaceodentata should be 13.0 mg/L, (overlying water). 

20 Chapter 6 Recommendations 
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26 

US Environmental Protection Agency/US Army Corps of Engineers. 

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References 27 
ECOLOGICAL RELEVANCE 

MOLECULAR / CELLULAR 

ENZYMES 

DNA I RNA 

MEMBRANES 

HlSlDPATHOLOGY 

PROXIMATE COMPOSITION 

ORGANISMIC 

OSMOREGULATlON 

BEHAVIOR 

METABOLISM 

SURVIVAL 

GROWTH 

REPRODUC’IION 

POPULATIONS I COMMUNITIES 

ENERGY FLOW 

NUTRIENT CYCLING 

INTRASPECIRC INTERACTIONS 

ABUNDANCE 

DIVERSITY 

SECONDS - HOURS MINUTES - DAYS DAYS - MONTHS 

- 

RESPONSE SENSITIVITY 

Figure 1. Sublethal end points within levels of biological organization 

80 

2 4 8 

NUMBER OF WORMS / BEAKER I’f WKS 

Figure 2. Effect of increasing intraspecific density on mean percent survival of Nereis 

(Neanthes) arenaceodentata in the absence of sediment. Error bars = standard 

error of the mean. Asterisks = significant difference (p < 0.05) from reference 

1X treatment (X = 2 worms/beaker) 
2 4 8 

NUMBER OF WORMS / BEAKER I’: WKS 

Figure 3. Effect of increasing intraspecific density on mean dry weights of Nereis 

(Neanfhes) arenaceocfentafa in the absence of sediment. Error bars = standard 

error of the mean. Asterisks = significant difference (p < 0.05) from reference 

1 X treatment (X - 2 worms/beaker) 

80 

20 

0 

2 4 8 

NUMBER OF WORMS / BEAKER I’: WKS 

Figure 4. Effect of increasing intraspecific density on mean percent survival of Nereis 

(Neanrhes) arenaceodenfafa in the presence of sediment. Error bars = standard 

error of the mean 
1 

0.2 

0 

E!BSl3 WKS NUMBER OF WORMS / BEAKER rn6 WKS 

Figure 5. Effect of increasing intraspecific density on mean dry weights of Nereis 

(Neanfhes) arenaceodentate in the presence of sediment. Error bars = standard 

error of the mean. Asterisks = significant difference (p c 0.05) from reference 

1 X treatment (X = 2 worms/beaker) 

loo 

80 

m3 WKS % SAND 

Figure 6. Effect of grain size on mean percent survival of Nereis (Neanthes) 

arenaceodentafa. Error bars = standard error of the mean 
4 

a3 E 

!i 

92 

3 

& 

01 

0 

5 30 60 90 100 

% SAND 

l!BIi3 WKS -6WKS 

Figure 7. Effect of grain size on mean dry weights of Nereis (Neanthes) 

arenaceodentata. Error bars = standard error of the mean 

% SURVIVAL DRY WT, mg 

100 

80 

60 

40 

20 

0 

30 25 20 15 10 

SALINITY, % 

Ei % SURVIVAL I DRY WT 

Figure 8. Effect of decreasing salinity on mean percent survival and mean dry weights of 

Nereis (Neanfhes) arenaceodentata. Error bars = standard error of the mean 
100 5 

80 4 

80 3 

40 2 

20 1 

0 0 

0 5 10 20 40 80 

NH3, mg/L 

I3 % SURVIVAL n DRYWT 

Figure 9. Effect of ammonium chloride (measured as total NH, mg/L) on mean percent 

survival and mean dry weights of Nereis (Neanthes) arenaceodenteta. Error 

bars = standard error of the mean 
rable 1 

Milestones in the Regulation of Dredged Material in San Francisco Bay 

CESPN initiates Dredge Disposal Study. Terminated in 1975. 

-Farallon Islands Marine Sanctuary and is 

nel ocean disposal site receives final designation by USEPA. It can receive only 

sing regulatory and environmental 
Table 2 

Milestones in Scientific Development of Sediment Toxicity Tests 

Table 3 

Total Ammonia and pH in Beakers Prior to Weekly Renewals During 3-Week 

Ammonia Toxicity Test 

: Values represent mean (standard deviation); n = 5. 

Treatment terminated because of loo-percent mortality. 
Table 4 

Water Quality and Percent Survival of Juvenile Worms During 96-hr Exposure to 

Hypoxic Conditions 

1 .a0 

’ Values represent mean (standard deviation). 

Table 5 

Water Quality and Percent Survival of Juvenile Worms During 96-hr Exposure to 

’ Values represent mean (standard deviation). 
Waterways Experiment Station Cataloging-in-Publication Data 

Moore, David W. 

Chronic sublethal effects of San Francisco Bay sediments on “Nereis 

(Neanthes) Arenaceodentata” : nontreatment factors / by David W. 

Moore, Thomas M. Dillon ; prepared for Department of the Army, US 

Army Corps of Engineers and US Army Engineer District, San Francisco. 

41 p. : ill. ; 28 cm. - (Miscellaneous paper ; D-92-4) 

Includes bibliographical references. 

1. Marine sediments - California - San Francisco - Testing. 

2. Dredging spoil - Environmental aspects - Evaluation. 3. Nereis. 

4. Biological assay. I. Dillon, Thomas M. II. United States. Army. 

Corps of Engineers. Ill. United States. Army. Corps of Engineers. 

San Francisco District. IV. U.S. Army Engineer Waterways Experiment 

Station. V. Long-Term Effects of Dredging Operations Program. 

VI. Title. VII. Series: Miscellaneous paper (U.S. Army Engineer Waterways Experiment Station) ; D-92-4. 

TA7 W34m no.D-92-4