Technical Document 2662

July 1994

Environmental Effects

from In-Water

Hull Cleaning of

Ablative Copper

Antifouling Coatings

A. O. Valkirs

B. M. Davidson

L. L. Kear

R. L. Fransham

Computer Sciences Corporation

Applied Technology Division

A. R. Zirino

J. G. Grovhoug

NCCOSC 

RDT&E Division

Approved for public release; distribution is unlimited.
		
		

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ACKNOWLEDGMENT

We sincerely express our appreciation to the following individuals for their contributions to this

marine environmental research effort: Scott Henderson prepared a detailed proposal and a sampling

plan design and coordinated field collections during the initial stages of this project. Jon Groves

assisted with many aspects of field work and performed San Diego Bay sediment grain size laboratory analyses. Special thanks are extended to members of the R/V ECOS crew including Bart Chadwick, Chuck Katz, and Andy Patterson, who furnished logistics support for oceanographic field data

collected and near realtime analyses. Dr. Ken Richter provided assistance with the Acoustic Doppler

Current Profiler (ADCP) data collection and interpretation. Additionally, Dr. Richter made available

to this project his unpublished sediment transect data for the Naval Station region of San Diego Bay.

Mike Dean (NAVSEA OOC5) and Harry Hodgkins (OCNR 12B) maintained outstanding support as

co-sponsors of this multi-year research investigation. Their solid support and timely funding administration is greatly appreciated. Dave Ives and the Seaward Marine Services, Inc., San Diego-based

dive crew aboard support vessel Jacqueline II provided very professional and much appreciated support. We especially appreciate their willingness to adjust their operational hull cleaning schedules to

accommodate our specialized sampling requirements. Analytical chemistry support at no cost to the

project was graciously provided by Drs. Stuart Belli (Vassar College), Martha Stallard, Rosa and

Antonio Sanchez (University of Extremadura, Spain), and Gioacchino Scarano (University of Pisa,

Italy). Kelton Clark provided valuable assistance with copper ISE/ASV field and analytical efforts.

Pete Stang furnished valuable experience and guidance for in situ hull AF paint leach rate measurements. Paul Kenis performed the Microtox analyses and Bill Wild performed the Qwik-Lite chlorophyll fluorescence assays on hull cleaning plume effluent. Our appreciation is also extended to Pete

Seligman for his managerial support and careful review of all aspects of this project. Mention of trade

names or commercial products does not constitute an endorsement or recommendation for use.
iii

EXECUTIVE SUMMARY

OBJECTIVE

The primary goal of this field study was to objectively evaluate the magnitude of environmental effects attributable to in-water hull cleaning operations in a typical Navy harbor. San

Diego Bay was selected for detailed analyses of vessels coated with copper-based antifouling

paints. This evaluation can be used to support dicussions with environmental regulators,

strengthen future compliance strategies, and provide a framework for future monitoring studies.

RESULTS

Six classes of naval surface vessels were surveyed during commercial hull cleaning operations in San Diego Bay during a two and one-half-year period. While total copper concentrations

in bay waters were rapidly elevated near hull cleaning operations, the biologically active species

of copper complexed rapidly, returning actual toxic copper potential to near ambient conditions

within minutes to hours after hull cleaning ceased. With the Ion Selective Electrode (ISE), a

good technique was developed for tracking and modelling a copper plume formed during hull

cleaning. The small changes in free copper ion activity detected by the electrode indicate that

most copper released during hull cleaning operations is in the particulate form and is rapidly

incorporated into bottom sediments. Estimates of toxicity from the hull cleaning plume to resident nontarget biota from site-specific bioassays were used to evaluate potential effects on the

marine ecosystem. These data are consistent with other studies and show toxicity thresholds

above the actual copper levels contributed to San Diego Bay by hull cleaning activities.

RECOMMENDATIONS

It is essential that the Navy employ a database to evaluate the environmental risks from

potential contaminant-introducing activities such as in-water hull cleaning and the general use of

antifouling paints.

Because of increasingly stringent environmental regulation and escalating concern over the

status of bays and estuaries, in-water hull cleaning has recently come under the scrutiny of regulatory agencies. We recommend that local, state, and federal regulatory personnel be shown these

data and encourage open discussions to revise compliance standards and regulations based on

realistic environmental concentrations.

Site-specific water quality criteria for copper should be developed using recent EPA methods

comparing reference water and site-specific water toxicity testing. This approach more effectively addresses the bioavailable copper on a site-specific basis and is consistent with our findings that less than the total amount of copper measured is in a form toxic to organisms.
v

CONTENTS

EXECUTIVE SUMMARY iii . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

ACRONYMS x . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

BACKGROUND 1 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

OBJECTIVES 1 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

TECHNICAL APPROACH 1 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

INTRODUCTION 3 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

TOXICITY, ENVIRONMENTAL CHEMISTRY, AND FATE 3 . . . . . . . . . . . 

REGULATORY STATUS 3 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

SAN DIEGO BAY—GENERAL CHARACTERISTICS 5 . . . . . . . . . . . . . . . . . . . . . . 

PHYSICAL DESCRIPTION 5 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

BIOTA 5 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

COPPER CONTAMINATION 6 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

BIOLOGICAL EFFECTS 11 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

OVERVIEW OF COPPER TOXICITY 11 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

HISTORICAL SAN DIEGO BAY DATA 14 . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

CHLOROPHYLL BIOASSAY 15 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

MATERIAL AND METHODS 15 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

RESULTS AND DISCUSSION 17 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

MICROTOX . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19

MONITORING DATA 21 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

BAY-WIDE SURVEYS 21 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

WATER SAMPLES 22 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

FILTERS 23 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

SEDIMENTS 23 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

WATER COLUMN MEASUREMENTS 23 . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

SEDIMENT MEASUREMENTS 27 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

SEDIMENT GRAIN SIZE ANALYSIS 27 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 
vi

LOADING FACTORS 31 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

BRUSHING ACTIVITY AND PLUME CONCENTRATIONS 31 . . . . . . . . . . 

USS DURHAM (LKA 114) 31 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

USS ENGLAND (CG 22) 31 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

USS MARVIN SHIELDS (FF 1066) 36 . . . . . . . . . . . . . . . . . . . . . . . . . 

USS FORT FISHER (LSD 40) 36 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

USS TUSCALOOSA (LST 1187) 36 . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

USS RANGER (CV 61) 36 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

ION SELECTIVE ELECTRODE/VOLTAMETRIC SPECIATION 36 . . . . . . . 

THE RELATIONSHIP BETWEEN CU++ AND TOXICITY 43 . . . . . . . . . . . . 

MEASUREMENTS IN SAN DIEGO BAY 44 . . . . . . . . . . . . . . . . . . . . . . . . . . 

USS DURHAM (LKA 114) 44 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

USS MARVIN SHIELDS (FF 1066) 46 . . . . . . . . . . . . . . . . . . . . . . . . . 

USS FORT FISHER (LSD 40) 47 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

USS TUSCALOOSA (LST 1187) 47 . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

USS RANGER (CV 61) 48 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

RESULTS AND DISCUSSION 48 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

Detection of Cu+ . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 49

RELEASE RATES/SHIP LOADING FLUX 49 . . . . . . . . . . . . . . . . . . . . . . . . . 

SEDIMENT DEPOSITION/BENTHIC FLUX CHAMBER 49 . . . . . . . . . . . . . 

ADDITIONAL NONNAVAL SITES 51 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

Municipal Yacht Harbors/Commercial Shipping 51 . . . . . . . . . . . . . . . . . 

PACO Ore Terminal 54 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

SDG&E South Bay Power Station 55 . . . . . . . . . . . . . . . . . . . . . . . . . . . 

Municipal/Naval Station Storm Water Runoff 55 . . . . . . . . . . . . . . . . . . 

DISCUSSION 57 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

CONCLUSIONS 61 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

REFERENCES 63 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 
vii

FIGURES

1. Naval Station sediment copper collection sites several years prior to

hull cleaning operations (from Yamamoto et al., 1975) 6 . . . . . . . . . . . . . . . . . 

2. Copper concentrations in San Diego Bay sediments, 1984–1989

(NOAA, 1991) 9 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

3. Sediment copper concentrations in San Diego Bay at Integrated Risk

Assessment stations, July 1993 (Richter, unpublished data). Copper

concentrations in mg kg-1 dry weight 10 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

4. Hull cleaning plume IC50 graph, USS TUSCALOOSA, March 1993 18 . . . . . . 

5. Hull cleaning plume IC50 graph, USS RANGER, July 1993 18 . . . . . . . . . . . . . 

6. Bay-wide sample station locations in San Diego Bay 21 . . . . . . . . . . . . . . . . . . . 

7. Mean bay-wide surface water dissolved copper concentrations in g

L-1, May 1991 24 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

8. Mean bay-wide midwater dissolved copper concentrations in g L-1, 

May 1991 24 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

9. Mean bay-wide deep water dissolved copper concentrations in g L-1, 

May 1991 25 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

10. Mean bay-wide surface water dissolved copper concentrations in g L-1, 

June 1992 25 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

11. Mean bay-wide surface water dissolved copper concentrations in g L-1, 

February 1993 26 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

12. Mean bay-wide deep water dissolved copper concentrations in g L-1, 

February 1993 26 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

13. Mean bay-wide sediment copper concentrations in mg kg-1 dry weight,

May 1991 28 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

14. Mean bay-wide sediment copper concentrations in mg kg-1 dry weight,

August 1991 28 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

15. Mean bay-wide sediment copper concentrations in mg kg-1 dry weight,

February 1993 29 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

16. USS DURHAM in-water hull cleaning, May 1991. Mean surface water

dissolved copper concentrations in g L-1 . . . . . . . . . . . . . . . . . . . . . . . . . . . . 32

17. USS DURHAM in-water hull cleaning, May 1991. Mean midwater

dissolved copper concentrations in g L-1 . . . . . . . . . . . . . . . . . . . . . . . . . . . . 33

18. USS ENGLAND in-water hull cleaning, August 1991. Mean surface

water dissolved copper concentrations in g L-1 . . . . . . . . . . . . . . . . . . . . . . . 34
viii

19. USS ENGLAND in-water hull cleaning, August 1991. Mean midwater

dissolved copper concentrations in g L-1 . . . . . . . . . . . . . . . . . . . . . . . . . . . . 35

20. USS MARVIN SHIELDS in-water hull cleaning, June 1992. Mean

midwater dissolved copper concentrations in g L-1 . . . . . . . . . . . . . . . . . . . . 37

21. USS FORT FISHER in-water hull cleaning, February 1993. Mean

midwater dissolved copper concentrations in g L-1 . . . . . . . . . . . . . . . . . . . . 38

22. USS FORT FISHER in-water hull cleaning, August 1991. Mean sediment

copper concentrations in mg kg-1 dry weight 39 . . . . . . . . . . . . . . . . . . . . . . . . . 

23. USS TUSCALOOSA in-water hull cleaning, February 1993. Mean

midwater dissolved copper concentrations in g L-1 . . . . . . . . . . . . . . . . . . . . 40

24. USS TUSCALOOSA in-water hull cleaning, August 1991. Mean

sediment copper concentrations in mg kg-1 dry weight 41 . . . . . . . . . . . . . . . . . . 

25. USS RANGER in-water hull cleaning, July 1993. Mean surface water

dissolved copper concentrations in g L-1 . . . . . . . . . . . . . . . . . . . . . . . . . . . . 42

26. Ion Selective Electrode response profile, USS DURHAM in-water

hull cleaning, 16 May 1991. (A) Initial hull cleaning pass; (B) slack

low tide; (C) hull cleaning operations end 44 . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

27. Sediment copper flux, Shelter Island, 19 June 1993. Copper

concentrations in g L-1. Flux rate: –81 g m-2 day-1 (r2 = 0.42) 52 . . . . . . . . . . 

28. Sediment copper flux, Shelter Island, 25 June 1993. Copper

concentrations in g L-1. Flux rate: –58 g m-2 day-1 (r2 = 0.64) 52 . . . . . . . . . . 

29. Sediment copper flux, Naval Station, 22 March 1993. Copper

concentrations in g L-1. Flux rate: –87 g m-2 day-1 (r2 = 0.82) 53 . . . . . . . . . . 

30. Sediment copper flux, NRaD Pier, 1990. Copper concentrations in

g L-1. Flux rate: –158 g m-2 day-1 (r2 = 0.89) 53 . . . . . . . . . . . . . . . . . . . . . . . 

31. Relative estimated copper loading from antifouling paint and other

sources into San Diego Bay per year 58 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

TABLES

1. Sediment copper concentrations in San Diego Bay prior to hull

cleaning operations. Values are mg kg-1 dry weight (Yamamoto et al.,

1975; Peeling, NOSC, unpublished data) 7 . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

2. Copper concentrations reported toxic and/or inhibitory to marine

organisms 11 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

3. Results of Microtox tests of cleaning plume effluent 20 . . . . . . . . . . . . . . . . . . 
ix

4. In-water hull cleaning surveys 22 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

5. Sediment grain size analysis summary. Data as percent of sample 29 . . . . . . . . . 

6. Results of samples collected before, during, and after hull cleaning

operation, 16 May 1991. All copper concentrations in parts per 

billion (ppb) 45 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

7a. Samples collected near and around pier area, USS DURHAM, 

17 May 1991. All copper concentrations in parts per billion (ppb) 46 . . . . . . . . 

7b. Results of bay-wide survey, 21 May 1991. All copper concentrations

in parts per billion (ppb) 46 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 

8. In situ antifouling paint copper release rate measurements 50 . . . . . . . . . . . . . . . 
x

ACRONYMS

g microgram

l microliter

m micron, micrometer

AA atomic absorption

ADCP acoustic Doppler current profiler

AF antifouling

APDC ammonium pyrrolidinedithiocarbamate

ASTM American Society for Testing and Materials

ASV anodic stripping voltametry

cm centimeter

CTD conductivity, temperature, depth

Cu copper

DIW deionized water

EC effect concentration

EDTA ethylene diamine tetraacetic acid

EPA United States Environmental Protection Agency

GFAA graphite furnace atomic absorption

HDPE high density polyethylene

HMDE hanging mercury drop electrode

HNO3 nitric acid

IC inhibition concentration

ISE ion selective electrode

kg kilogram

L liter

LC lethal concentration

LOEC lowest observable effect concentration

m meter

M moles

M megohm

MIBK methyl isobutyl ketone

ml milliliter

MLC median lethal concentration

mV millivolt
xi

NATO North Atlantic Treaty Organization

NAVSEA Naval Sea Systems Command

NCCOSC Naval Command, Control and Ocean Surveillance Center

ng nanogram

NOAA National Oceanic and Atmospheric Administration

NOEC no observable effect concentration

NPDES National Pollutant Discharge Elimination System

NRaD NCCOSC RDT&E Division

NRCC National Research Council of Canada

NURP National Urban Runoff Program

OCNR Office of the Chief of Naval Research

pCu logarithm of the reciprocal of the copper ion concentration

PE polyethylene

pH logarithm of the reciprocal of the hydrogen ion concentration

ppb parts per billion

ppm parts per million

SDG&E San Diego Gas and Electric Company

TLM median tolerance limit

WRCB State of California Water Resources Control Board
1

BACKGROUND

With the desired extension of U.S. Navy vessel drydocking periods to 5 to 7 years, the

requirement for in-water hull cleaning of the new ablative copper antifouling coatings has

increased greatly. Because of increasingly stringent environmental regulation and escalating

concern over the status of bays and estuaries, in-water hull cleaning has come under scrutiny of

regulatory agencies. There is increasing likelihood that use and maintenance of copper antifouling paints will be regulated by measures such as release rate restrictions, discharge limits, debris

cleanup, and monitoring requirements.

It is essential that the Navy develop a base of information that can be used to evaluate the

environmental risks from its copper-producing activities such as in-water hull cleaning and

general use of antifouling paints. Studies performed about 15 years ago to evaluate the fate and

effects of hull cleaning wastes from conventional copper-based paints provided some useful

input for the database, but similar data need to be obtained for the newer ablative copper paints

that are now used on most Navy vessel hulls. Past and present experience shows that ablative

paints may require cleaning at some point during their operational cycle. Environmental

assessment information will be needed to (1) address compliance issues with state and federal

regulators on an ecological risk basis; (2) support formulation of remediation strategies if

necessary; and (3) develop, improve, and optimize hull husbandry practices.

OBJECTIVES

 To define the overall environmental loading and mass balance of copper in typical Navy

harbor environs and to evaluate the magnitude and fate of those materials derived from

in-water hull cleaning operations.

 To examine the possible toxicity of hull cleaning by-products of copper ablative coatings

and to identify the biological effects of chemical species of that material.

TECHNICAL APPROACH

Field efforts for these studies were performed in San Diego Bay, where many U.S. Navy

vessels receive in-water hull cleaning. Additionally, the existing environmental database is

extensive, since NRaD and other organizations have performed many water quality, biological,

and ships husbandry studies in this harbor.

During the course of this investigation, six detailed measurement/sampling series were

accomplished during in-water hull cleaning operations. Hull cleaning effluent plumes were

monitored for several hours while copper levels were measured. These measurements assist in

defining the composition, amounts, and fates of typical hull cleaning effluents. The Naval

Command, Control and Ocean Surveillance Center (NCCOSC) RDT&E Division’s (NRaD’s)

marine environmental survey craft, R/V ECOS, with equipment capable of obtaining real-time

measurements of supporting oceanographic parameters, supported the San Diego Bay survey

operations. Harbor-wide surveys of copper levels and related oceanographic parameters were

performed during periods when hull cleanings were not taking place to characterize background

levels and ambient copper distribution.

The fate and effects of three primary by-products of hull cleaning were evaluated:

 Dissolved copper—mostly released during and shortly after hull cleaning activities from

brushed and fragmented paint surfaces: lower levels may be released from the same
2

sources after cleaning and some may leach copper from debris on and in bottom sediments. Maximal effects are likely pulsed, having relatively short-term effects.

 Particulate copper—present largely as paint chips associated with bottom sediments;

some fine material is suspended in the water column for varying lengths of time.

Potential effects are probably long-term in nature, depending on dissolution rate of the

copper and degree of uptake of particulate and dissolved copper fractions by sedimentdwelling organisms.

 Organic fouling debris—present mostly on and in sediments near hull cleaning areas

(derived from fouling growth brushed from hulls). Effects are expected to be primarily

short-term, being caused by autotrophic consumption of oxygen and production of

dissolved nutrients during decomposition of the organic material.

In the aquatic environment, dissolved copper actually exists in many forms and compounds,

referred to here as copper species. In most studies of copper toxicity, “total copper,” which

includes copper of all species, is measured and is assumed to be responsible for any noted toxic

effects. In reality, considerable differences exist in the biological availability and toxicity of

various copper species. Thus, although total copper values may be high in an area, the amount

of toxic copper present may be low, because most copper is tightly bound in complex organic

compounds or is adsorbed onto particulate material.

The measurement of biologically active and ligand exchangeable copper was an important

aspect of this program. These are the two groups of copper species having the greatest biological availability and mobility. Total dissolved copper concentrations were measured in water

samples using standard atomic absorption techniques. Additionally, approximately 20% of all

water samples was analyzed for copper speciation using electrochemical (anodic stripping

voltametry) techniques. In later phases of the program, toxicity of the more commonly encountered copper species was tested for toxicity using a standard laboratory bioassay test (Microtox). Total particulate copper (associated with filtered particles larger than 0.45 m in size)

was measured for approximately 50% of all water samples.
3

INTRODUCTION

TOXICITY, ENVIRONMENTAL CHEMISTRY, AND FATE

Trace element speciation in surface waters and its ecological implications was explored

during a NATO conference (Leppard, 1983). The toxicity of essential and non-essential metals

was also discussed. It is generally acknowledged that essential elements are less toxic than

non-essential elements, with trace metals becoming toxic when the nutritional supply exceeds

the optimal amount by a factor of 40 to 200 fold. However, copper is considered to be toxic at

relatively low levels. Copper toxicity seems to be higher for lower organisms than for more

advanced organisms. The high toxicity of copper has been tentatively correlated with the

electronegativity, or other properties, of its salts or of chelates. The more electronegative metals

and all the divalent transition metals are particularly active in the chemical inactivation of

enzymes (Albergoni and Piccinni, 1983).

The toxicity of copper in environmental waters and sediments is directly linked to its

availability, and hence its physico-chemical form. Metals may vary widely in their natural

environmental forms including “...the hydrated ion; simple organic and inorganic complexes;

stable organic and inorganic complexes; adsorbed on inorganic colloids; adsorbed on organic

colloids; bound in various ways to large particles which include solid organics, minerals and

mixtures of various kinds of these. In terms of transport, the chemical species can be: truly

dissolved; insoluble but suspended; or a component of colloids having intermediate characteristics between those of dissolved substances and sedimenting particles” (Leppard and Burnison,

1983). Colloidal forms of copper complexes have particular relevance to this study. In coastal

seawater, the combined forms of organic and inorganic colloidal copper complexes accounted

for 56% of total copper (Batley, 1983).

In sediments, the issues of availability to organisms and physico-chemical form are complex

as well. In a study of the partitioning of trace heavy metals in sediments, “... the total concentrations of a metal in sediments provides little indication of the potential interactions of the

sediments with the abiotic [non-living] and biotic [living] components...” (Tessier and Campbell,

1987). Heavy metals in sediments may be associated with clay surfaces, clay structural matrices,

humic acids, surfaces of iron and manganese hydrous oxides, detrital and freshly precipitated

carbonates, nodules, detrital organic matter, and crystalline and amorphous sulfides. Where

copper is present in sediments containing sulfide, it is primarily a precipitated sulfide and not

bioavailable (Loring, 1981, 1982). When the molar sum of the acid volatile sulfides exceeds the

molar sum of the non-iron heavy metals in sediments, the heavy metals in the sediments are not

available, and hence non-toxic to aquatic species (DiToro et al., 1990).

As might be expected, the nature of copper in interstitial waters is also complex. Approximately 80% of the dissolved copper in interstitial water from anoxic sediment from Narragansett

Bay was complexed with natural dissolved organic matter, or associated with colloidal organic

matter (Elderfield, 1981). It was reported that insoluble sulfide was a major factor controlling

copper availability resulting in low interstitial water copper concentrations, and a low flux of

copper from sediment to the interstitial water (Elderfield, 1981; Elderfield et al., 1981).

REGULATORY STATUS

The water quality criterion for copper in marine waters is a 1-hour average not to exceed

2.9 g L-1 (U.S. EPA, 1985). This criterion has been adopted by the state of California in its
4

enclosed bays and estuaries plan as the state water quality objective for the protection of marine

organisms (Water Resources Control Board [WRCB], 1991). In the state of California’s ocean

plan for the protection of marine organisms in open waters, a 6-month median concentration of

3 g L-1 is targeted. A daily maximum concentration of 12 g L-1 and an instantaneous

maximum concentration of 30 g L-1 are also allowed (WRCB, 1990). It should be noted that

these target concentrations reflect total copper, and therefore are conservative since the active

component or bioavailable copper in a natural water sample would certainly be less than total

copper.

The most recent EPA view of the water quality criterion for copper in marine waters has

changed, however. The 1-hour average concentration limit of 2.9 g L-1 now may address that

copper fraction which passes through a 0.45-micron filter operationally defined as dissolved

copper (U.S. EPA, 1992). The EPA acknowledges that this fraction still contains colloidal

copper complexes and its bioavailability remains to be defined. This change is, however, a

significant development regarding the estimation of bioavailable copper in the aquatic environment. The EPA has further stated that a ratio of copper toxicity in ambient site water and

toxicity measured in reference water (Narragansett Bay) may be used to adjust site water copper

criteria (water effect ratio). This would have the effect of raising the allowable copper concentration in waters having less bioavailable copper than that in the reference water. The ratios

measured in seawater have not been particularly large, however, being on the order of a factor of

two (Delos, 1993, personal communication). These recent EPA recommendations have not as

yet been endorsed by the WRCB.
5

SAN DIEGO BAY—GENERAL CHARACTERISTICS

PHYSICAL DESCRIPTION

In a recent study addressing the distribution and fate of organotins from antifouling paint

sources in U.S. harbors (Grovhoug et al., 1987), San Diego Bay was described as a semienclosed water body extending approximately 28 kilometers in length and varying from 0.5 to

4.6 kilometers in width, with a depth of generally less than 4.5 meters. The sediment is characterized by primarily grey, brown, or black mud, silt, gravel, and sand. In many of the dredged

areas, the mud layer has been removed exposing the underlying sandy strata (Kram et al., 1989).

A channel 7.5 to 20 meters deep runs lengthwise through most of the bay. Currents generally

flow in the direction of the channel with velocities ranging from 0.5 to 3 knots dependent on

tidal fluctuations (Kram et al., 1989). The salinity is near oceanic throughout most of the bay

due to minimal freshwater input. Seligman et al. (1986) have discussed the bay in terms of four

regions characterized by current velocity patterns and use characteristics. The northern region

has relatively high current velocities and rapid flushing rates. This region includes most of the

shipping channels and dredged areas of the bay. The southern region is largely shallow having

low-velocity currents and longer water residence times. Nursery grounds for fish and invertebrate species are located in this region. The naval region includes several ship berthing areas

and is moderately well-flushed. Marinas and commercial boat basins compose a fourth region of

the bay and are generally characterized by moderate to dense aggregations of vessels in enclosed

embayments with reduced flushing characteristics (Seligman et al., 1989).

In the context of the general current velocities seen in the bay, the Submarine Support

Facility should have the highest flushing rate of naval facilities located in San Diego Bay due to

its location near the entrance channel. The carrier pier has a moderately high flushing rate due to

its location in and parallel to the deep channel. Water circulation at the Naval Station is reduced

due to its inner position relative to the long axis of the bay and the interference of piers and

ships.

Based on tidal dispersion, Ridley (1959) calculated an exchange ratio of 0.06 for San Diego

Bay, or a 6% reduction of a contaminant per tidal cycle assuming little mixing. Using this

figure, an 80% reduction in contaminants present would be achieved in 13 tidal cycles

(7.5 days). If a higher degree of mixing is assumed, the 80% flushing level might require

18 tidal cycles (9 days).

BIOTA

San Diego Bay is an important spawning area for ocean fishes, with the bay habitats

providing an integral element in the interconnected food web of adjacent ocean waters. Benthic

studies in the vicinity of the Naval Station reveal communities indicative of a moderately

disrupted habitat, however. Frequently observed benthic organisms include the snail Nassarius

tegulus, the burrowing shrimp Callianassa spp., the gobiid fish Clevelandia ios, and numerous

polychaete worms (U.S. Army Corps of Engineers, 1975a). Fish and invertebrate species tend to

be most abundant on the western side of the bay, where generally the water is shallower and

there is improved water quality. Black croaker (Cheilotrema saturnum) and bonito (Sarda

chiliensis) were the most numerous fish collected during a Naval Ocean Systems Center baywide survey (Peeling, 1974) in both the bay in general, and a sampling site near Pier 8 at the
6

Naval Station. Thirty-eight fish species were identified. Spiny lobsters (Panulirus interruptus)

also move into San Diego Bay during their spawning season (March–July).

The principal vegetation in most of the bay includes several species of algae and the eel

grass, Zostera marina. Much of the original marshland surrounding the bay has been replaced

by urban development. The southern portion of the bay is relatively undisturbed by industrial

activity.

COPPER CONTAMINATION

To evaluate the potential effects of copper released from hull cleaning activities, this study

focused on copper values in water and bay sediments. Some historical data concerning copper

values in the naval region is provided prior to initiation of hull cleaning operations in the

mid-1970s. Data collected by the California Regional Water Quality Control Board in 1972 (in

Peeling, 1974) indicate that sediment copper values off the end of the mole pier near Pier 8 at the

Naval Station ranged from 57 to 170 mg kg-1. The Army Corps of Engineers reported a

sediment copper concentration of 57 mg kg-1 off Pier 3 at the Naval Station in 1971 (in Peeling,

1974). During 1973 and 1974, a heavy metal survey of sediments was conducted in San Diego

Bay at two locations in the Naval region where hull cleaning activities would later take place

(Yamamoto et al., 1975). Figure 1 shows the locations of the sampling stations in the Naval

region. Copper concentrations in sediment from the Naval Station region are given in table 1.

Figure 1. Naval Station sediment copper collection sites several years prior

to hull cleaning operations (from Yamamoto et al., 1975).
7

Table 1. Sediment copper concentrations in San Diego Bay prior to hull cleaning

operations. Values are mg kg-1 dry weight (Yamamoto et al., 1975; Peeling, NOSC,

unpublished data).

Inactive Fleet Sites (see figure 1) Copper

Station 1 231

Station 2 197

Station 3 207

Station 4 264

Station 5 271

Station 6 284

Station 7 253

Active Fleet Sites (see figure 1)

Station 19 133

Station 20 360

Station 21 301

Station 22 189

Station 23 86

Station 24 111

Station 25 82

Station 26 40

Station 27 42

Peeling, NOSC, Unpublished Data

1C 25

1A 6.3

1 1.2

3 11

5 12

6 64

7B 20

8 270

9 39

9A 60
8

The highest copper concentrations in sediment were found at the Naval Station and inactive

fleet areas, followed by the carrier piers and the submarine support facility. The Naval Station

sediments averaged 246  83 mg kg-1, while the carrier and submarine facilities collectively

averaged 72  3 mg kg-1 copper. Water concentrations ranged from 3 to 15 g L-1 at the

inactive fleet site, and from 3 to 16 g L-1 at the active fleet site (Yamamoto, et al., 1975).

During later studies, sediment copper concentrations averaged 415 mg kg-1, while average water

copper concentrations in San Diego Bay were reported to be about 3 g L-1 (Salazar et al., 1980;

Zirino et al., 1978; Kenis et al., 1978). This latter value is similar to the lower copper concentrations measured in surface waters from the naval region over a 1-year period which ranged from

2.2 to 23 g L-1 (Krett, 1980). These earlier studies cited serve as a historical reference point for

later comparisons of copper concentrations in water and sediment samples collected during hull

cleaning operations at the Naval Station.

Additional data for sediment copper measurements made in the 1980s are summarized in

Woodward–Clyde (1991) from data cited in NOAA (1991). Sediment copper data from the

Naval Station ranged from 92 to 241 mg kg-1 (dry weight) during 1984 to 1986 (figure 2).

The copper values measured in sediment at the Naval Station (Site 5) are very similar to the

mean sediment copper value of 246 mg kg-1 for Naval Station sediments from the data of

Yamamoto et al. (1975) indicating little change in sediment copper concentration during an

interval of a decade. More recent and abundant data (San Diego Bay Integrated Risk Assessment, K. Richter unpublished data, 1993, figure 3) indicate that sediment copper values may

have increased at Naval Station Piers 3–4, 11–12, and 13 to levels ranging from approximately

360 to 480 mg kg-1 dry weight (figure 3). Sampling took place on 1–2 July 1993. Hull cleaning

operations took place during May 1991 and February 1993 (Pier 12), and August 1991, June

1992, and February 1993 (Pier 13). Probable causes for the observed increases are varied,

however. Storm runoff from the Naval Base and the proximity of Pier 13 to the PACO Ore

terminal, where copper ore was spilled into San Diego Bay for a number of years, complicate

isolation of copper inputs for these areas. The high copper value measured at station B-1

between Piers 3 and 4 was not likely to be due to hull cleaning activity because of the distance

from Piers 12 and 13 and direction of the prevailing circulatory current pattern. Since this

station is located close to the Naval Station at the head of Piers 3 and 4, it is quite possible that

storm runoff has contributed significantly to copper input at this area. Gadbois (1992) reported

copper in storm runoff ranging from 0.32 to 1.9 mg L-1 in samples collected from areas 10 and

14 located near the head of Pier 4. The variable copper concentrations measured in storm runoff

complicate predictions of input. Unfortunately, no “end of pipe” flow volumes are available for

the Pier 3–4 area precluding any estimate of input. Considering the time span over studies done

in 1971 to the present, it is quite likely that urban development in San Diego has contributed

significantly to the input of copper to the Naval Station and the bay in general.
9

Figure 2. Copper concentrations in San Diego Bay sediments, 1984 – 1989 (NOAA, 1991).
10

Figure 3. Sediment copper concentrations in San Diego Bay at Integrated Risk Assessment

stations, July 1993 (Richter, unpublished data). Copper concentrations in mg kg-1 dry

weight.
11

BIOLOGICAL EFFECTS

OVERVIEW OF COPPER TOXICITY

A general summary of copper toxicity to marine organisms is provided in table 2. These data

are provided to give an indication of the copper levels found toxic to various species under

various testing conditions.

Table 2. Copper concentrations reported toxic and/or inhibitory to marine

organisms.

Organism Effect

Concentrations

(ppm, mg L-1 ) Reference

Algae

Mixed diatoms Toxic 0.050 Zeituon et al., 1969

Dinoflagellates 2-day LC50 0.020 Saifullah, 1978

Mixed phytoplankton Inhibition of growth 0.100 Thomas et al., 1977

Exuviaella sp. Inhibition of growth 0.025 Mandelli, 1969

Gymnodinium breve Toxic 0.030 Marvin et al., 1961

Laminaria hyperborea Inhibition of growth (28-day

exposure)

0.050 Hopkins and Kain, 1971

Invertebrates

Colonial hydroid, Campanularia

flexuosa

11-day growth rate inhibition 0.010–0.013 Stebbing, 1976

Colonial hydroid, Campanularia

flexuosa

Enzyme inhibition 0.0014 Moore and Stebbing, 1976

Colonial hydroid, Eirene viridula 14- to 21-day growth rate

inhibition

0.030–0.060 Karbe, 1972

Ctenophore, Mnemiopsis mccrodyl 24-hr LC50 0.017-0.029 Reeve et al., 1976

Pleurobrachia pileus 24-hr LC50 0.033 Reeve et al., 1976

Polychaetes Lethal 0.120-0.250 McLusky and Phillips, 1975

Abnormal larvae 0.010 Reish, 1964

Nereis vireus 4-day LC100 0.500 Raymont and Shields, 1963

Nereis vireus 2-day LC100 1.5 Raymont and Shields, 1963

Cirriformia spirabracha 26-day LC50 0.040 U.S. EPA, 1980

Phyllodoce maculata 9-day LC50 0.080 McLusky and Phillips, 1975

Neanthes arenaceodentata 28-day LC50 0.044 Pesch and Morgan, 1978

Neanthes arenaceodentata 28-day LC50 0.100 Pesch and Morgan, 1978

Larval annelids, mixed species 24-hr LC50 0.089 Reeve et al., 1976
12

Organism Effect

Concentrations

(ppm, mg L-1 ) Reference

Echinoderms

Sea urchin, Arbacia punctatum 58% decrease in sperm

motility

0.300 Young and Nelson, 1974

Paracentrotus lividus Retardation of pluteal larval

growth 4-day test

0.010-0.020 Bougis, 1965

Crustaceans

Acartia clausi 2.5-hr LC50 5.0 Corner and Sparrow, 1956

Acartia clausi 2-day LC50 0.034-0.082 Moraitou and Apostolopoulou,

1978

Acartia tonsa 6-day LC50 0.009-0.073 Sosnowski et al., 1979

Acartia tonsa 24-hr LC50 0.104-0.311 Reeve et al., 1976

Euchaeta marina 24-hr LC50 0.188 Reeve et al., 1976

Metridia pacifica 24-hr LC50 0.176 Reeve et al., 1976

Phialidaum sp. 24-hr LC50 0.036 Reeve et al., 1976

Undinula vulgaris 24-hr LC50 0.192 Reeve et al., 1976

Copepod nauplii, mixed species 24-hr LC50 0.090 Reeve et al., 1976

Labidocera scotti 24-hr LC50 0.132 Reeve et al., 1976

Barnacle larvae 2-hr LC50 200 Corner and Sparrow, 1956

Barnacle larvae Inhibition of attachment 0.024 U.S. Naval Institute, 1952

Carcinus maenas 12-day lethal 1-2 Raymont and Shields, 1963

Cancer magister zoea 96-hr LC50 0.049 Bretler et al., 1984

Cancer maenas 48-hr LC50 100 Portmann, 1968

Crangon crangon 48-hr LC50 30 Portmann, 1968

Leander squilla Toxic 0.500 Raymont and Shields, 1963

Pandalus montagui 48-hr LC50 0.200 Portmann, 1968

Lobster, Homarus americanus 13-day LC50 0.056 McLeese, 1974

Molluscs

Acmaea scabra 3-day LC50 0.100 Marks, 1938

Cardium edule 48-hr LC50 1.0 Portmann, 1968

Busycon canaliculatum 77-day LC50 0.470 Betzer and Yevich, 1975

Nassarius obsoletus Decrease in O2 consumption,

3 days

0.100 MacInnes and Thurberg, 1973

Crassostrea gigas 96-hr TLM 1.9 Fujiya, 1960

Crassostrea gigas embryos 48-hr EC50 0.0053 Bretler et al., 1984

Crassostrea virginica 48-hr LC50 0.103 Bretler et al., 1984

Crassostrea virginica larvae 12-day LC50 0.046 Calabrese et al., 1977

Oyster larvae Lethal 3.0 Reish, 1964

Oyster larvae 48-hr LC50 0.103 Calabrese et al., 1973

Oyster larvae Inhibition and Tolerance Level 0.8-2.5 Reish, 1964
13

Organism Effect

Concentrations

(ppm, mg L-1 ) Reference

Oyster larvae 48-hr MLC 0.001-0.003 Connor, 1972

Oyster larvae 96-hr TLM 0.01 Mandelli and McIlhenny, 1971

Mya arenaria Toxic 0.02 Pringle et al., 1968

Mya arenaria 7-day LC50 0.035 Eisler, 1977

Mytilus edulis 96-hr LC50 0.28 Abel, 1976

Mytilus edulis 17-day LC100 200 Marks, 1938

Mytilus edulis embryos 48-hr EC50 0.0058 Bretler et al., 1984

Mytilus edulis 7-day LC50 200 Scott and Major, 1972

Mytilus edulis embryos Mean Acute Value 0.0058 U.S. EPA, 1985

Mercenaria mercenaria 10-day LC50 0.016 Calabrese et al., 1977

Mercenaria mercenaria 77-day LC50 0.025 Shuster and Pringle, 1968

Bay scallop, Argopecten irradians 42-day EC50 (growth) 0.0058 U.S. EPA, 1980

Argopecten irradians 119-day LC100 0.005 U.S. EPA, 1980

Black abalone, Haliotis cracherodiiHistopathological gill abnormalities, 4-day treatment> 0.032 Martin et al., 1977

Red abalone, Haliotis rufescens Histopathological gill abnormalities, 4-day treatment> 0.032 Martin et al., 1977

Fish

Fundulus heteroclitus 50% inhibition of hatching 0.250 Gardner and LaRoche, 1973

Fundulus heteroclitus 96-hr LC50 3.2 Jackim et al., 1970

Fundulus heteroclitus Histopathological lesions,

21-day test

<0.500 Gardner and LaRoche, 1973

Fundulus heteroclitus Enzyme inhibition, 4-day

exposure

0.600 Jackim, 1973

Menidia menidia Histopathological lesions,

4-day test

 <0.500 Gardner and LaRoche, 1973

Clupea harengus pallasi, embryo Incipient LC50, 6-day exposure

0.033 Rice and Harrison, 1977

Clupea harengus pallasi, larva Incipient LC50, 2 day exposure 0.900 Rice and Harrison, 1977

Brevoortia tyrannus 14-day LC50 0.610 Engel et al., 1976

Leiostomus xanthurus 14-day LC50 0.160 Engel et al., 1976

Micropagen undulatus 14-day LC50 0.210 Engel et al., 1976

Lagodon rhomboides 14-day LC50 0.150 Engel et al., 1976

Pleuronectes platessa 4-day LC50 0.750 Saward et al., 1975

Pseudopleuronectes americanus Histopathological lesions,

14-day exposure

0.180 Baker, 1969

Atlantic salmon, adults in fresh

water

Lethal 48 Sprague, 1964

Juvenile Coho salmon 96-hr LC50 0.06-0.074 Lorz and McPherson, 1977
14

Organism Effect

Concentrations

(ppm, mg L-1 ) Reference

Oncorhynchus tschawytscha Acute toxicity 0.04 Hazel and Meith, 1970

Atherinops affinis, fertilization,

embryo

EC50 0.024-0.190 Anderson et al., 1991

Atherinops affinis, development NOEC 0.032-0.123 Anderson et al., 1991

Atherinops affinis, larval survival LC50 0.212-0.288 Anderson et al., 1991

Atherinops affinis, larval survival NOEC 0.090-0.160 Anderson et al., 1991

HISTORICAL SAN DIEGO BAY DATA

Several studies have contributed data on copper toxicity to indigenous species in San Diego

Bay. Krett (1980) reported that soluble copper water concentrations measured by anodic

stripping voltametry (ASV) at the Naval Station and Shelter Island yacht basin averaged 4.0 g

L-1 compared to a site near the open bay side of Shelter Island that averaged less than 1 g L-1.

Although the copper concentrations were significantly higher at the Naval Station and Shelter

Island sites, phytoplankton productivity and biomass were not reduced in those regions relative

to the open bay Shelter Island pier site. Diversity indices revealed no significant differences

among sites; however, phytoplankton genera reported to be sensitive to copper were absent from

the Shelter Island inlet and Naval Station sites. Subsequent laboratory bioassay testing indicated

that the phytoplankton genera from the Shelter Island inlet site were more copper tolerant. Six

genera were present at ambient copper concentrations decreasing to four genera when the copper

concentration was increased to 40 g L-1 by addition of copper sulfate. The pier phytoplankton

genera decreased from 12 genera at ambient copper levels to four genera when the copper

concentration was increased to 40 g L-1. Productivity and biomass also decreased drastically at

5 g L-1 copper. Field and laboratory evidence therefore indicated that the phytoplankton

species exposed to higher ambient copper concentrations were generally more tolerant to copper

exposure (Krett, 1980).

Johnston (1990) reported dissolved (ASV) copper concentrations of 2.6 g L-1 at the Naval

Ocean Systems Center pier and 8.9 and 11 g L-1 within the Shelter Island yacht basin. The

mean number of fouling species per area was higher where lower concentrations of copper and

organotin compounds were measured in water samples.

Salazar and Salazar (1985) reported that sediment collected from the commercial basin area

of San Diego Bay and used in bioassay tests to determine its toxicity was acceptable for ocean

disposal under existing guidelines administered by EPA and the Army Corps of Engineers. The

species tested were a copepod (A. tonsa), a mysid (A. sculpta), and a bottom fish (C. stigmaeus).

Copper was measured at 210 mg kg-1 (ppm wet wt.) in the sediment tested. In a bioaccumulation test, copper was significantly accumulated in clam (M. nasuta) and polychaete worm

(N. arenaceodentata) tissues over control concentrations. The clams tested accumulated 3.3 mg

kg-1 copper, while the polychaete worms accumulated 16 mg kg-1 copper. The significant

accumulation noted indicates that copper was bioavailable to the test species in sediment from

this area of the bay. Percent clay was 14.3 percent in this area, which is similar to that measured

in sediments from the Naval sites sampled in this study (8.4 to 13.2% clay). (See table 5 in the

“Monitoring Data” section of this report.)
15

Copper is present in very high concentrations at a site approximately 300 yards south of the

Naval Station’s southern boundary (Pier 13). Copper ore was stockpiled and loaded onto ships at

the National City Marine Terminal from 1979 to 1985. Sediment copper concentrations as high

as 58,269 mg kg-1 dry weight were measured near the wharf (Woodward–Clyde Consultants,

1991). The Woodward–Clyde report extensively defines the spatial distribution of copper in the

sediment and discusses physical-chemical factors that would act to bind copper in the sediments

and therefore render it less bioavailable to organisms. The report further indicates that the

crystalline sulfides in the copper-ore bind the copper into a highly insoluble nontoxic form.

Bioassay tests with nine different bioassay test organisms and 14 biotic response characteristics

were used to evaluate the toxicity of sediments from areas of high and low sediment copper

concentration. Tests were conducted with flatfish, shrimp, sea urchin eggs and embryos, clams,

worms, two different amphipods, fish larvae, and oyster larvae. Eight of the nine species tested

exhibited no toxicity under standard test conditions. One of the amphipod species not found in

San Diego Bay exhibited a toxic response independent of the copper concentration in the

sediment tested. Other studies of numbers and types of species present in the sediments showed

that differences and similarities between numbers, types, and diversity of species in the area

were not related to sediment copper concentrations. The report concluded that the 1000 mg

Cu kg-1 remediation objective defined by the Water Quality Control Board, San Diego

Region, for the copper containing sediments in this area was highly protective of aquatic life

(Woodward–Clyde Consultants, 1991).

CHLOROPHYLL BIOASSAY

This section of the study provides environmental toxicity data for copper on the potential

effects of an effluent collected during the underwater hull cleaning of USS TUSCALOOSA

(LST 1187) coated with International Paint Company formulation BRA-540 paint, and USS

RANGER (CV 61) coated with International Paint Company formulation BRA-540/542 paint.

The Qwik-Lite chlorophyll assay was used to produce the results discussed in this report.

The Qwik-Lite chlorophyll assay uses microalgae as test organisms in a static non-renewal

exposure system. The premise of this assay is that a toxic response of microalgae to chemical

exposure can be a change in growth rate or biomass. This change can be quantified indirectly by

measuring chlorophyll fluorometrically. This bioassay was conducted in accordance with the

American Society for Testing and Materials (ASTM) Standard Guide for Conducting Static

96-hr Toxicity Test with Microalgae (E1218) (ASTM, 1992). Relative fluorescence (expressed

as a percent of control) and an inhibition concentration (IC5072) were used as endpoints. The

IC5072 is defined as a statistical or graphically estimated concentration that is likely to cause a

50% inhibition of one or more biological processes in 72 hours.

MATERIAL AND METHODS

The test medium was seawater effluent discharged from a hull cleaning device as it was

being used in the hull cleaning process. The effluent was filtered through 0.45-micron filters

prior to bioassay testing. The samples were collected in an acid-washed 5-gallon Nalgene

carboy by a method that directed a small portion of the discharge effluent of the scamping device

through a 3/4-inch hose over a distance of approximately 25 feet. Thus, the toxicity potential

was a conservative estimate based on samples collected in the centroid of the effluent discharge

of a hull cleaning machine prior to dispersion due to currents.
16

The instrumentation used for measuring in-vivo chlorophyll via relative fluorescence from

algal cells was a Turner Model 112 Fluorometer equipped with a 110-853 (Blue T-5) lamp, a

photomultiplier type R-136 (red sensitive), sliding window orifices 1X, 3X, 10X, and 30X, a

5–60 filter for excitation, and a 2–64 filter for emission to detect fluorescence at > 640 nm. The

optimum sensitivity for chlorophyll  measurements is obtained at an excitation wavelength of

430 nm and emission wavelength of 663 nm. The instrument was blanked between readings

with filtered (0.22 m) seawater, and samples were read on a static digital readout. Borosilicate

glass vials and 125-ml flasks were used as test containers. Prior to testing, all borosilicate

containers were soaked in RBS, 10% nitric acid and finally triple-rinsed with deionized water.

All vials and flasks were seawater-aged prior to use.

Isochrysis sp. was obtained from the North East Pacific Culture collection at the University

of British Columbia, Canada. Monocultures were maintained in Enriched Seawater Medium in

accordance with appendix A2 of ASTM Standard E1218 in 2-liter borosilicate erlenmeyer flasks

under a light regime of 12:12 hours (light:dark) at approximately 100 foot-candles delivered

from cool white fluorescent bulbs. Ambient temperature was 22 to 25 centigrade.

The same amount of algal inoculate was exposed to different concentrations of the initial

effluent for a period of 72 hours. Three replications at six different concentrations were tested.

Tested effluent copper concentrations in the first assay were 6.25% (14 ppb), 12.5% (28 ppb),

50.0% (110 ppb), 75.0% (165 ppb), and 100% (220 ppb) of effluent as the test article. The

control groups were exposed to 0% of effluent. The test was repeated for the second assay using

the following effluent copper concentrations: 6.25% (15 ppb), 12.5% (30 ppb), 25.0% (60 ppb),

50.0% (120 ppb), 75.0% (180 ppb), and 100% (240 ppb) effluent. Copper concentrations were

determined by GFAA following the bioassay procedure.

At the beginning of each bioassay, the same amount of algal stock culture (150 l) was

introduced into three 125-ml erlenmeyer flasks containing 25 ml of dilution water for the

controls and each of six different concentrations of the effluent mixture as treatments. The

dilution water was San Diego Bay water collected off of NCCOSC RDT&E’s (NRaD’s)

Pier 136. The water had been filtered and enriched with the same media as the stock cultures,

except for micronutrients (trace metals). The control group received no effluent. The effluent

mixture did not exhibit background fluorescence or color when tested as a reagent control.

The filtered effluent was clear and exhibited no phase separation. There was approximately

7.5 mm of headspace in each flask. All flasks were read immediately after introduction of the

algae into the flasks (t=1) and at 24-hour intervals thereafter for a period of 72 hours. Nominal

concentrations were analyzed for inorganic copper after the bioassay was completed. Fifty ml of

each effluent concentration and test control water (0% effluent) were collected in acid-washed

nalgene bottles and stabilized with 0.2% Ultrex nitric acid. Samples were then analyzed by

graphite furnace atomic absorption (GFAA) utilizing a direct seawater vaporization methodology. Results were adjusted to coincide with known dilution factors.

The following deviations from prescribed procedures occurred during this investigation:

Assay was completed after 72 hours vice 96 hours. Previous assays indicated that

significant differences in exposure groups can be determined within the 72-hr period.

Also, growth of the inoculate in control groups normally exceeds the range of the

fluorometer after 72 hours.
17

Samples were not filtered immediately after collection. Initial sampling procedures

failed to account for continued leaching of particulates that apparently occurred in the

unfiltered water contained in carboys. This leaching resulted in increased concentrations

of copper. Comparison of values from the 18- and 6-hr filtered samples and on-site

filtered samples collected from the plume in the same location indicated that copper had

approximately doubled in the filtered fraction during the 18- and 6-hr storage periods.

Some alteration of dissolved and complexed copper species may have occurred from

that characteristic of the initial sample.

The dilution of the 25% effluent concentration in the first bioassay was incorrect.

Results of the 25% effluent exposure group were eliminated due to human error in

making the dilution.

There was no attempt to fully characterize the effluent. A complete analysis or a full

characterization of the effluent of all chemical components was not performed, thus

other factors may have been present that could interact with available copper and/or

contribute to the toxic loading.

The Enriched Seawater Media micronutrient (trace metal and ethylene diamine

tetraacetic acid [EDTA]) solution was not added to the dilution water. This was done to

avoid chelating the metal(s) in the environmental sample and possibly affecting their

bioavailability.

Mean relative fluorescence, standard deviation, and coefficient of variation were calculated

for each control and effluent concentration replicate. Relative fluorescence calculated as a

percentage of control values was plotted versus exposure time. The IC5072 was then estimated

by plotting percent reduction of chlorophyll fluorescence versus concentration of the effluent,

fitting a sigmoidal dose-response curve to the data, and then determining the concentration

corresponding to the 50% reduction level. No Observable Effect Concentration (NOEC) and

Lowest Observable Effect Concentration (LOEC) values could not be estimated because

significant differences were found between the control groups and all exposure groups concentrations making it difficult to identify a NOEC.

RESULTS AND DISCUSSION

USS TUSCALOOSA Hull Cleaning Plume Assay: Based on published literature the IC50

for algae has been estimated to be approximately 50 ppb (Cu). Results are displayed in figure 4.

Response across concentrations was measured. Algal growth rate decreased as effluent concentration increased and growth was arrested at 75% (165 ppb (Cu)) effluent exposure. The

observed IC50 value was approximately 42 ppb (Cu).

USS RANGER Hull Cleaning Plume Assay: Results are displayed in figure 5. Response

across concentrations was measured, and algal growth decreased as effluent concentration

increased, and growth was again arrested at 75% (180 ppb (Cu)) effluent exposure. The

observed IC50 value was approximately 45 ppb (Cu).

Both bioassays demonstrated a concentration-response where relative fluorescence decreased

with increased effluent exposure. This is a direct measure of chlorophyll  presence in each

sample, which is an indication of algal growth (or number of viable algal cells present in the

sample) relative to effluent exposure. IC50 values for both assays ranged between 42 to 45 ppb.
18


	









Figure 4. Hull cleaning plume IC50 graph, USS TUSCALOOSA, March 1993. 
	









Figure 5. Hull cleaning plume IC50 graph, USS RANGER, July 1993.
19

The results of this test should be viewed as one test of effluent copper toxicity. Additional

testing is necessary to confirm the results, and, if possible, should include more complete

chemical analysis of the effluent. Also, special caution should be taken to ensure that samples

are processed upon collection as quickly as possible. The available information on the effluent

mixture demonstrates toxicity to this species of microalgae similar to literature values in a

controlled laboratory setting. The steep response curve indicates a rapid toxic response to the

effluent.

Since holding the unfiltered effluent water sample greatly increased the 0.45-m filtered

water copper concentration, the importance of determination of copper toxicity based on

available copper is emphasized. The recent EPA approach involving bioassay testing of ambient

and reference waters to develop a water-effect ratio would likely raise the site-specific copper

water quality criteria in an area such as San Diego Bay where dissolved organic matter is more

likely to interact with free copper ion than in water from a cleaner reference site.

MICROTOX

The Microtox bacterial bioassay was used to measure the toxicity of water samples

collected from the hull cleaning plumes of three ships. Samples were collected during the

cleaning operations within the brushing plume area. Microtox bioassays were performed on

filtered and/or on unfiltered samples. Additionally, a reference water sample was collected

approximately midway between the Naval Station and the Amphibious Base in the center of the

bay (SDCH4). Toxicity values are expressed as the percent of the original seawater sample that

yielded either a 20% or 50% (EC20 or EC50) reduction in light output after 5- and 15-minute

exposures using the Microtox system. Light-emitting bacteria were employed as the test

source. An EC50 of 100% means the seawater sample was tested with no dilution. An EC20 of

25% refers to a 1:3 dilution of the sample with reference seawater. Typically, an EC50 is used to

express toxicity. This is the concentration that produces a 50% reduction in light output relative

to a seawater control sample (reference sample) after 5 or 15 minutes. Data from these tests are

summarized in table 3.

When Microtox tests reveal no measurable toxicity relative to controls using an undiluted

test sample, an EC20 may be obtainable. For example, a filtered water sample from the cleaning

plume of USS RANGER did not exhibit a 5-minute EC50 using an undiluted sample (100%),

while a 5-minute EC20 was recorded with a 58% sample (see table 3). The lower percent effect

level (EC20) provided a measurable toxic effect.

Plume samples from USS FORT FISHER and the filtered samples from USS TUSCALOOSA exhibited no measurable toxicity with undiluted samples. The unfiltered sample from

USS TUSCALOOSA exhibited EC20 values of 93 and 48% at 5 and 15 minutes, respectively,

demonstrating slight toxicity. USS RANGER plume samples exhibited the greatest toxic

responses measured among cleaning samples collected. The unfiltered sample was the most

toxic tested exhibiting 5- and 15-minute EC50 values of 73 and 43%, respectively. The undiluted

filtered sample exhibited no measurable toxicity after 5 minutes, while after 15 minutes, an EC50

of 95% was recorded. EC20 values of 58 and 56% were measured at 5 and 15 minutes, respectively.

The Microtox testing was performed to assist in evaluation of the potential plume toxicity.

However, due to the nonspecific nature of the test, no clear assessment of copper toxicity is

possible from the Microtox data alone. The results indicate that the plume samples were

generally not acutely toxic.
20

Table 3. Results of Microtox tests of cleaning plume effluent.

Filtered Water Unfiltered Water

EC20 EC50 EC20 EC50

5 min 15 min 5 min 15 min 5 min 15 min 5 min 15 min

USS RANGER 58% 56% *** 95% 73% 43%

SD=9 SD=5 SD=3 SD=6 SD=2

n=3 n=3 n=3 n=3 n=3

USS TUSCALOOSA *** *** *** *** 93% 48% *** ***

SD=6 SD=3

n=3 n=3

SDCH4 *** *** *** ***

USS FORT FISHER *** *** *** ***

 *** insufficient toxic response to calculate EC50 or EC20
21

MONITORING DATA

BAY-WIDE SURVEYS

Bay-wide water and sediment samples were collected to compare copper concentrations in

samples collected during hull cleaning activity to those in open bay areas. Samples were

collected onboard the R/V ECOS by a flow-through (5 liters per minute) pumping system

composed of Teflon tubing and an all-plastic centrifugal pump head. Samples collected from

the hull cleaning plume were collected by a submerged rubber hose connected to an impellerless

deck pump onboard an 18-foot skiff. The following procedures discuss sample preparation and

analysis for bay-wide survey samples as well as for hull cleaning samples. Sample stations are

shown in figure 6. Sample collection dates and codes per ship cleaned are listed in table 4.







	















Figure 6. Bay-wide sample station locations in San Diego Bay.
22

Table 4. In-water hull cleaning surveys.

Vessel Paint Type Date Survey Code

USS DURHAM (LKA 114) ABC-3 15–16 May 91 SDC1

USS ENGLAND (CG 22) BRA-540 8–9 August 91 SDC2

USS MARVIN SHIELDS (FF 1066) ABC-3/BRA-540 22 June 92 SDC3

USS FORT FISHER (LSD 40) F-121 23 February 93 SDC4

USS TUSCALOOSA (LST 1187) BRA-540 26 February 93 SDC5

USS RANGER (CV 61) BRA-540 22 March 93 SDC6

WATER SAMPLES

Seawater samples were collected in 125-ml precleaned (I-Chem) Nalgene HDPE bottles

and were filtered using 115-ml capacity Nalgene analytical filter units, with 0.45-m cellulose

nitrate filters. Filters were retained for analysis from surveys 1 and 2 for particulate [Cu], which

was added to the dissolved fraction [Cu] for total [Cu]. For surveys 3 through 6, nonfiltered

seawater was retained for total [Cu], and particulate [Cu] was determined indirectly by subtracting the dissolved fraction [Cu]. All water samples were acidified to an approximate pH of 1.6

(0.2% v/v Ultrex HNO3) and stored frozen. The term “dissolved” is used in this report to refer

to the sample passing the 0.45-m filter unit. It is so used to facilitate comparison with literature

data and is not meant to imply that this fraction is actually that containing only dissolved copper.

The discussion of ion-specific electrode measurements that follows in this document addresses

dissolved copper in detail.

Water samples were analyzed using a Perkin-Elmer Zeeman 5100 Graphite Furnace

Atomic Absorption Spectrometer. L’vov platforms were used in pyro-coated tubes. Survey 1

and 2 water and survey 2 dome leachate samples were analyzed by direct injection using

Pd(NO3)2 and Mg(NO3)2 as matrix modifiers. Concentrations were determined from a 6-point

calibration curve (0, 5, 10, 20, 50, 100 g L-1). Standards were created instrumentally from 20

and 100 g L-1 stock solutions made serially from a 1000 g ml-1 reference standard solution.

Recovery percentages and recalibrations (based on 20 g L-1 standard) typically were performed after every 10 samples.

Prior to analysis, surveys 3 through 6 water and dome samples were extracted using the

APDC-MIBK (ammonium pyrrolidinedithiocarbamate; methyl isobutyl ketone) method, for the

purpose of matrix cleanup, and, in the case of the SDC3 dome leachate samples, for dilution-extraction. A 10-ml volume of APDC solution (generally 1% w/w) was made daily from crystalline APDC in 18-M double-deionized water (DIW) (any reference to DIW in this section refers

to this water), and cleaned 3 to 4 times with 1-ml aliquots of MIBK prior to use. The APDC

solution was used at 0.125-mg APDC per ml sample extracted (i.e., 25 l of 1% APDC per ml

sample). Survey samples were extracted at the proportion of 2 ml sample: 2 ml MIBK, as were

dome leachate samples from SDC4-6. SDC3 dome leachate sample: MIBK ratios varied

depending on the expected concentration, from 1:3 to 0.2:4. Survey 3 water and dome leachate

sample concentrations were determined using method of addition calibration. Additions of 5,

12.5, and 25 g L-1 were made to the first of each group of survey samples to develop the
23

calibration curve for that group. Dome samples were processed in three groups and concentrations added to the first sample in each group varied depending on the expected concentration

range. Surveys 4 through 6 water and dome sample concentrations were determined from a

6-point calibration curve generated using extracted standards made from MIBK-cleaned

seawater (0, 5, 10, 20, 50, and 100 g L-1). Samples out of range of the calibration curve were

diluted instrumentally with MIBK.

Nonfiltered plume samples were diluted 0.1 ml:5 ml DIW (pre-acidified DIW, 0.2% HNO3),

and concentrations were determined from a 4-point calibration curve generated from filtered,

pre-acidified seawater standards (0, 1000, 2000, and 3000 g L-1) that were then diluted in the

same manner as the samples. Extraction was not necessary because of the large dilution factor.

FILTERS

Filters were removed from the filter units with Teflon forceps and stored frozen in 50-mm

disposable petri dishes. Because the filters themselves contained some copper, 12 filters from

the only lot used for surveys 1 and 2 were digested and analyzed to determine this background

[Cu]. All filters were digested with 10-ml Ultrex HNO3 in lined digestion vessels (CEM).

Digestates were diluted 1:9 with DIW. Concentrations were determined from a 6-point calibration curve generated from standards (0, 5, 10, 20, 50, and 100 g L-1) made instrumentally from

20 and 100 g L-1 stock solutions (in 10% HNO3 v/v in DIW), which were made serially from a

1000 g ml-1 reference standard solution.

SEDIMENTS

Sediments were stored frozen in precleaned (I-Chem) Nalgene HDPE bottles or PE

zipper closure bags. After thawing, samples were homogenized before and between sub-sampling. Three aliquots of each sample were dried at 250C, ground with a Diamonite mortar and

pestle, and placed in X-ray cells with mylar window film. Samples were analyzed using a

Spectrace 9000portable X-Ray Fluorescence Spectrometer for 200 seconds on the 109Cd

source. The National Research Council of Canada (NRCC) standard reference material PACS-1

was analyzed routinely for comparative purposes.

Sediment samples were collected with either a Van Veen stainless steel sampler from the R/V

ECOS, or by a cylindrical corer fitted with a plastic inner cylinder. Samples were placed in

zip-lock plastic bags and stored frozen until grain size analysis was performed. Samples were

passed through a series of standard screens, and subsequently hydrometer (ASTM No. 152H)

readings were taken to define the percentage sand, silt, and clay fractions (Black, 1965).

WATER COLUMN MEASUREMENTS

Water column samples were collected concurrently with hull cleaning operations in order to

compare background levels of copper throughout the bay with those measured in the vicinity of

hull cleaning operations. The bay-wide water column copper data are summarized in figures 7

through 12. Copper measured in the bay-wide samples is referred to as “dissolved” copper again

simply to indicate that this fraction passed a 0.45-m filter and to facilitate comparisons with

literature values. Measurements of dissolved copper are discussed in the “Ion Selective Electrode” section of this document.
24

Figure 7. Mean bay-wide surface water dissolved copper concentrations in g L-1, May

1991.

Figure 8. Mean bay-wide midwater dissolved copper concentrations in g L-1, May 1991.
25

Figure 9. Mean bay-wide deep water dissolved copper concentrations in g L-1, May 1991.

Figure 10. Mean bay-wide surface water dissolved copper concentrations in g L-1, June 1992.
26

Figure 11. Mean bay-wide surface water dissolved copper concentrations in g L-1,

February 1993.

Figure 12. Mean bay-wide deep water dissolved copper concentrations in g L-1,

February 1993.
27

Of the 58 measurements (sum of measurement intervals over all stations reported in figures 7

through 12) made throughout the bay during the three hull cleaning operations, 24 were at or

below 2.9 g L-1, which is the EPA 1-hour average water quality criterion for copper. The

Shelter Island (Station 9) and Glorietta Bay (Station 8) yacht harbor stations exhibited intermediate to high copper values relative to other stations. The highest concentrations recorded in these

areas were 6.9 and 5.6 g L-1, respectively. These values are similar to the mean dissolved

copper concentration reported by Krett (1980), but not as high as the 8.9 to 11 g L-1 dissolved

copper levels reported in Johnston (1990). Both the Krett and Johnston studies measured

“dissolved” copper by ASV in water samples from Shelter Island. Labile copper values

measured at pH 2.2 by ASV, and copper extracted on Chelex 100-AA and measured by atomic

absorption compared very well (Zirino et al., 1978), suggesting that labile copper concentrations

measured by ASV at pH 2 may also be similar to copper concentrations measured by atomic

absorption after separation using a 0.45-m filter. It appears that “dissolved” copper concentrations may not have changed a great deal in the bay over a period of 10 years.

Background water column copper levels in the main channel adjacent to the Naval Station

(Station 5) ranged from 2.8 to 5.8 g L-1 with an overall average value of 3.8 g L-1. The

majority of the main channel samples exceeded the EPA water quality criterion of 2.9 g L-1. As

mentioned above, approximately 50% of the samples analyzed during the bay-wide study

exceeded this criterion indicating that copper concentrations are elevated throughout the bay. If

compared to earlier studies, it appears that this elevation may have been so for a decade or

possibly longer. The constant input by leaching of copper from the antifouling paints of

pleasure, commercial, and military vessels is a likely primary source of copper in the bay. This

input will be discussed further in this document under the topic of “Loading Factors.”

SEDIMENT MEASUREMENTS

Copper concentrations in bay-wide sediment samples collected during hull cleaning operations are summarized in figures 13 through 15. Concentrations were quite variable exhibiting no

particular patterns. The sediment copper concentrations measured at the main channel station off

the Naval Station (Station 5) ranged from 132 to 268 mg kg-1 and were similar to those reported

by Yamamoto et al. (1975) and NOAA (1991) from sediment samples collected yet closer to the

Naval Station in the early 1970s and 80s. Copper in sediment collected from Station 8 (table 1)

in the early 1970s was much lower (28 mg kg-1) than that reported in the present study at the

main channel off the Naval Station (Station 5) and approximate to the location of Station 8.

Sediment copper concentrations at the Shelter Island yacht harbor and PACO Ore terminal were

relatively high, ranging from 133 to 212 and 140 to 647 mg kg-1, respectively.

SEDIMENT GRAIN SIZE ANALYSIS

Sediment grain size analysis was performed to identify sample composition and permit

comparison to literature values. A sediment grain size analysis summary is presented in table 5.

Sediments at the main channel station near the Naval Station were relatively high in percentage

clay indicating a high capacity for binding copper.
28

Figure 13. Mean bay-wide sediment copper concentrations in mg kg-1

dry weight, May 1991.

Figure 14. Mean bay-wide sediment copper concentrations in mg kg-1

dry weight, August 1991.
29

Figure 15. Mean bay-wide sediment copper concentrations in mg kg-1 dry weight,

February 1993.

Table 5. Sediment grain size analysis summary. Data as percent of sample.

Coarse

Sand Sand Silt Clay

Sample

Number >2 mm

1.00–

2.00

0.500–

1.00

0.250–

0.250

0.125–

0.250

0.063–

0.125

0.046–

0.063

0.026–

0.046

0.015–

0.026

0.009–

0.015

0.005–

0.009

0.003–

0.005

0.001–

0.003 <0.001

Bay Entrance Channel

01–S1 1.3 0.3 0.6 2.4 71.9 19.9 0.3 0 0 0 2.5 0 0 0.2

01–S2 0.9 0.3 0.4 5.8 76 17.3 0.5 <1

01–S3 0.7 0.3 0.4 1.6 73.4 21.7 0.8 <1

Shelter Island Bay Channel

02–S1 0.3 0.4 1.6 5.1 55.7 17.1 2.5 1.0 2.9 0.7 0.3 2.5 0.7 7.1

02–S2 0.1 0.1 0.2 2.6 53.8 23.1 2.6 1.7 1.8 2.9 2.1 1.1 1.1 4.6

02–S3 0.8 0.4 0.8 3.6 48.2 24.4 2.9 1.4 1.0 1.8 2.4 1.4 1.5 7.0

Harbor Island Bay Channel

03–S1 1.6 1.8 4.1 15.3 32.1 21.7 3.3 5.1 3.6 1.8 1.8 1.1 1.8 10.5

03–S2 1.6 1.7 3.0 9.8 30.7 23.6 4.4 3.7 3.7 3.4 2.2 1.8 1.5 8.9

03–S3 1.8 1.8 5.1 19 33.9 16.4 3.0 1.8 1.6 2.8 1.0 2.1 1.0 7.2

Coronado Bay Bridge

04–S1 0.3 0.7 1.5 9.2 33.8 21.4 4.4 6.6 2.6 3.1 3.5 2.6 2.2 14.9

04–S2 * * sand fraction

lost * 4.0 1.1 3.5 3.1 3.6 3.5 4.1 14.3

04–S3 0.3 0.8 1.8 10.3 45.3 15.2 2.9 1.1 2.1 3.6 1.0 1.5 1.1 10.3
30

Table 5. Sediment grain size analysis summary. Data as percent of sample (continued).

Coarse

Sand Sand Silt Clay

Sample

Number >2 mm

1.00–

2.00

0.500–

1.00

0.250–

0.250

0.125–

0.250

0.063–

0.125

0.046–

0.063

0.026–

0.046

0.015–

0.026

0.009–

0.015

0.005–

0.009

0.003–

0.005

0.001–

0.003 <0.001

Naval Station Bay Channel

05–S1 0.3 0.3 0.3 4.8 29.2 11.4 3.6 6.0 8.3 5.5 5.0 2.8 5.0 17.9

05–S2 0.4 0.0 0.4 3.3 18.4 13.1 2.0 11.6 9.1 7.6 5.1 5.0 5.0 25.3

05–S3 0.3 0.1 0.3 1.3 12.1 15.0 5.1 9.8 10.3 8.3 7.7 3.1 4.7 21.6

South Basin

06–S1 5.9 1.0 2.0 13.4 26.5 12.8 3.6 0.8 3.6 3.2 3.9 3.6 4.0 18.6

06–S2 2.1 0.8 0.9 9.2 30.9 13.5 4.2 0.9 3.3 4.2 4.2 3.4 4.2 21.4

06–S3 2.4 1.0 2.7 10.0 27.5 16.5 4.8 4.1 3.2 3.3 3.3 1.6 3.6 16.7

Paco Ore Terminal

07–S1 0.5 0.1 0.3 2.5 23.6 20.6 3.6 2.5 5.3 4.4 5.7 4.99 4.5 20.3

07–S2 0.6 0.2 0.8 4.2 25.1 16.3 4.1 6.3 4.5 5.4 4.5 3.6 4.9 17.6

07–S3 0.1 0.1 0.2 1.3 13.6 16.5 5.7 7.1 7.0 5.2 5.7 4.2 5.2 25.0

Glorietta Bay Entrance

08–S1 0.1 0.1 0.1 0.9 20.2 46.5 7.2 1.5 3.6 3.2 3.6 2.2 2.5 9.7

08–S2 0.1 0.1 0.1 0.7 12.1 40.0 8.2 5.1 7.3 5.1 4.3 1.7 4.7 14.5

08–S3 0.1 0.1 0.1 1.1 14.8 47.5 7.0 0.8 4.6 5.1 3.0 2.0 1.5 13.6

Shelter Island Yacht Harbor

09–S1 0.1 0.1 0.1 1.3 13.3 15.9 5.7 6.9 11.8 8.7 6.0 4.4 3.9 20.0

09–S2 0.1 0.1 0.1 1.0 12.7 17.9 6.5 9.0 10.4 8.5 5.0 3.1 4.5 16.7

09–S3 0.0 0.05 0.05 1.2 26.8 22.8 4.8 6.0 6.8 5.6 2.4 2.4 2.4 12.4

USS DURHAM (LKA 114)

S4 0.2 0.2 1.7 13.3 9.4 2.1 12.1 12.6 8.5 5.7 4.8 5.2 21.1

S5 5.8 7.2 6.7 9.9 28.7 5.9 2.4 2.0 6.0 4.8 4.8 3.2 4.4 12.4
31

LOADING FACTORS

BRUSHING ACTIVITY AND PLUME CONCENTRATIONS

The discrete sample data collected during hull cleanings are summarized in figures 16

through 25. Only filtered (“dissolved”) sample copper concentrations are presented since this

fraction most closely approximates the copper available to organisms. It is not intended,

however, to imply that all copper in the filtered samples is available. (See Ion Selective Electrode discussion, p. 44 and tables 7a and 7b.) Reference to the EPA marine water quality criteria

for copper of 2.9 g L-1 is made to characterize the existing and post cleaning copper levels

relative to the current regulatory standard.

USS DURHAM (LKA 114)

Surface water dissolved copper concentrations decreased rapidly with distance from the

ship’s hull during cleaning operations (figure 16). Before cleaning and after cleaning surface

water copper concentrations were essentially the same. All surface water copper concentrations

measured were greater than the EPA water quality criteria of 2.9 g L-1, including those

measured in surface water before cleaning operations commenced. Samples collected at

midwater depth remained essentially the same before and after cleaning operations, ranging from

2.5 to 8.7 g L-1 (figure 17). Comparatively, samples collected before cleaning were similar to

those collected after cleaning 1/2 to 2 hours later. Samples collected along the middle transect

(where most of the measurements were made) exhibited a decrease in copper concentration from

19.1 to 5.3 g L-1 at 18 and 96 m from the ship’s hull, respectively. Four of 27 samples collected

before, during, and after hull cleaning at midwater depth were at or below the copper water

criteria of 2.9 g L-1.

If the copper concentrations during hull cleaning and distance from the hull are regressed, the

mean water copper concentration prior to cleaning (4.3 g L-1) is met at 101 m from the hull.

Data from the middle transect at midwater depth were used since these measurements gave the

best correlation with copper concentration and distance. The area of hull cleaning influence may

be estimated to be a cylinder with a radius of 101 m and a depth of 9 m. The area is calculated to

be 288,280 m3, or approximately 0.001% of the bay volume (taken to be 2.5 x 108 m3).

USS ENGLAND (CG 22)

During the hull cleaning of USS ENGLAND, surface water dissolved copper concentrations

ranged from levels approximating before and after hull cleaning concentrations to levels nearly

twice initial concentrations (figures 18 and 19). All surface water copper concentrations were

above 2.9 g L-1. Midwater copper concentrations measured during hull cleaning were not

generally highly elevated above pre-cleaning values. At 42 m from the hull, along the middle

transect, a sample collected during cleaning was nearly three times the pre-cleaning level. At

10 m from the hull, a post-cleaning sample was approximately twice the pre-cleaning concentration. The sample collected during cleaning at 10 m was, however, less than the concentration

measured prior to cleaning. Clearly the fixed-point sampling performed gives only an approximate characterization of the water column copper concentration. Many more discrete samples

would be necessary to clearly define the entire area where elevated copper concentrations

occurred. All samples measured were above 2.9 g L-1 copper.
32

Figure 16. USS DURHAM in-water hull cleaning, May 1991. Mean surface water

dissolved copper concentrations in g L-1.
33

Figure 17. USS DURHAM in-water hull cleaning, May 1991. Mean midwater dissolved

copper concentrations in g L-1.
34

Figure 18. USS ENGLAND in-water hull cleaning, August 1991. Mean surface water

dissolved copper concentrations in g L-1.
35

Figure 19. USS ENGLAND in-water hull cleaning, August 1991. Mean midwater

dissolved copper concentrations in g L-1.
36

USS MARVIN SHIELDS (FF 1066)

Less copper data are available for USS MARVIN SHIELDS hull cleaning operation. This

paint-test vessel was painted primarily with equal distributions of ABC-3 and BRA-540 (figure

20). Samples collected during hull cleaning were generally slightly higher in copper concentration than those collected prior to cleaning. At 2 m from the hull, approximately twice as much

copper was measured during cleaning than prior to cleaning. Dissolved copper concentrations in

all samples, including pre-cleaning samples, were above 2.9 g L-1.

USS FORT FISHER (LSD 40)

Copper data are available only during USS FORT FISHER hull cleaning operation (figure 21). Dissolved copper concentrations measured ranged from 4.1-5.1 g L-1. These levels

are similar to pre-cleaning levels measured during other hull cleanings, suggesting that little

dissolved copper was released during USS FORT FISHER hull cleaning. This pattern was

expected, as USS FORT FISHER was coated with Formula 121 antifouling paint 6 years

previously.

Sediment samples collected prior to cleaning and after hull cleaning operations did not show

any indication of increased copper in sediments as a result of cleaning operations (figure 22).

The notably higher sediment copper value measured at point E (1037 mg kg-1) suggests that

paint particles may have been abundant in this sample.

USS TUSCALOOSA (LST 1187)

Copper data collected during the hull cleaning of USS TUSCALOOSA ranged from 1.2 to

8.4 g L-1 (figure 23). Less copper was measured during this hull cleaning operation than

during others. It is possible that the paint coating on this vessel was nearly exhausted and well

beyond its operational life. Five of eight samples measured had less than 2.9 g L-1 dissolved

copper.

Of five prior and after cleaning sediment sample comparisons, only one location (A)

indicated a clear difference in sediment copper values after cleaning operations (figure 24). The

other locations had similar copper concentrations before and after cleaning operations. The

sediment copper concentrations measured at the site of USS TUSCALOOSA and USS FORT

FISHER hull cleanings were higher than those measured during the bay-wide surveys off the

Naval Station (Station 5), but similar to values measured in this same area several months later

(figure 3; Richter unpublished data). The sediment copper data collected in the Naval area at

Piers 12 and 13 are higher than those measured in these areas in the mid-1970s (Yamamoto et al.,

1975) and may indicate a long-term increase due to ship presence and cleaning activity or

proximity to the PACO Ore Terminal.

USS RANGER (CV 61)

Dissolved copper concentrations measured in hull cleaning samples collected near USS

RANGER were also similar to previously measured ambient values ranging from 2.4 to

5.6 g L-1 (figure 25). Seven of nine samples collected were less than 2.9 g L-1 copper.

ION SELECTIVE ELECTRODE/VOLTAMMETRIC SPECIATION

“Dissolved“ copper can exist in estuarine or ocean water in many forms: as the free ion or as

a labile or inert complex. The complex itself may be either organic or inorganic, or possibly,
37

Figure 20. USS MARVIN SHIELDS in-water hull cleaning, June 1992. Mean midwater

dissolved copper concentrations in g L-1.
38

Figure 21. USS FORT FISHER in-water hull cleaning, February 1993. Mean midwater

dissolved copper concentrations in g L-1.
39

Figure 22. USS FORT FISHER in-water hull cleaning, August 1991. Mean sediment

copper concentrations in mg kg-1 dry weight.
40

Figure 23. USS TUSCALOOSA in-water hull cleaning, February 1993. Mean midwater

dissolved copper concentrations in g L-1.
41

Figure 24. USS TUSCALOOSA in-water hull cleaning, August 1991. Mean sediment

copper concentrations in mg kg-1 dry weight.
42

Figure 25. USS RANGER in-water hull cleaning, July 1993. Mean surface water

dissolved copper concentrations in g L-1.

both. The inert forms are most likely colloidal (Mackey and Zirino, 1994). The toxicity of the

different fractions may vary; indeed, the work of Sunda and Guillard (1976) and Anderson and
43

Morel (1978) have shown that it is the concentration of free ion (and not the total concentration

of Cu) that correlates best with plankton toxicity.

To assess the environmental impact of hull cleaning operations, we have applied some newly

developed speciation tests to samples collected during the hull cleaning procedure. Similarly, we

have used the Orion Cu ion selective electrode (ISE) to detect the release of Cu++ during hull

cleaning and to determine the base or ground level of free Cu++ at various locations in San Diego

Bay.

The relationship between the millivolt output of the Orion Cu++ ISE and free Cu++ was

established by calibrating the electrode with Cu++ activity buffers prepared according to Belli

and Zirino (1993). These workers also showed that the electrode was able to provide accurate

measurements of the free Cu++ activity or concentration in seawater media and demonstrated

that the problem known as the “chloride interference problem” was likely due to the incorrect

usage of the electrode by previous workers.

We have also measured hull cleaning and ambient waters in San Diego Bay for certain

copper fractions according to a speciation scheme published by Scarano et al., (1990). This

voltammetric method detects the “labile” fraction of Cu present at pH 8, the “ligand exchangeable” fraction, which includes the labile fraction and that portion of the “bound” fraction that

would exchange with ethylenediamine (En-exchangeable, Scarano et al., 1992), and the “inert”

fraction, e.g., that portion of the sample that can only be detected by lowering the pH to 2.

It should be noted that the speciation scheme above separates the fractions by increasing the

“strength” of the chemical treatment, and if a sample is composed entirely of labile Cu, then all

three voltammetric fractions would have identical concentrations. On the other hand, the

potentiometric measurements do indeed measure the free Cu++ ion. Even the most labile

complexes, such as CuCl2, are not detected by the ISE. Thus, Cu++ activity values detected by

potentiometry are much smaller than any of the fractions measurable by voltammetry.

Inorganic models may be constructed that describe the speciation of seawater in terms of its

inorganic components (Zirino and Yamamoto, 1972). All of these models essentially predict that

the concentration of the free Cu++ ion is approximately 1% of the total copper concentration.

Therefore, for a seawater concentration of 510-8 M (3.2 g L-1), the activity of the free ion

should be approximately 510-10 (0.032 g L-1), or pCu = 9.3.

THE RELATIONSHIP BETWEEN Cu++ AND TOXICITY

As was stated above, the argument for Cu++ ion activity being a more reliable measure of

toxicity than the total copper ion concentration goes back to the work Sunda and Guillard (op.

cit.). They showed that cultures of the estuarine diatom Thalassiosira pseudonana (clone 3H)

and of the estuarine green alga Nannochloris atomus in highly chelated media showed growth

rate inhibition at calculated Cu++ activities above 310-11 (1.9 ng L-1). Growth rate inhibition

was not related to the total Cu concentration that ranged from 10-3 M to 10-6 M (63.5 mg L-1 to

63.5 g L-1). Cu++ ion activity was calculated because it could not be measured directly at that

time. Two years later, Anderson and Morel (op. cit.) repeated Sunda and Guillard’s experiment

using cultures of the marine dinoflagellate Gonyaulax tamarensis. They also found that a Cu++

activity level of 410-11 (2.5 ng L-1) also inhibited motility. They pointed out that for some

marine environments, this value was relatively close to the value that could be calculated

assuming only inorganic complexation.
44

MEASUREMENTS IN SAN DIEGO BAY

USS DURHAM (LKA 114)

Cu++ activity measurements were made in the field during the last five hull cleaning

operations. Initial measurements were made during the hull cleaning operations of USS

DURHAM. Figure 26 shows the raw millivolt response of the Cu++ ISE during the hull cleaning

event of May 16, 1991. For this operation, the sensor was actually placed on the R/V ECOS, 9

m from the side of USS DURHAM. Seawater near USS DURHAM was collected by holding

the open end of a 50-m Teflon hose near the hull and aspirating the sample through the hose to

the electrodes on the ECOS. The effect of brushing on the ambient water is clearly evident in the

figure, producing a peak from about 10:30 to 15:00. The peak in potential (mV) values also

coincides with a peak in the plot of the discrete samples collected at the same time and analyzed

by ASV (table 6). However, pCu values could not be assigned at this time because seawatercopper buffers had not yet been developed.

Figure 26. Ion Selective Electrode response profile, USS DURHAM in-water hull

cleaning, 16 May 1991. (A) Initial hull cleaning pass; (B) slack low tide; (C) hull

cleaning operations end.

Nevertheless, assuming an ideal electrode response of 29.6 mV decade-1, we see that during

the hull cleaning operation, Cu++ values rose by about two orders of magnitude. The high values

did not persist, however, and dissipated after the hull cleaning ceased (table 6).

That the electrode response reflected the dissolution of hull cleaning residue is confirmed by

table 6, which shows the results of the ASV analysis of the discrete samples collected at various

distances from the hull. It is clearly evident that the ambient water has been significantly

(though temporarily) impacted. The “labile” content of the water next to the ship is seen to

increase by two orders of magnitude, in agreement with the ISE results.
45

Table 6. Results of samples collected before, during, and after hull cleaning operation, 16 May 1991. All copper concentrations in parts per billion (ppb).

Sample Labile EnexchangableAcidexchangableAA (F) Comments

3 METERS FROM SHIP HULL

CSU-67 217 227 30.1 Approx. 30 min after hull clean. ops.

began

CSU-85 12.4 26.2 151.2 12.1 Approx. 45 min after scrubber pass

CSU-103 3.5 6.3 10.1 4.6 Approx. 40 min after hull clean. ops.

ended

CSU-121 6.0 10.3 28.9 11.1 Approx. 2 hrs after hull clean. ops.

ended

21 METERS FROM SHIP HULL

FSU-11 4.9 8.9 49.2 2.9 Pre hull cleaning (5/15/91)

FSU-73 16.3 16.3 12.2 6.7 Approx. 30 min after scrubber pass

FSU-127 4.6 10.9 12.7 3.6 Approx. 2 hrs after hull clean. ops.

ended

45 METERS FROM SHIP HULL

HSU-33 1.8 4.7 9.1 Pre hull cleaning (5/15/91)

HSU-55 2.4 3.6 11.8 3.8 Pre hull cleaning (5/16/93)

HSU-95 9.4 15.6 67.3 10.8 Approx. 45 min after scrubber pass

HSU-131 5.7 9.0 4.2 Approx. 2 hrs after hull clean. ops.

ended

Discrete samples were also collected during the week of the USS DURHAM hull cleaning

operation and were measured for the various fractions: labile, ethylenediamine exchangeable,

and acid exchangeable. Some were associated with the hull cleaning operations while others

were taken around San Diego Bay in a bay-wide survey. The first are presented in table 7a,

while the results of the bay-wide survey are presented in table 7b. With the exception of a few

discrepancies, a clear pattern emerges from all the samples:

labile < ion exchangeable < acid-exchangeable

We notice that data from the bay-wide survey indicate that the acid-exchangeable fraction is

approximately twice the labile fraction with the En-exchangeable fraction being somewhere in

between. This pattern appears to be typical of Cu in San Diego Bay waters and is supported by

additional data obtained by Venderweel (in preparation). However, near ship operations, the acid

exchangeable value is often five times higher than the En-exchangeable fraction. We suggest

that this result is due to the presence of particulate copper, presumably from paint particles.
46

Since the paint is largely of copper oxide, the oxide particles would be dissolved quickly by the

acid and not by the ethylenediamine. In general, analysis of the same samples by atomic

absorption (AA) did not produce the very high values, but gave concentrations similar to the

En-exchangeable fraction. Because the AA determination of the samples occurred after filtration

through a 0.45-m filter, this also suggests the presence of particulates.

Table 7a. Samples collected near and around pier area, USS DURHAM, 17 May

1991. All copper concentrations in parts per billion (ppb).

Sample Labile En-exchangable Acid-exchangable Location

ZMU-1 6.1 8.2 14.8 Outside of pier

ZMU-5 9.6 16.0 20.9 Outside of pier

ZMU-11 4.5 5.9 8.2 End of pier

ZMU-14 13.0 17.6 21.5 Inside pier area

ZMU-20 12.2 15.0 20.5 Inside pier area

Table 7b. Results of bay-wide survey, 21 May 1991. All copper concentrations in

parts per billion (ppb).

Sample Labile EnexchangableAcidexchangableAA (F) Location

CSD1SU-29 5.3 8.8 10.3 4.7 Bay-mouth

CSD2SU-32 1.6 2.0 4.5 3.6 Center of channel btw Shelter

Is. and NAS

CSD3SU-35 2.1 3.6 N/A 1.8 Center of channel, south of

Harbor Is.

CSD4SU-38 2.3 3.1 5.0 3.3 Center of channel, under

Coronado Bridge

CSD5SU-41 3.9 4.8 6.3 4.0 Off of San Diego Naval Station

CSD6SU-44 3.7 3.9 2.1 1.9 South San Diego Bay

CSD7SU-47 1.7 2.4 2.0 2.1 Overload Facility

CSD8SU-50 4.7 6.2 9.7 3.7 Glorietta Bay

CSD9SU-53 1.6 2.9 3.1 3.1 Shelter Island

USS MARVIN SHIELDS (FF 1066)

Participation in the USS MARVIN SHIELDS hull cleaning survey was limited to making 

Cu ISE measurements before, during, and after the hull cleaning operation (June 22, 1992).

Once again, the R/V ECOS was stationed about 25 m off MARVIN SHIELD’s beam, and the 

Cu ISE was aboard the ECOS. Using a 25-m Tygon hose, water was sampled at 2, 5, 10, and

20 m from the hull and pumped to the ECOS where it was measured for copper activity. This

time, the electrode was calibrated for activity by using a single ethylenediamine buffer and

varying the pH. The electrode potential became more positive after the hull cleaning event.
47

Also measurements closest to the ship were more positive, indicating more available copper.

Unlike for USS DURHAM, only a 15-mV increase was registered, indicating a six-fold increase

in free Cu ion next to the hull during the event. pCu values (the negative logarithm of the free

Cu++ activity) were in the range of approximately 12 with high Cu++ of 11.8 (equal to 1.610-12

M, or 0.1 ng L-1) to a low of 12.7 (equal to 2.010-13 M, or 0.013 ng L-1). The potential of the

electrode returned to background value after about 30 minutes. No samples were analyzed by

voltammetry.

USS FORT FISHER (LSD 40)

For the USS FORT FISHER hull cleaning event, the electrode was mounted on a CTD

(conductivity, temperature, depth) probe (IDRONAUT 301 – Milan, Italy) with direct readout

to a portable computer via a cable. The Cu-electrode-equipped CTD, which also measures pH

and dissolved oxygen, was hand-lowered in the water from a small boat situated in the hull

cleaning regions. Profiles were taken throughout the hull cleaning operations at various

distances from the hull.

Calibrations of the electrode were effected using the Belli and Zirino (1993) buffers. The

IDRONAUT 301 possesses a reference electrode of unusual stability for a field device. This

makes it possible to compare Cu ISE readings over long periods of time. We carried out calibrations of the electrode in the laboratory and in the field over a 6-month period and found that

the Cu ISE electrode potential readings in freshly made buffers varied only by a few millivolts.

There is also very little drift with the pH electrode associated with the CTD. Thus, a secondary

measure of the stability of the reference is whether the pH electrode on the CTD returns the pH

value of the fresh ethylenediamine and glycine buffers (8.22). Electrode potentials were related

to Cu++ ion activity using the formula:

pCu = (–65 mV – (– potential reading))/(29.6 mV) + 13.33

Three approximately equidistant stations were located along the starboard side of USS FORT

FISHER and were sampled approximately three times during the hull cleaning. Profiles were

made from surface to near bottom. In summary, all profiles for all variables indicated a wellmixed water column. Cu ISE values varied from about –13 mV ( pCu = 11.6) nearest the ship, at

the onset of the hull cleaning, to approximately –50 mV (pCu = 12.8) during a lull in the

operation. Similar changes and ranges were noted during the USS MARVIN SHIELDS

operation.

USS TUSCALOOSA (LST 1187)

The procedure used for sampling during the USS TUSCALOOSA survey was very similar to

that used for USS FORT FISHER. Four stations were set up approximately 10 m from the

starboard hull and sampled throughout the hull cleaning procedure. Measurements were made

by lowering the CTD to near bottom at several distances (2, 5, and 15 m) from the hull.

For all stations at all depths, the electrode potentials varied between –30 mV and –33 mV,

giving a pCu of approximately 12.1. No copper was detected by ISE from this hull cleaning

operation. All stations around USS TUSCALOOSA were well-mixed. No samples were

collected for analysis by voltammetry.
48

USS RANGER (CV 61)

The final hull cleaning survey was of USS RANGER. As in the last two surveys, stations

were set up along the port side, under the flight deck and sampled periodically during the hull

cleaning operation. Sampling was effected at each station by manually lowering the CTD from a

small boat.

All stations showed a nearly isopycnal water column and Cu ISE potentials varied from

–11 mV to –14 mV, giving a pCu value of approximately 11.6. A sample of the “red water”

discharged by the hull cleaning “brush” was trapped and measured directly with the electrode on

the CTD. The potential immediately moved toward the positive and in a few minutes, the

potential rose to +85 mV (pCu = 8). No samples were collected for analysis by voltammetry.

RESULTS AND DISCUSSION

The results of our work with the Cu ISE indicate that the free Cu++ level in San Diego Bay is

approximately 10-12 M (0.06 ng L-1; pCu = 12) and that the in situ hull cleaning procedure may

temporarily raise the pCu to about 11, but that the impacted waters quickly return to the ambient

level in a matter of hours or less. This strongly suggests that the hull cleaning debris is largely

particulate and that its contact with the ambient waters is brief before settling to the bottom. In

the cases of USS TUSCALOOSA and USS RANGER, no impact from hull cleaning could be

noted at all, even a few meters from the hull.

A pCu value of near 12 is at least one order of magnitude below the toxicity value for

microalgae established by Sunda and Guillard (1976) as well as Anderson and Morel (1978).

Thus, by this criterion, the temporary changes in pCu due to hull cleaning cannot be construed to

be a danger to microalgae. In fact, a pCu of 12 lies in the optimum region established by Sunda

and Guillard (1976), for the growth of Thalassiosira pseudonana. Higher values of Cu++ are

toxic, and lower values lead to a nutrient deficiency.

Interesting is the relatively large fraction of labile Cu present in the bay as determined by

stripping voltammetry. This fraction is present at pCu from 7 to 8. At this point, it is important

to define labile in terms of the over-voltage applied to the electrode. For this work, we applied a

reducing potential of –1.0 V to the hanging mercury drop electrode (HMDE). This constitutes

an over-voltage of approximately 500 mV over the reduction of copper in organic free seawater.

Scarano et al. (1992) showed that for Tyrrhenian Sea samples, this over-voltage did not produce

significantly higher currents than an applied potential of 500 mV, e.g., there were no additional

voltammetrically labile compounds to be reduced in addition to those already reduced at 500

mV. In San Diego Bay, the labile fraction turned out to be anywhere from 50 to 70% of the

ethylenediamine exchangeable fraction. This is exactly the same fraction observed for the

Tyrrhenian Sea samples.

Thus, the labile fraction in San Diego Bay represents easily reduced copper compounds;

compounds as easily reduced as CuCl2. This poses two questions: (1) Is the labile fraction as

toxic as Cu++? and (2) What is its composition? What raises the pCu to 12 and is labile to

voltammetry? (In San Diego Bay, inorganic complexation, as predicted by speciation models,

can only take the pCu to about 10).

The answer to the first question appears to be “no.” The labile fraction is not toxic to

microalgae. From the criteria developed by Sunda and Guillard (1976) and Anderson and Morel
49

(1978) for cultures, the values found in San Diego Bay would completely inhibit growth, and

thus, there would be no microalgae in the bay! Thus, either the labile fraction is not toxic or 

the existing microflora is very resistant. Indeed, Krett (1980) showed that copper tolerant

species of microalgae existed in the bay. Nevertheless, Krett did note toxicity at about

210-7 M (12.7 g L-1 = ppb) ionic copper for plankton samples from the Shelter Island 

inlet, a relatively polluted area. In any case, it does not appear that the labile fraction is toxic 

to microalgae in the bay.

Even a tentative answer to the second question, “What is the physico-chemical nature of the

labile fraction?” is more difficult and awaits further speciation work.

Detection of Cu+

It should be pointed out that our entire discussion has been in terms of the cupric (Cu++) ion,

the stable form in oxygenated seawater, while copper in antifouling paints is in the cuprous or +1

state. Nevertheless, our results are still valid because (1) we expect the cuprous form to oxidize

spontaneously in the water; (2) ASV does not distinguish between the two valence states; and

(3) our laboratory tests showed that the Cu ISE also responds well to additions of cuprous ion.

RELEASE RATES/SHIP LOADING FLUX

The in situ copper release rates measured with a closed recirculating polycarbonate plastic

dome system (Lieberman et al., 1985) attached to the ship’s hull by divers are presented in

table 8. Water is circulated through the closed system under a slight negative vacuum during the

period of 1 hour while samples are withdrawn for copper analysis. The copper concentration

increases with time in a linear fashion as more and more copper is leached into the closed

system. The release rate is then calculated as a linear function with time.

The copper leach rates summarized in table 8 are low and less than the required leach rate

(10 g/cm2/day) to prevent fouling of hulls (Preiser et al., 1977). The physical appearance of the

hulls tested supported this assumption as they were heavily fouled. Divers communicated that

paint appeared to be missing in some hull areas. The age of the paint tested was likely responsible for the low leach rates measured.

SEDIMENT DEPOSITION/BENTHIC FLUX CHAMBER

As has been stated previously, the availability of copper is an essential aspect of its potential

toxicity both in water and in sediments. In water samples copper partitioning and availability

have been investigated by separation techniques and chemical measurements. The availability of

copper in sediments is controlled by unique factors as well. It is generally known that if

interstitial water concentrations are higher than water concentrations above the sediments (water

column concentrations) then a pollutant contained in the sediment is available via a positive flux.

Hence, in addition to the issue of increased copper in the water column from hull cleaning

operations, that of increasing sediment copper concentrations and possible availability to

sediment dwelling organisms is also of concern.
50

Table 8. In situ antifouling paint copper release rate measurements.

USS RANGER (CV 61)

Before Brushing After Brushing

Paint Type g Cu cm2 day-1 Loading g day-1 g Cu cm2 day-1 Loading g day-1

BRA-540 1.2 158 3.8 500

5.7 748

USS FORT FISHER (LSD 40)

Paint Type g Cu cm2 day-1 Loading g day-1

F-121 Samples 4.0 188

Not Taken Due to 3.5 166

Dome Failure 9.6 454

USS TUSCALOOSA (LST 1187)

Paint Type g Cu cm2 day-1 Loading g day-1 g Cu cm2 day-1 Loading g day-1

BRA-540 3.1 100 3.0 96

2.6 84 2.9 94

USS MARVIN SHIELDS (FF 1066)

Paint Type g Cu cm2 day-1 Loading g day-1 g Cu cm2 day-1 Loading g day-1

BRA-540 5.9 52 5.5 48

5.7 50 2.8 25

APC-3 1.1 10

1.2 11 3.2 28

USS ENGLAND (CG 22)

Paint Type g Cu cm2 day-1 Loading g day-1 g Cu cm2 day-1 Loading g day-1

BRA-540 3.6 112 5.0 155

4.6 143 5.5 171

4.7 146 5.4 168

6.0 186 4.8 149
51

Since evaluating the availability of a pollutant in a sediment sample introduces problems

associated with disturbance of its physical-chemical structure and hence potentially artifactual

results, the Navy has developed a novel in situ sampling device that measures the positive or

negative flux of a contaminant from a sediment (Chadwick and Stanley, 1993; Chadwick et al.,

1993). This is accomplished by isolating a volume of water above the sediment, drawing off

samples from this volume over time, and analyzing these samples for increases or decreases in

contaminant concentration. An increasing contaminant concentration indicates that the contaminant is effusing from the sediment, while a decreasing concentration indicates that the contaminant is partitioning to the sediment. The chamber, sensors, sampling devices, control systems,

power supply, and deployment/retrieval equipment are all mounted on a tripod frame that can be

placed in situ and left to sample a site for as long as 7 days.

Sediment copper flux data collected in San Diego Bay at the Shelter Island yacht basin in

1993, at the Naval Station in 1993, and at the NRaD pier in 1990 are presented in figures 27

through 30. Although the dates and exact areas of investigation are not synoptic with the naval

hull cleaning sites, the data serve to indicate that no positive flux of copper was calculated from

these areas. The Shelter Island sediment site is relatively high in percent clay and would be

expected to bind copper. The naval site near Paleta Creek is close to the sites where ship hulls

were cleaned. Sediments collected at the hull cleaning sites also had relatively high percentage

clay content (table 5). The copper deposited in the underlying sediments after hull cleanings

would not likely become available given the occurrence of natural binding processes judging

from the in situ copper flux data collected.

ADDITIONAL NONNAVAL SITES

Municipal Yacht Harbors/Commercial Shipping

Among several principle sources of copper input to San Diego Bay, that from municipal

yacht harbors and commercial shipping must be considered significant. The presence of several

thousand pleasure craft (currently 7600 as of April 1993, San Diego Harbor Police census) in

various yacht basins throughout the bay, and routine commercial shipping activity leads to a

constant input of copper into the surrounding waters as a result of leaching release from copper

containing antifouling paints. The application of copper from antifouling paint to pleasure craft

was estimated, together with that of commercial vessels and military vessels, to total 50 metric

tons per year (Young et al., 1979) in San Diego Bay. A total of 84,000 liters of copper containing antifouling paint was estimated to be applied to pleasure craft, commercial, and military

vessels on a yearly basis. The contribution by pleasure craft is calculated to be approximately

10,000 liters per year. Since this information was collected in 1973, we have used recent yacht

census data (April, 1993) to recalculate the copper applied to pleasure, commercial, and military

vessels on a yearly basis. The total amount of copper applied using an updated pleasure craft

census is calculated to be 56 metric tons per year. The number of pleasure craft has increased

from 3600 in 1973 to 7600 presently. The estimated yearly application of copper-based antifouling paint to pleasure craft has increased to approximately 22,000 liters. We have assumed that

commercial and military inputs have remained similar, and that the copper content of antifouling

paints has also remained similar to those used in 1973.
52

Figure 27. Sediment copper flux, Shelter Island, 19 June 1993. Copper concentrations

in g L-1. Flux rate: –81 g m-2 day-1 (r2 = 0.42).

Figure 28. Sediment copper flux, Shelter Island, 25 June 1993. Copper concentrations

in g L-1. Flux rate: –58 g m-2 day-1 (r2 = 0.64).
53

Figure 29. Sediment copper flux, Naval Station, 22 March 1993. Copper concentrations

in g L-1. Flux rate: –87 g m-2 day-1 (r2 = 0.82).

Figure 30. Sediment copper flux, NRaD Pier, 1990. Copper concentrations in g L-1.

Flux rate: –158 g m-2 day-1 (r2 = 0.89).
54

With the input of copper to the bay from antifouling paints applied to pleasure, commercial,

and military vessels, it is not surprising that several sites have been reported to have total copper

concentrations in excess of the 2.9 g L-1 Federal water quality criteria level. Elevated copper

concentrations have been reported in the yacht harbor regions and Naval Station ( Zirino et al.,

1978; Krett, 1980; Johnston, 1990). The dissolved copper concentrations measured in this study

during hull cleaning activity (figures 16 through 25) were generally 10 g L-1 or less and

compared with dissolved copper concentrations previously measured in the yacht harbor and

Naval Station areas. The large contingent of pleasure craft in addition to commercial and

military vessel presence can be expected to release copper into the bay on a continuing basis. A

review of our bay wide survey data indicates that copper values will likely continue to persist

throughout the bay at a concentration of 5 g L-1 or less (figures 7 through 12).

PACO Ore Terminal

The PACO ore terminal represents a unique historical copper input to the bay since the form

of copper deposited in the surrounding sediments was a ore concentrate (cupric ferrous sulfide)

from ship loading activity over-spill (Woodward–Clyde, 1991). The Regional Water Quality

Control Board issued a Cleanup and Abatement Order (N. 85-91) to PACO Terminals, Inc.,

calling for the abatement of all copper discharge from the site and cleanup of copper-containing

sediments in the site vicinity (Woodward–Clyde, 1991). Sediment copper concentrations as high

as 58,269 mg kg-1 dry weight were measured near the pier area, thus causing concern (Woodward-Clyde, 1991). The San Diego Port Authority has taken responsibility for the cleanup

process and has agreed to clean the contaminated sediment area to a 1000 mg kg-1 (ppm)

isopleth (San Diego Port Authority, 1993, personal communication).

In the summary document cited above (Woodward–Clyde, 1991), the toxicity and bioavailability of copper from this site was evaluated. The findings of this report indicated that the

copper present in the ore deposited in the surrounding sediment of the PACO ore terminal was

not bioavailable and not toxic to marine species tested (including Pacific oyster embryos). Eight

of nine species tested exhibited no toxicity response to sediments under standardized laboratory

test conditions. The only toxic response was exhibited by the amphipod Rhepoxynius abronius.

This response was independent of the sediment copper concentration. Natural populations of the

mussel Mytilus edulis grow near the terminal in an area where some of the highest sediment

copper concentrations were measured, suggesting that the sediment copper is not available/toxic

to the mussels attached to structures in the water column. Additionally, mussels planted in the

vicinity of the terminal as part of the California Mussel Watch Program, and collected from piers

near the terminal, did not contain tissue copper concentrations significantly different from those

in mussels collected from other areas in San Diego Bay (Woodward–Clyde, 1991).

The Woodward–Clyde study (1991) concluded that the 1000 mg kg-1 copper remediation

objective for the sediments in the vicinity of the PACO ore terminal loading facility is highly

protective of aquatic life and the designated beneficial uses of the bay. A yet higher sediment

copper concentration could be accepted as well and still be protective to marine species and

other resources. The principle reason for a lack of harmful effects from the elevated sediment

copper concentration was concluded to be the unavailability and nontoxicity of the cupric iron

sulfide in the copper ore. This form of copper is highly insoluble in anoxic sediments typified

by those beneath the thin oxidized sediment layer at the surface. Additionally, there are several

precipitation, complexation, and sorption reactions that convert copper into nontoxic forms in

both anoxic and oxic waters limiting its availability to marine species. Some data are available
55

for water column copper concentrations in this area (WESTEC, 1986). Samples measured 1 and

2 m above the sediment surface at high and low tides in an area where the sediment copper

concentrations were high (19,800 mg kg-1 dry weight) ranged from 3 to 31 g L-1 as total

particulate copper (WESTEC, 1986). “Soluble” (that fraction passing a 0.45-m filter) copper

ranged from < 2-8 g L-1. The filtered water sample values are similar to those measured in our

bay-wide surveys which ranged from 2.1 to 4.8 g L-1 (figures 7 through 12).

SDG&E South Bay Power Station

In addition to input from antifouling paints, copper enters San Diego Bay waters via effluent

water from the San Diego Gas and Electric (SDG&E) Company South Bay power plant. The

effluent water must be monitored for heavy metals, including copper, at least semi-annually.

Waste-stream effluent water containing copper from boiler cleanings (removal of deposits from

the inside of the tubes) or boiler fireside washes (removal of deposits from the outside of the

tubes) represents a worst-case example of effluent. Although treated prior to discharge, this

effluent may still contain g L-1 copper concentrations. Effluent metal cleaning water would be

expected to have higher copper levels than effluent water generated during normal operations.

Total copper from waste-stream metal washing effluent water measured on a semiannual basis

during 1991–92 ranged from 0.03 to 0.31 lb day-1 (SDG&E, personal communication). No

copper was detected during one sampling interval (EPA 220.1 atomic absorption direct aspiration

measurement). The mean value of copper discharged during five monitoring intervals was

0.16 lb day-1. The NPDES permit allows a 6 month median discharge level of 25 lb day-1.

These discharge concentrations are clearly within allowed limits and compare well with the

majority of calculated copper loadings from hull cleaning activity (table 8).

Municipal/Naval Station Storm Water Runoff

Although it is a relatively infrequent event, storm runoff water may deposit a great deal of

copper into the bay during rainy periods. Data from the EPA National Urban Runoff Program

(NURP) summarized in Pitt and Field (1990) indicate that the median total copper concentration

in storm-water drainage from large U.S. cities was 35 g L-1. The concentration of copper in

apparently soluble forms exceeded 10 g L-1 in some samples. During a storm, a great deal of

copper would wash directly into San Diego Bay from the urban area via storm drains. An

estimate by the U.S. Army Corps of Engineers (1975b) indicated that the runoff from city streets

in the San Francisco Bay region could introduce 80 pounds of copper an hour to estuarine

waters. It would not seem an unreasonable assumption that storm runoff from the city of San

Diego would introduce a similar quantity of copper during storm periods.

Recent data from the Naval Station San Diego (Gadbois, 1992) indicate that storm runoff

from the Naval Station contributes effluent containing a relatively high copper concentration to

the adjacent bay area via storm drains. The runoff effluent is high in contaminants because there

is no centralized conveyance for mixing prior to release into the bay. Gadbois (1992) reported

that 70 to 80% of precipitation in the Naval Station area can be expected to enter the bay as

runoff due to the high degree of paved surface and low surface permeability. If the copper runoff

data reported in Gadbois (1992) are compared with data reported in the EPA nationwide urban

runoff program, it is clear that a relatively high level of copper is discharged into the bay

immediately surrounding the Naval Station during storm activity.

The mean copper concentration in storm runoff reported in the NURP study (U.S. EPA,

1983) ranged from 25 g L-1 in a 52% industrial area to 36 g L-1 in a 100% industrial area.
56

The copper concentrations measured in storm runoff from 14 separate sites in the Naval Base

reported in Gadbois (1992) ranged from 120 to 18300 g L-1 and averaged 190 g L-1. The

majority of measurements reported by Gadbois (1992) were above the state water copper

concentration objective of 2.9 g L-1 by nearly two orders of magnitude. Action is being taken

by state regulatory agencies to minimize pollutants in storm runoff. It is likely that some

reduction of copper input to the surrounding bay area will occur from implementation of best

management practices to minimize inputs. Some degree of copper input will, however, most

likely continue from storm runoff. If best management practices decrease input by an order of

magnitude, then the storm runoff input of copper would be similar to the highest copper

concentrations measured in the plume samples collected 3 meters from a ship hull. If the copper

concentration in storm water runoff from the Naval Station were decreased by two orders of

magnitude, the input would be similar to copper concentrations measured farther from ship hulls

during cleaning operations, as well as similar to concentrations measured in the bay-wide

monitoring surveys.
57

DISCUSSION

Our bay-wide water column copper measurements and hull cleaning surveys have demonstrated that copper concentrations may exceed water quality criteria throughout the bay. Copper

input due to hull cleaning operations was determined to be a short-term spatial and temporal

input. In the open bay main channel off the Naval Station, water samples averaged 3.8 g L-1.

Collectively, the average dissolved copper concentration in water samples collected from all

ships studied before all hull cleanings was 5.8 g L-1, and 6.0 g L-1 after hull cleanings at

Piers 12 and 13 within the Naval Station. While hull cleaning operations resulted in locally

elevated water column copper concentrations in some instances, clearly, during a period of

several hours, overall average concentrations returned to ambient levels.

The widespread presence of copper in the bay is not surprising when the historical presence

of pleasure, commercial, and military vessels is considered. The persistent input of copper from

antifouling paints (56 metric tons of copper applied yearly) may be considered a historical and

continuing source of copper on a bay-wide scale. Although it is difficult to calculate the actual

input from leaching processes due to a great deal of uncertainty regarding vessel size, usage, and

paint type, a great deal of copper enters the bay from antifouling paints on a daily basis.

Additionally, the input of copper from urban storm-water runoff, existing areas of polluted

sediment, and power plant operations further contributes to the copper level of bay waters. With

the exception of the input from the PACO Ore Terminal area that will undergo remedial action,

the other sources of input are expected to continue on a similar scale. We do not, therefore,

predict that copper levels in the bay are likely to decrease, and those recently measured may be

considered characteristic of present as well as future levels.

To put the input of copper from Naval hull cleaning operations into San Diego Bay in

perspective, we have composed a graphical summary of yearly inputs from several principal

sources (figure 31). Clearly, pleasure craft and the naval fleet are the major sources of copper

from antifouling paints to the surrounding bay waters. The inputs from a power plant and

commercial shipping are quite small. The Naval hull cleanings (29 events in 1993) accounted

for a comparatively small input to bay waters.

Several assumptions were made to develop these estimates. A common leach rate of 10

g/cm2/day was assumed for all small craft, naval vessels, and commercial vessels based on an

estimate of the required amount of copper leachate necessary to control fouling (Preiser et al.,

1977). Since paints may vary in their copper content and leach rate, a single estimate was used

for all vessel sources. Wetted hull areas were estimated from data reported in Grovhoug et al.,

(1989) for pleasure and commercial vessels. An average wetted hull area of 35,000 sq. ft. was

assumed for naval vessels since this would average both very large and smaller vessels. We

assumed that of the 51 naval vessels currently homeported in San Diego, 40 might be in port.

The area of cleaning influence described by a cylinder with a radius of 101 m and depth of 9 m

multiplied by the average copper concentration from the mid-depth and mid-hull transect of the

USS DURHAM cleaning was used to estimate hull cleaning input. These particular copper

values were used because they have the best and most complete regression data available. Data

reported for NPDES discharges from the SDG&E South Bay power plant were used to estimate

the yearly power plant input. We did not estimate urban copper runoff since such an estimate is

difficult due to rainfall variability in San Diego. The urban runoff input would, of course, lessen

the percentages of the other input sources.
58

Figure 31. Relative estimated copper loading from antifouling paint and other sources

into San Diego Bay per year.

With regard to current water column copper concentrations, the degree of toxicity is of

fundamental importance since concentrations are frequently above the water quality criteria.

The ion-selective electrode data reported in this study strongly suggest that activity of Cu++ (that

which may be considered most available and toxic to organisms) is an extremely small fraction

(concentrations as low as 0.06 ng L-1) of the total copper measured. The dissolved copper

fraction was elevated during cleaning activity, but quickly returned to ambient levels in a matter

of a few hours. The hull cleaning debris was largely particulate copper as evidenced by comparison of total and particulate copper concentrations. Particulate copper was frequently 50 to 80%

of the total copper measured and is not considered available to water column organisms. The

algal test IC50 estimates of 42 to 45 g L-1 copper from the hull cleaning effluent collected in

this study emphasize the likely non-availability of copper released from cleaning operations to

resident algal species, which have been shown to be resistant to copper at the Naval site as well

(Krett, 1980). If copper in the hull cleaning effluent was more available (i.e., present to a greater

degree in the dissolved phase than ISE measurements have shown), then the algal IC50 concentrations should have been much lower than those calculated in this study. The algal IC50

values determined in this study are in agreement with literature values cited in table 2 that

summarizes the widely varying toxicity of copper to marine species.

If the estimates of dissolved copper in our study are accurate, and the labile fraction measured by ASV is not toxic, then the site specific water quality criterion for the surrounding

waters of the Naval Station would be expected to be well above the general value of 2.9 g L-1.

This assumption should be tested using the EPA toxicity testing protocols discussed in the

regulatory status section of this report. If our assumptions are accurate, a positive water effect

ratio could be determined via toxicity testing using reference and site-specific waters. If this

ratio were simply a factor of two, then the resultant site-specific water quality criteria would be

2.9  2 = 5.8 g L-1 copper (fraction passing a 0.45-m filter). Such a concentration is
59

essentially the average copper water concentration measured at the hull cleaning sites before and

after hull cleaning operations.

The particulate copper released by the hull cleaning operations settles to the bottom in the

vicinity of the parent ships. Our benthic flux data measured at the Paleta Creek site near Pier 8

at the Naval Station indicated that no positive copper flux was occurring. Since this area is very

near and similar in sediment particle size characteristics to the sites at the Naval Station where

hull cleanings occurred, we would not expect a positive flux of copper from the hull cleaning

sites, given a period of equilibration where natural processes would act to bind the copper

deposited in the underlying sediments. Only one sediment sample measured at the hull cleaning

sites registered a total copper concentration above 1000 mg kg-1, which is the clean-up limit

acceptable at the PACO Ore Terminal determined by the Regional Water Quality Control Board.

Therefore, the hull cleaning operations have not resulted in unacceptable deposits of copper to

the underlying bay sediments.

During the hull cleaning process, biomass debris is released and settles to the sediment.

Previous unpublished hull cleaning data indicated that the biomass settling to the sediment was

not sufficient to cause adverse effects of burial of indigenous species. Subsequent increases in

biological oxygen demand due to release of hull debris also was not a significant problem.
60
61

CONCLUSIONS

 Copper concentrations in San Diego Bay water and sediment have apparently remained

similar over at least the last 10 years, including areas near the Naval Station.

 Initial and post-cleaning copper concentrations were above the EPA’s 2.9-ppb water

quality criteria concentration. During hull cleaning operations, initial and post-cleaning

copper concentrations were well below the 42 to 45 ppb EC50 toxicity threshold determined for algae using the Qwik-Lite chlorophyll bioassay. The data indicate that

cleaning activity did not result in copper concentrations toxic to algae. ISE measurements indicate that dissolved copper concentrations were well below values reported to

be toxic.

 Sediment benthic flux measurements indicated that copper in sediment near the hull

cleaning area at the Naval Station would not likely become available to the overlying

water column due to a negative flux measured. It may be assumed due to the proximity

and similar sediment characteristics that copper in sediments from the hull cleaning area

would also bind to the sediments.

 Only a single sediment sample collected at the hull cleaning sites exceeded the copper

level accepted as protective of marine species in the bay (1000 ppm). Only one beforeand-after comparison of copper in sediment collected from the hull cleaning sites

indicated an increase in copper from cleaning activity. Hull cleaning activities do not

appear to significantly increase toxic levels of copper in bay sediments.

 Inputs of copper to San Diego Bay are varied and will likely result in continued levels of

copper similar to those found during this study. The occasional cleaning of Navy ship

hulls did not result in increased copper concentrations beyond a very brief period of time;

nor were copper concentrations measured above levels toxic to algae. This suggests that

other marine species would not be adversely affected at the copper concentrations

measured.
62
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spermatozoa. Biol. Bull., 147, 236.

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from desalination plants: The effect of copper content, heat and salinity. Water Res. and

Develop., U.S. Department of the Interior, Washington, D.C.

Zirino, A. and S. Yamamoto. 1972. A pH-dependent model for the chemical speciation of

copper, zinc, cadmium and lead in seawater. Limnol. Oceangr., 17(5), 661–671.

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by automated anodic stripping voltametry. Environ. Sci. Technol., 12, 73–79.
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ENVIRONMENTAL EFFECTS FROM IN-WATER HULL CLEANING OF

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