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Nov 3, 1998 - *Environmental and Occupational Health Sciences Institute (EOHSI). Piscataway, NJ 08854. +Nu Horizon Enterprises, Inc. Cranford, NJ 07016.

Prepared for: The International Copper Association

By: P. G. Georgopoulos*, A. Roy*, M. J. Yonone-Lioy +, R. E. Opiekun*, and P. J. Lioy*

*Environmental and Occupational Health Sciences Institute (EOHSI) Piscataway, NJ 08854 +

Nu Horizon Enterprises, Inc. Cranford, NJ 07016

March 2001

EOHSI is a joint project of UMDNJ-Robert Wood Johnson Medical School and Rutgers The State University of New Jersey


Table of Contents


OBJECTIVE AND OVERVIEW...................................................................................... 1

1.1 Objective .........................................................................................................................1 1.1.1 What This Document Contains ...................................................................................... 5 1.1.2 What Should Come Next................................................................................................ 6 1.2

Description of the Relational Database Management System for Copper Environmental Distribution and Exposure Studies (RDMS-CEDES) ......................................................7 1.2.1 Using the RDMS-CEDES CD-ROM ............................................................................... 8

2 2.1

INTRODUCTION: COPPER AND MAN ...................................................................... 15 Copper and Technology: From the Copper Age to the Information Age .......................15

2.2 Copper and Biology: Essentiality and Toxicity ..............................................................15 2.2.1 Health Effects in Humans............................................................................................. 16 2.3 Copper and Policy.........................................................................................................19 2.3.1 Guidelines and Recommendations for Copper Intake ................................................. 20 2.3.2 Notes on Recent Regulatory Activities ......................................................................... 21 3 3.1

A BRIEF OVERVIEW OF THE PHYSICAL, CHEMICAL AND BIOLOGICAL PROPERTIES OF COPPER ........................................................................................ 27 Physical and Chemical Attributes of Copper .................................................................27

3.2 Biological Functions of Copper .....................................................................................29 3.2.1 Biochemical Functions ................................................................................................. 29 3.2.2 Physiologic Functions .................................................................................................. 32 4

ENVIRONMENTAL RELEASES OF COPPER ........................................................... 39


Atmospheric Releases ..................................................................................................39

4.2 4.2.1 4.2.2 4.2.3

Releases to Wastewater ...............................................................................................40 Copper in Cooling Systems.......................................................................................... 42 Releases to Land ......................................................................................................... 43 Databases for Environmental Copper Releases .......................................................... 43




General Concepts .........................................................................................................49


Atmospheric Dynamics of Copper.................................................................................51


Hydrospheric Dynamics of Copper ...............................................................................52

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5.3.1 5.3.2 5.3.3 5.3.4 5.3.5 5.3.6 5.3.7 5.3.8

General Discussion ...................................................................................................... 52 Fluxes of Copper in the Hydrosphere .......................................................................... 53 Chemistry of Copper in the Hydrosphere..................................................................... 57 Drainage from Urban Areas ......................................................................................... 60 Copper in Groundwater................................................................................................ 61 Mathematical Modeling of Metal Speciation................................................................. 61 Copper in Seawater ..................................................................................................... 63 Copper in Sediments.................................................................................................... 63

5.4 Copper in the Lithosphere and Pedosphere .................................................................66 5.4.1 Soil (Pedospheric) Dynamics of Copper and Soil-Water Interactions .......................... 67 5.5 5.5.1 5.5.2 5.5.3 6

Environment-Biota Interactions: Bioconcentration and Bioaccumulation ......................70 Aquatic Biota ................................................................................................................ 70 Terrestrial Biota............................................................................................................ 73 Copper in the Anthroposphere ..................................................................................... 73 COPPER LEVELS IN THE ENVIRONMENT AND HUMAN TISSUES: OVERVIEW OF SELECTED FIELD STUDIES ...................................................................................... 81

6.1 Atmospheric Concentrations .........................................................................................81 6.1.1 Precipitation ................................................................................................................. 84 6.1.2 Fog ............................................................................................................................... 84 6.2 6.2.1 6.2.2 6.2.3 6.2.4 6.2.5 6.2.6

Copper Concentrations in the Hydrosphere..................................................................84 Marine Water................................................................................................................ 85 Estuarine Water ........................................................................................................... 87 River Water .................................................................................................................. 87 Lake Water................................................................................................................... 89 Groundwater ................................................................................................................ 90 Sediments .................................................................................................................... 91

6.3 Copper Concentrations in Soils and Terrestrial Biota .................................................100 6.3.1 Copper at Hazardous Waste Sites............................................................................. 104 6.4 Copper Levels in Indicator Biota .................................................................................104 6.4.1 Sewage Sludge .......................................................................................................... 107 7

ASSESSING HUMAN EXPOSURE, DOSE AND RISK ............................................ 109


Exposure Potential and Pathways ..............................................................................109

7.2 7.2.1 7.2.2 7.2.3

Environmental Exposures ...........................................................................................109 Inhalation.................................................................................................................... 109 Ingestion..................................................................................................................... 110 Dermal........................................................................................................................ 110


Occupational Exposures .............................................................................................110

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7.4 Dietary Exposures to Copper: Drinking Water and Food ............................................112 7.4.1 Copper in Food .......................................................................................................... 112 7.4.2 Copper in Drinking Water........................................................................................... 115 7.5 7.5.1 7.5.2 7.5.3

From Exposure to Dose: Bioavailability of Dietary Copper .........................................122 Interactions of Copper with Other Components of the Diet........................................ 122 Copper Biokinetics and Metabolism........................................................................... 124 A Note on Unusually Susceptible Populations (Based on ATSDR, 1990) ................. 127


Modeling Health Risks Due to Copper Exposures ......................................................128


BIOMARKERS OF COPPER EXPOSURE AND EFFECTS ..................................... 135

8.1 Biochemical Indices of Copper Status ........................................................................135 8.1.1 Biomarkers of Exposure Susceptibility and Effect...................................................... 135 8.2

Copper Biomarkers Used in Recent Human Health Studies.......................................138

8.3 Copper Biomarkers and Exposure Markers in Populations ........................................141 8.3.1 Copper Levels in Human Hair .................................................................................... 142 8.3.2 Copper Levels in Blood and Serum ........................................................................... 143 9



APPENDIX A: A COMPILATION OF COPPER DATA IN FOOD, DRINKING WATER, AND AIR .................................................................................................................... 151






REFERENCES........................................................................................................... 173

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List of Tables Table 1: International Regulatory Practices for Copper in Drinking Water .................................................24 Table 2: Copper Containing Proteins Found in Humans.............................................................................34 Table 3: Selected Industrial Copper Compounds and their Properties.......................................................35 Table 4: Copper Releases to Water and Land, 1987 to 1993 (in pounds)..................................................44 Table 5: Summary of Copper Concentrations in Environmental Media as Reported in the Copper Sourcebook 1998 (Harrison, 1998)..............................................................................................75 Table 6: Copper Concentrations in Environmental Media and Biota as Reported in The Handbook of Trace Elements (Pais & Benton Jones, 1997).............................................................................76 Table 7: Copper Concentrations in Environmental Media and Biota as Reported in Nriagu (Nriagu, 1979b) .....................................................................................................................................................77 Table 8: Sediment Component Classes......................................................................................................78 Table 9: Tissue and Body Copper Levels in Healthy Adults and Adults with Wilson's Disease ...............145 Table 10: Copper Content of Human Tissues and Body Fluids ................................................................146 Table 11: Concentrations of Copper in Air ................................................................................................151 Table 12: Concentrations of Copper in Water ...........................................................................................153 Table 13: Copper Content of Selected Foods ...........................................................................................154 Table 14: Copper Content of Selected Foods per 100 Grams..................................................................156

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List of Figures Figure 1: Main Switchboard of RDMS-CEDES ...........................................................................................10 Figure 2: Reports Switchboard of RDMS-CEDES ......................................................................................11 Figure 3: Criteria for Selection of Articles in RDMS-CEDES.......................................................................12 Figure 4: Forms Switchboard of RDMS-CEDES.........................................................................................13 Figure 5: Sample Report from RDMS-CEDES Showing Descriptors and Data from Selected Articles......14 Figure 6: Locations of Copper Mines in the U.S. ........................................................................................45 Figure 7: Locations of Environmental Release of Copper in the United States ..........................................46 Figure 8: Environmental Releases of Copper in the United States (by County). ........................................47 Figure 9: Conceptual Multilevel-Multiscale Model of Environmental and Human Exposure Dynamics of Copper .........................................................................................................................................79 Figure 10: Prototype Implementation of a Model of Global Distribution and Fluxes of Copper in the Environment, Using the STELLA (Structured Experimental Learning Laboratory with Animation) Simulation Software .....................................................................................................................80

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Acknowledgments The authors wish to thank the International Copper Association, NYC, NY (#TPT0569A and B99) for funding this review, and in particular Dr. Scott Baker of ICA for his insights and encouragement. The authors also wish to thank Dr. Gustavo Lagos and Dr. Herbert Allen for reviewing an early version of the manuscript and providing most valuable comments and suggestions. Finally, appreciation is extended to Ms. Linda Everett for assisting with the literature database and the preparation of the final version of this manuscript, to Mr. Ioannis Georgopoulos and Dr. Vikram Vyas for providing useful information on publications and databases relevant to this effort, and to Mr. Michael Hennelly of ICA for his editorial comments.

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The objective of this work is to provide a selective review and evaluation of currently available information on environmental and biological chemodynamics (i.e. physicochemical transformations and transport dynamics) of copper and of its levels and distribution (observed or estimated) in environmental media and human tissues. An essential component of this evaluation effort has been the development and testing of an appropriate structure for the flexible organization and management of the above information. Such a structure has been implemented computationally in the form of a prototype, useroriented, relational database, named Relational Database Management System for Copper Environmental Distribution and Exposure Studies (RDMS-CEDES). This easily expandable and upgradable prototype database (available in MS Access format) provides access to data summaries, literature references and other information relevant to studies focusing on copper levels in environmental media and human tissues, and on the processes affecting these levels. This monograph accompanies and supplements the above prototype relational database, providing both (a) an overall background of issues relevant to copper release and distribution in the environment and to associated potential human exposures, and (b) overview, summaries and discussion of the specific information contained in RDMS-CEDES. The selective review attempted in this work provides a basis for developing an evolving conceptual — and eventually quantitative — model, integrating both phenomenological and mechanistic aspects of copper dynamics and environmental exposure potential. It is also expected to provide a scientific basis for the design of future research to fill identified gaps in data, specifically needed to test hypotheses that can establish source-to-human receptor relationships. The above mentioned model of copper dynamics would provide a tool for identifying and evaluating the influence of •

biogeochemical processes, and

human activities

on the distribution and accumulation of copper in different environmental media. This tool can be expected to enable establishment of a rational basis for quantifying estimates of exposure and dose to human populations, to specific individual cases of concern, and to other relevant components of ecosystems.

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It should be noted that clearly it is not the objective of this work to provide an exhaustive review of available data on copper in the environment, nor to review the biological impact and effects of copper on humans and on terrestrial and aquatic ecosystems. The Copper Research Information Flow (CRIF) Project maintains the International Copper Association (ICA) Reference Collection - the most extensive and current collection of published materials on copper as it relates to human health and the environment. CRIF is supported by the International Copper Association and based in the Department of Earth and Ocean Sciences at the University of British Columbia (UBC) in Vancouver, Canada. The Project now contains over 74,000 citations on copper, provides access to major online databases (e.g., Chemical Abstracts, Toxline, Medline, Biosis, NTIS, Embase) and 5,000 e-journals. The database is updated and augmented regularly to ensure currency. Led by Dr. Brenda Harrison at UBC, the service can be accessed by contacting Dr. Harrison at [email protected] Readers interested in an introduction to the ecological role and effects of copper can consult excellent recent reviews on these subjects (Lewis, 1995a; Lewis, 1995b; Ecological Planning and Toxicology Inc., 1998). An introduction to issues associated with copper toxicity and related risks, covering literature sources up to the late 1980s, is provided in the 1990 ATSDR Toxicological Profile for Copper (TP-90-08). Good summary reviews of the role of copper as a micronutrient in the human diet are available (see e.g. Turnlund, 1999; Bogden & Klevay, 2000; WHO-IPCS, 1998; National Institute of Medicine, 2001). Additional useful information can also be found in these references: •

Adriano, 1986, Trace Elements in the Terrestrial Environment. 533 pp. Springer-Verlag, NY. (Chapter 6. Copper. pp. 181-218.)

Howell & Gawthorne, 1987, Copper in Animals and Man. Vol. A. 125 pp. Vol B. 140 pp. CRC Press.

Kies, 1989, Copper Bioavailability and Metabolism. Vol. 258 in Adv. Experimental Medicine and Biology. 307 pp. Plenum Press, NY.

Linder, 1991a, Biochemistry of Copper. 525 pp. Plenum Press, NY.

Linder, 1991b, Nutrition and metabolism of the trace elements. in Nutritional Biochemistry and Metabolism: With Clinical Applications. Elsevier Science Publishing, Amsterdam, The Netherlands.

Finally, since 1997, a number of new documents have been published which contain useful review/summary information on copper: •

He et al., 1997, Spatial and temporal patterns of acidity and heavy metals in predicting the potential for ecological impact on the Le An river polluted by acid mine drainage. Sci. Total Environ.

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Kennish, 1997, Practical Handbook of Estuarine and Marine Pollution. 524 pp. CRC Press, Boca Raton.

O'Dell & Sunde, 1997, Handbook of Nutritionally Essential Mineral Elements. 692 pp. Marcel Dekker, NY. (Chapter 8, Copper by E.D. Harris. pp. 231—274.)

Pais & Benton Jones, 1997, The Handbook of Trace Elements. 223 pp. St. Lucie Press.

Richardson, 1997, Handbook of Copper Compound and Applications. 432 pp. Marcel Dekker, NY.

Riveros-Rosas et al., 1997, Personal exposure to elements in Mexico City air. Sci. Total Environ. 198:79-96.

Tobias et al., 1997, Establishment of the background levels of some trace elements in soils of NE Spain with probability plots. Sci. Total Environ. 206:255-265.

WHO-IPCS, 1998, Copper - Environmental Health Criteria 200. World Health Organization and the International Programme on Chemical Safety

Budd et al., 1999, The Keweenaw current and ice rafting: Use of satellite imagery to investigate copper-rich particle dispersal. J. Great Lakes Research 25 (4):642-662.

Joseph, 1999, Copper: Its Trade, Manufacture, Use, and Environmental Status. 451 pp. ASTM International.

Kerfoot & Nriagu, 1999, Copper mining, copper cycling and mercury in the Lake Superior ecosystem: An introduction. J. Great Lakes Research 25 (4):594-598.

Kerfoot et al., 1999, Anthropogenic copper inventories and mercury profiles from Lake Superior: Evidence for mining impacts. J. Great Lakes Research 25 (4):663-682.

Kerfoot & Robbins, 1999, Nearshore regions of Lake Superior: Multi-element signatures of mining discharges and a test of Pb-210 deposition under conditions of variable sediment mass flux. J. Great Lakes Research 25 (4):697-720.

Kolak et al., 1999, Nearshore versus offshore copper loading in Lake Superior sediments: Implications for transport and cycling. J. Great Lakes Research 25 (4):611-624.

Landner & Lindestrom, 1999, Copper in Society and in the Environment: An Account of the Facts on Fluxes, Amounts and Effects of Copper in Sweden. Second revised edition. 329 pp. Swedish Environmental Research Group.

Leone & Mercer, 1999, Copper Transport and its Disorders: Molecular and Cellular Aspects. Vol. 448. New York, NY: Kluwer Academic/Plenum Publishers.

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Mansilla-Rivera & Nriagu, 1999, Copper chemistry in freshwater ecosystems: an overview. J. Great Lakes Research 25 (4):599-610.

DiToro et al., 2000, The Biotic Ligand Model, Copper in the Environment and Health. New York, NY: International Copper Association.

Eisler, 2000, Handbook of Chemical Risk Assessment: Health Hazards to Humans, Plants, and Animals. Vol. 1. Metals. 738 pp. Lewis Publishers. Boca Raton. (Chapter 3: Copper. pp. 93-200.)

Jeong et al., 2000, Release of Copper from Mine Tailings on the Keweenaw Peninsula. J. Great Lakes Research 25 (4):721-734.

Landner et al., 2000, Copper in Sewage Sludge and Soil, Copper in the Environment and Health. New York, NY: International Copper Association.

Multhaup & Hermann, 2000, Copper in the Pathogenesis of Neurodegenerative Disorders: A Literarature Summary, Copper in the Environment and Health. New York, NY: International Copper Association.

National Research Council, 2000, Copper in Drinking Water. 147 pp. National Academy Press, Washington, D.C.

Samet, 2000, A Technical Guide for the Study of Acute Gastrointestinal Effects of Copper in Drinking Water: Methods for Public Health Investigators, Copper in the Environment and Health. New York, NY: International Copper Association.

Georgopoulos et al., 2001, Environmental copper: Its dynamics and human exposure issues. Journal of Toxicology and Environmental Health Part B, 4:341-394.

Lagos, 2001, Corrosion of Copper Plumbing Tubes and the Release of Copper ByProducts to Drinking Water - A Literature Summary, Copper in the Environment and Health. New York, NY: International Copper Association.

Parametrix Inc. & EPT, 2001, Acclimation and Adaptation of Terrestrial Organisms to Metals in Soil. New York, NY: International Copper Association.

National Institute of Medicine, 2001, Dietary Reference Intakes for Vitamin A, Vitamin K, Arsenic, Boron, Chromium, Copper, Iodine, Iron, Manganese, Molybdenum, Nickel, Silicon, Vanadium, and Zinc (Chapter 7 - Copper). Washington D.C.: Institute of Medicine, Food and Nutrition Board.

The special issue of American Journal of Clinical Nutrition (AJCN, 1998) should also be mentioned here. It contains the scientific reports presented and discussed at the International Conference on Genetic and Environmental Determinants of Copper Metabolism, sponsored by

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the National Institutes of Health (NIH) and the University of Chile, held at the NIH Stone House, March 18-20, 1996. The purpose of the conference was to critically review genetic and environmental factors that determine copper metabolism, as well as their implications for copper deficiency and excess in humans. Finally, it should be noted that up-to-date information regarding many issues and activities involving copper (including production, economic, environmental, etc. data) can be accessed through the web pages, and the numerous links thereof, of the Copper Development Association ( ), the International Copper Association ( and the International Copper Study Group ( Also, more recent information on evolving methods and databases for assessing copper exposures can be found in the forthcoming report “A Framework and Data Sources for the Assessment of Exposures to Copper” (Georgopoulos et al., 2002). More details are provided on the website of the Center for Exposure and Risk Modeling (CERM), in the section devoted to copper ( 1.1.1

What This Document Contains

In order to provide a conceptual “road map” to the organization of this monograph, we include a brief description of the contents of each chapter. •

The current chapter sets forth the objectives and organization of this monograph and of the accompanying relational database management system for copper environmental distribution and exposure studies (RDMS-CEDES). In addition, it provides up-to-date information sources unavailable when this document was written, and recommendations for what needs to be done next.

Chapter 2 discusses the history of copper use and summarizes its human health effects, in terms of essentiality, deficiency and excess (toxicity). It also discusses worldwide governmental copper regulation/policy/guidelines for pollutant discharge, human exposure and dietary intake.

Chapter 3 provides a brief review of copper’s essential physical, chemical and biological attributes, as a background for the subsequent discussion of environmental and biological copper chemodynamics.

Chapter 4 concerns copper production, use and disposal.

Chapter 5 provides a brief overview of environmental and biological copper chemodynamics, including the atmospheric, hydrospheric, and soil-water interactions, as well as environment-biota interactions.

Chapter 6 provides an overview of selected field studies on copper levels in the environment, from household dust to marine sediments, and in human tissue.

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Chapter 7 discusses human exposure, dose, and risk assessment under various conditions and by various routes, and copper health-risk modeling.

Chapter 8 concerns biomarkers of copper exposure and effects, and

Chapter 9 sets forth conclusions and recommendations for future research.


What Should Come Next

During the development of this report it was realized that a very significant component of existing data on copper have not been incorporated in the peer-reviewed literature. There are several reasons for this, the main one being that important information on copper is often collected through studies not focusing on copper. As a result, data directly related to copper often do not become the subject of specific analysis and/or do not get published in the peer-reviewed literature; rather, they can only be found in federal and state agency reports and/or in electronic data files. To address this problem, a number of databases containing temporal and spatial information on copper distribution were identified, mostly from such United States federal agencies such as USGS, the Centers for Disease Control, and US EPA. Such databases are: • • • • • • • • • • • •

USGS Water Quality Monitoring Network (WQN) MT2 Data USGS National Geochemical Atlas Data US NOAA Ocean Resources Conservation and Assessment (ORCA) Data Trends and Status, Mussel Watch and Benthic Surveillance US EPA Environmental Monitoring Assessment Program (EMAP) US EPA Toxics Release Inventory US EPA NHEXAS US EPA AIRS US EPA SDWIS US EPA STORET CDC NHANES II/III Geochemical Atlases of Hungary, Slovakia, Sweden, etc. SWAD (European Surface Water Database)

In most instances, analysis of the data in the above databases has not yet appeared in the peer reviewed literature. Appendix B of this document provides a summary description of the contents of these databases. It is our strong recommendation that an effort be made to retrieve and organize material from these databases, with an appropriate focus on copper.

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1.2 Description of the Relational Database Management System Environmental Distribution and Exposure Studies (RDMS-CEDES)



The prototype RDMS-CEDES was developed using, as its starting source of information, the Copper Information Sourcebook 1998. The Sourcebook lists 4,746 articles (i.e. the results of the 1997 search selected for inclusion in the ICA Reference Collection, which, as mentioned earlier, contains more than 50,000 items) spanning a wide array of scientific disciplines, encompassing human health, ecology, agriculture, the environment, etc. In addition, the Sourcebook also contains tabular compilations of the gross aspects of the copper monitoring data (e.g., min-max ranges, etc.) found in the articles cited. Owing to the wide variety of studies cited in the Sourcebook, not all were of equal value for the purposes of this monograph. Further, there is considerable variability in the types and quality of reported copper concentration data. As a consequence of such informational limitations and constraints, it was decided to concentrate on copper-related data reported within the context of the design of individual studies. The database provided on the CD-ROM attempts to offer a perspective on the content of a subset of the literature cited in the Sourcebook that met the criteria (provided in the final draft of the CDROM) of focus, relevance, and study design, needed for inclusion in the present assessment. More specifically, the articles listed in the 1998 Sourcebook were first screened based upon their titles, and over 400 articles, or about 8%, were identified as potentially containing information on environmental copper that met the selection criteria of focus and relevance. Thus, the selected articles are a subset of those used to compile the tables of environmental monitoring data in the 1998 Sourcebook. These 400 plus articles were then reviewed, and evaluated, and 112 of them were selected based on the following criteria: • • •

adequate definition of study design; full characterization of the analytical methods used in the study; inclusion of adequate summaries; and

explicit discussion of the conclusions derived from the data, especially those which enhance understanding of spatial and temporal trends in environmental copper distribution. Salient information from these articles was summarized in RDMS-CEDES. It should be noted that of the limited number of studies that met the evaluation and analysis criteria, only six were designed specifically to study copper-related problems or issues. It should be further noted that a number of the reviewed articles were found to be relevant to more than one media, therefore supporting the approach of using a relational database as a convenient tool for cross-referencing information in these articles. The relational data management approach also provides flexibility in structure that can prove to be advantageous for future alternative classifications of the data. The articles selected for inclusion in the prototype RDMS-CEDES were characterized according to: • • •

Focus, Type of Study, Geographical Location, COPPER: Environmental Dynamics and Human Exposure Issues Page 7


• • • • •

Setting, Polluted or Pristine Environment, Environmental Media Studied, whether Flux was considered, Temporal Scale of the Study and

Spatial Scale of the Study. The ‘Focus’ field indicates whether the focus of the paper was mainly: • • • •

copper (CU), inorganic ions (II), toxic heavy metals (Boon, 1994), organics (O), or

essential trace metals (TM). The type of study is also denoted in the database by classifying research as: • • • • •

analytical (A), statistical (S), modeling (M), field research (F), laboratory-based research (L), or

toxicological study (T). Location was also used as a category in the database, classifying each study by continent. The continent identifiers are defined in the “Copper Key.txt” file located on the CD-ROM. In addition to geographical information, the database further subdivides each study by • • •

setting (e.g., agricultural, rural, urban, etc.), condition of the environment (e.g., polluted or pristine), type of media investigated (e.g., air, plants, water, etc.) and

whether or not environmental flux was considered in the study. The Spatial Scale and Temporal Scale fields in the database provide an indication of the scope of the study, and the resolution of data in space and time. All of these categories are fully defined in the “Copper Key.txt” file on the CD-ROM. 1.2.1


The RDMS-CEDS CD-ROM requires the user to have Microsoft Access (97 or 2000) software. To activate the database, the user should double-click on the file “copper_select_frontend.” Once

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the database is loaded, the user is presented with a Main Switchboard page (see Figure 1) through which the database can be accessed and sorted by the criteria mentioned above (see Figure 3). The ‘Create/Print’ option allows the user to create reports that are formatted for printing, while the ‘View Copper Data’ option allows the user to view copper references and data interactively. •

Selecting the ‘Create/Print Reports’ option allows the user to o view and print all of the references selected from the 1998 Copper Information Sourcebook for this study, and o search the 112 selected references using criteria available through pull-down menus on a form. If the user allows an asterisk to remain in place of any given criterion, all articles for that criterion will be displayed. Selecting the ‘View Copper Data’ option allows the user to view selected references based on the Focus and Environmental Media attributes of the reference. Double-clicking any of the selected references allows the user to view a summary of the data in the reference. Selecting ‘EXIT’ will allow the user to return to the Main Switchboard.

Please note: All variables are defined in the Copper Key.txt file on the CD-ROM. The following figures (Figure 1 to Figure 5) show screenshots of the “switchboards” and sample forms and reports of the prototype RDMS-CEDES.

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Figure 1: Main Switchboard of RDMS-CEDES This screen appears when the database is initially opened. Select “Create/Print Reports” to create a printable list of references. Select “View Copper Data” to view the data in one or more of the selected papers reviewed in this report.

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Figure 2: Reports Switchboard of RDMS-CEDES This screen appears when “Create/Print Reports” is selected in the Main Switchboard. Select “Search Selected Articles” to create a list of references of the selected papers reviewed in this study. Select “References: Copper Information Sourceboook 1997” to create a list of all references in the 1997 Sourcebook.

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Figure 3: Criteria for Selection of Articles in RDMS-CEDES This screen appears when “Search Selected Articles” is selected in the Report Switchboard Screen. Create a list of References of the articles reviewed in this report based on criteria chosen from the drop-down lists. Choosing “*” in the drop-down list will select all articles for those criteria.

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Figure 4: Forms Switchboard of RDMS-CEDES This screen appears when “View Copper Data” is selected in the Main Switchboard. Select “View Selected Articles and Data” to choose articles for which data are to be viewed.

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Figure 5: Sample Report from RDMS-CEDES Showing Descriptors and Data from Selected Articles This screen appears when “View Selected Articles and Data” is selected in the Forms Switchboard page. Double-click any reference in the top list of references to view the environmental and human data in the reference.

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2 INTRODUCTION: COPPER AND MAN 2.1 Copper and Technology: From the Copper Age to the Information Age

Copper, along with gold, is one of the first metals utilized by humans, about 10,000 years ago, owing to the natural occurrence of its elemental form — readily available lumps or leaves in exposed rock formations (“native copper”).* In fact, the fashioning of simple tools by hammering and heating native copper signaled the end of the Stone Age. The series of steps required for early humans to proceed from a knowledge of the requirements for melting metal to the intentional practice of ore reduction by heat in the presence of carbon is still open to speculation, as is the fabrication of copper alloys, believed to have occurred around 3000 to 2000 B.C. The discovery of coal or charcoal firing rather than wood, which was necessary to obtain sufficiently high temperatures to melt the native metal, eventually led to the mining of copper ores. From early workings discovered in the Sinai, mining is known to have begun in about 3800 B.C. Mines operating in Cyprus around 3000 B.C. were later taken over by the Roman Empire, and the metal product was called cyprium, later simplified to cuprum, the origin of the Latin name still used for the metal. It is a fact that for many centuries human technological advancement basically reflected progress in the processing and utilization of copper. Subsequently, iron and its alloys dominated metal utilization though copper usage remained extremely important; indeed copper and its alloys have always been used extensively for plumbing infrastructure, cooking utensils and artwork. It should be mentioned that recognition of the biological impact of copper and its compounds (see Section 2.2) took place quite early in history; the ancient Egyptians used copper salts as biocides. Because of copper’s excellent electrical conductivity, the advent of electricity and of telecommunications expanded its range of applications. Today’s information age has given copper new status as a “high tech” metal thanks to its potential for use in microelectronics applications, thus replacing aluminum in the manufacturing of upcoming generations of computer microprocessors. (It is interesting to note that the code name of the successor of the ubiquitous Pentium processor in personal computers has been code-named the “coppermine” processor, to emphasize its high-tech potential. In fact the name reflects marketing considerations since the particular processor is not copper-based at this point.)

2.2 Copper and Biology: Essentiality and Toxicity

Copper is an essential micronutrient for plants and animals, including humans; it is involved in the function of several enzymes and other proteins needed in a wide range of metabolic processes. At the same time, high levels of copper can be detrimental to life, thus providing a means for controlling unwanted organisms. The biological benefits of exposure to copper reflect this balance between essentiality and toxicity. *

This accumulation of elemental copper via normal geologic processes is due to its low standard reduction potential of +0.158 V, significantly below hydrogen in the electromotive series.

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The recognition of the impact of copper on biological processes took place quite early in human history. Indeed, it is believed that both the ancient Egyptians and the Chinese were using copper salts for therapeutic purposes. More detailed information on such practices is available from around 400 B.C., when Hippocrates prescribed copper compounds for pulmonary and other diseases. The use of copper compounds to treat disease peaked in the 19th century and then declined when the treatments were unsuccessful. 2.2.1

Health Effects in Humans

It is beyond the scope of this monograph to review the rapidly expanding knowledgebase on the biological significance of copper, and the variety of human health effects associated with either deficient or excessive copper intake. Instead, relevant facts pertaining to the historical evolution and the current status of the understanding of copper toxicology and health effects, based partially on Turnland’s study (Turnlund, 1999), are presented to provide perspective for the nonspecialist. Table 3 provides a representative list of copper compounds with summary information on their toxicological characterization (typically, relative to occupational settings). The reader interested in further details can consult the extensive specialized literature documented in the 1998 ICA Sourcebook; a brief introduction to the health effects associated with copper essentiality and toxicity can be found in the 1990 ATSDR Toxicological Profile for Copper; for a more up-to-date summary one can consult a more recent review, such as Turnland’s study (Turnlund, 1999), the Reports of ICA Project No. 223, and the WHO-IPCS 1998 Environmental Health Criteria 200 document on copper (WHO-IPCS, 1998). It should be mentioned that, in general, deficiency is considered to be a greater concern than toxicity. • •

• •

Late 19th century: Copper was identified as a normal constituent of blood, and its toxicity was described; By 1900, an anemia that could not be prevented by iron supplements had been observed in animals kept on a whole milk diet. In 1928 it was reported that this anemia in rats was controlled by iron only when copper supplements were also given. Experiments in several animal species produced similar results suggesting that copper-deficiency anemia occurs in all species. Human disease was first linked to copper metabolism shortly after Wilson's disease was described in 1919, long before 1953 when the condition was recognized as a genetic error of metabolism. As early as 1930, a relationship between anemia in humans and copper deficiency was suspected, but because copper supplements improved hemoglobin synthesis in only some cases, the hypothesis was not well accepted. Conclusive evidence of copper deficiency in humans was not substantiated until 1964. Menkes' disease, another anemia disorder, was described in 1962 and was recognized as a copper absorption disorder in 1972. Since about 1950, an increasing number of diseases, not specifically disorders of copper metabolism, have been associated with altered, usually increased, levels of copper in blood or other tissues. Numerous studies have demonstrated that copper is required for o infant growth,

COPPER: Environmental Dynamics and Human Exposure Issues Page 16


• •

o host defense mechanisms, o bone strength, o red and white blood cell maturation, o iron transport, o cholesterol and glucose metabolism, o myocardial contractility, and o brain development An official dietary copper recommendation for “an estimated safe and adequate daily dietary intake” was first introduced in 1979 and modified in 1989. Toxicity resulting from excess copper intake has been observed in numerous studies in a variety of animal species, such as sheep, cattle, pigs, rats, and poultry, as well as in humans.

Copper Deficiency

Copper deficiency can result in the expression of an inherited defect such as Menkes disease or in an acquired condition. Acquired deficiency is a clinical syndrome that occurs mainly in infants, although it has also been described in both children and adults. This deficiency can be the consequence of decreased copper stores at birth, inadequate copper supply, inadequate copper absorption, increased requirements, and increased losses (Amsden et al., 1978). Clinically evident copper deficiency is a relatively infrequent condition in humans. The most frequent clinical manifestations of acquired copper deficiency are anemia, neutropenia (compromised immune response), and bone abnormalities that include osteoporosis and fractures (Shaw, 1992; Olivares & Uauy, 1996a; Olivares & Uauy, 1996b).

Copper Excess: Acute Toxicity

Acute copper toxicity is infrequent in humans, and usually is a consequence of • • • • •

accidental consumption by children, ingestion of several grams in suicide attempts, application of copper salts to burned skin, drinking water from contaminated water supplies, or consumption of acidic food or beverages that were stored in copper containers.

Acute symptoms include (Knobeloch et al., 1994; Turnlund, 1999; Araya et al., 2001) • • • • •

salivation, epigastric pain, nausea, vomiting, and diarrhea.

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Vomiting and diarrhea usually prevent more serious manifestations of copper toxicity that can include • • • • • • • •

coma, shock, oliguria (dimished urine secretion), hemolytic anemia, acute renal (kidney) failure with tubular damage, hepatic necrosis (liver cell death), vascular collapse, and death.

It should be mentioned that the threshold for acute adverse gastrointestinal effects from copper in drinking water has not been precisely established (see also WHO-IPCS, 1998 for additional references on the matter).

Copper Excess: Chronic Toxicity

Chronic toxicity in humans is associated principally with Wilson disease, with the occurrence of infantile cirrhosis in areas of India (Indian childhood cirrhosis), as well as with isolated clusters of cases in other countries related to excessive copper intake (Anderson et al., 1992; Patel & Bhattacharya, 1995; Horslen et al., 1994). The most likely explanation for chronic toxicosis appears to be a genetically determined defect in copper metabolism combined with a high copper intake (Anderson et al., 1992; Alderdice & McLean, 1982; Aldini et al., 1987; Olivares & Uauy, 1996a; Miller et al., 1995, Müller et al., 1996). Chronic copper toxicosis has also been observed in dialysis patients following months of hemodialysis when copper tubing was used, and in vineyard workers using copper compounds as pesticides. The amount of oral copper intake required to produce toxic effects is not well established, but liver damage in infants has been reported as potentially related to consuming water with 2 to 3 mg copper/l in early infancy (Muller-Hocker et al., 1988). An extremely wide range of oral copper dosages, beginning at 0.07 mg/kg per day, has been associated with gastrointestinal effects (Turnlund, 1999). In addition to Wilson's disease, certain other diseases are associated with accumulation of toxic levels of copper in the liver and other tissues, even without excessive intake. Subtle deleterious effects of high dietary copper have also been observed: LDL cholesterol increased when copper supplements were given to men. On the other hand, however, lower copper intake has been implicated with other variables such as heightened cholesterol in some studies as a possible risk factor for cardiovascular disease. Copper plays a critical role in neurologic diseases; there is speculation that copper-induced production of hydroxy radicals may contribute to neurodegeneration in Alzheimer's disease (Turnlund, 1999).

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2.3 Copper and Policy

The majority of human exposures to copper are associated with the oral uptake pathway (ingestion of drinking water and food). However, in special situations, particularly in occupational settings, adverse exposures can occur via inhalation and dermal contact. Around the world, policies and guidelines for copper exposure and intake attempt to balance essentiality and toxicity. In the United States, copper and its compounds released by human activities and/or present in various environmental media are regulated according to the following environment and health protection acts: •

• • •

CERCLA: Under the Comprehensive Environmental Response, Compensation, and Liability Act of 1980 (CERCLA, United States Public Law 96-510), as amended by the Superfund Amendments and Reauthorization Act (SARA, United States Public Law 99499) releases of listed substances at or above their Reportable Quantities (RQs) must be reported to the National Response Center. RQs are set on the basis of aquatic toxicity, acute mammalian toxicity, ignitability, reactivity, chronic toxicity, and carcinogenicity, with possible adjustment on the basis of biodegradation, hydrolysis, and photolysis. The list of CERCLA hazardous substances and their RQs can be found in 40 CFR 302.4. (Further information is available from the RCRA/Superfund Hotline: 1-800-424-9346). RCRA: The Resource Conservation and Recovery Act (RCRA), administered by EPA's Office of Solid Waste (OSW), addresses the issue of how to safely manage and dispose of the huge volumes of municipal and industrial waste generated nationwide. FIFRA: The Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA) creates a statutory framework under which EPA, through a registration process, regulates the development, sale, distribution, and use of pesticides. NPDWR: The National Primary Drinking Water Regulations (NPDWR) under the Safe Drinking Water Act, Subparts B and G (codified in 40 CFR Part 141) list Maximum Contaminant Levels (MCLs) for certain chemicals. The MCL is the maximum permissible level of a contaminant in public drinking water systems. MCLs are based on health factors, but are also required by law to reflect the technological and economic feasibility of removing the contaminant from the water supply. (Further information is available from the Safe Drinking Water Hotline: 1-800-424-4791).

Additionally, copper is also regulated by the Clean Water Act, through inclusion in the Priority Pollutants List (PPL): •

CWA: The Clean Water Act regulates the discharge of pollutants into waterways by industrial, municipal, and other sources. These sources are subject to effluent limitations based on guidelines and water quality standards. The Priority Pollutants List (PPL) consists of approximately 125 pollutants. EPA has developed water quality criteria for all of them.

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Table 1 provides an overview of existing international regulations and guidelines for controlling the levels of copper present in drinking water. 2.3.1

Guidelines and Recommendations for Copper Intake

Frank copper deficiency in humans is very rare, which suggests that the current dietary intake usually suffices to prevent copper deficiency. It has been suggested that the usual copper intake is marginal and may not support optimal health, but data in this area are conflicting and not sufficient to support that hypothesis. Due to the difficulty in measuring copper status via factors such as zinc, carbohydrate and vitamin C intake, that affect copper bioavailability (see following chapters), it is very difficult to establish exact requirements for copper intake. For this reason, the Subcommittee on the Tenth Edition of the RDAs could not establish an RDA for copper and instead recommended a safe and adequate range of copper intake. While copper is viewed as a nutrient that is under effective homeostatic control in humans, adequate dietary supplementation with copper in relation to total parenteral nutrition, and in relation to nutrient-nutrient, hormonenutrient and nutrient-pharmaceutical interactions remain important areas for study. Copper depletion/repletion studies to establish the minimum requirements for healthy humans were done only recently, with a copper intake low enough to produce systematic reduction in copper status. Relatively few cases of frank copper deficiency have been reported, and these were accompanied by confounding factors, such as malnutrition, malabsorption, and excessive gastrointestinal losses. Hence, their value in establishing a minimum requirement for healthy individuals is limited. The most relevant example of a long-term diet containing less than the minimum copper requirement may be the following (Higuchi et al., 1988) as reported in Turnlund (1999): An enteral diet containing 15 µg copper/100 kcal (0.56 pmol/J) produced copper deficiency in six of six severely handicapped patients between the ages of 4 and 24 years after they had consumed the diet for 12 to 66 months. Serum copper values of 1.8 to 7.2 µmol/L (11.7-45.7 µg/dL) and ceruloplasmin values of 30 to 125 mg/l (3-12.5 mg/dL) were discovered, accompanied by other manifestations of copper deficiency. These values increased to within the normal range after 3 months of copper supplementation. By extrapolation (though it may not be valid to extrapolate to healthy adults, from these growing, severely handicapped individuals) copper deficiency could be expected to develop eventually, if the diet contained 15 µg copper/100 kcal, or 0.44 mg copper/2900 kcal for men and 0.29 mg copper/1900 kcal for women (0.56 pmol/J). This example, combined with one study in which healthy young men maintained copper balance and status at 0.79 mg/day (12 µmol/day), and another in which young men did not maintain status at 0.37 mg/day (6 µmol/day) (Turnlund, 1999), suggests that the minimum copper requirement for men is somewhere between 0.4 and 0.8 mg/day (6-12 µmol/day). An adult basal copper requirement within that range, 0.6 mg/day (9 µmol/day) for women and 0.7 mg/day (11 µmol/day) for men, was suggested by WHO (World Health Organization, 1996). In a recent report, the National Institute of Medicine (National Institute of Medicine, 2001) has set a new RDA for copper at a minimum of 0.9 mg/day for adults from food alone and an upper limit of 10 mg/day of copper sourced from a combination of food, drinking water and supplements.

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Most multivitamin dietary supplements on the market today include 2 mg of copper, the midpoint of the Safe and Adequate Range of Intake recommended by the FNB.

NRC Recommendations

The daily dietary copper intake recommended by the U.S. National Research Council (NRC) has been 1.5 to 3 mg (24-47 µmol) for adults, and 0.4 to 0.6 mg (6-9 µmol) for infants 0 to 6 months of age, to 1.5 to 2.5 mg (24-39 µmol) for children over 11 years of age (Aalbers et al., 1987). WHO Recommendations

WHO (World Health Organization, 1996) estimated the normative requirement for women to be 0.7 mg/day (11 µmol/day) and for men, 0.8 mg/day (12.5 µmol/day). After adding margins of safety to the individual requirement, including all dietary conditions, variations in usual intakes, and individual variability, a recommendation of 1.25 mg/day (19 µmol/day) was derived. 2.3.2

Notes on Recent Regulatory Activities

European Union

In 1998, Council Directive 98/83/EC on the quality of water intended for human consumption (adopted by the Council on 3 November 1998) established a new drinking water standard of 2 mg/l for copper (from 3 mg/l). The directive entered into force on 25 December 1998; member states were given two years to transpose the directive into national legislation and five years to ensure that drinking water complies with the standard set.

United States

In 1998, the California Environmental Protection Agency (EPA) was asked to develop 75 Public Health Goals (PHGs) over the following three years. The California EPA Office of Environmental Health Hazard Assessment has already published the PHG for copper in drinking water as 0.17 mg/l, even while acknowledging the scarcity of good data and the scientific weakness of the study on which this value is based. This PHG is substantially lower than the federal drinking water maximum contaminant level goal (MCLG) of 1.3 mg/l, which is itself 35% lower than the World Health Organization provisional drinking water guidance level of 2 mg/l. The California PHG, which is advisory, became effective in 1999 and the industry is concerned about the potential for a derivative regulatory state action level. Further, under California law, consumer notification will be required under the U.S. Lead and Copper Rule if PHG exceedences are found during residential tap water monitoring. Simultaneously, the California Department of Housing and Community Development (CalDHCD) is considering whether to continue the ban on the use of cross-linked polyvinyl chloride (CVPC) plumbing tube. To make this determination, CalDHCD is preparing an Environmental Impact Report (EIR)

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that will compare the economics, technology and risks of copper and plastics. The PHG, as well as existing and future water quality, will be given careful consideration in the EIR analysis. In December 1999, the U.S. EPA made minor revisions to the Lead and Copper Rule, effective April 11, 2000. These revisions, the Lead and Copper Rule Minor Revisions or LCRMR, streamline requirements, promote consistent national implementation, and in many cases, reduce burden for water systems. The LCRMR do not change the action levels of 0.015 mg/l for lead and 1.3 mg/l for copper, or Maximum Contaminant Level Goals of 0 mg/l for lead and 1.3 mg/l for copper, established by the 1991 Lead and Copper Rule (“the rule”). They also do not affect the Rule's basic requirements to optimize corrosion control and, if appropriate, to treat source water, deliver public education, and replace lead service lines. As part of the LCRMR rulemaking process, the Agency collected additional data pertaining to exclusion of transient non-community water systems from the Rule’s requirements. EPA concluded that it is still appropriate to continue this exclusion because the Agency believes there are de minimus (Andersson et al., 1991) non-carcinogenic adverse health effects resulting from exposure to lead in drinking water at such systems. All water system operators and managers of community water systems (CWSs) and non-transient non-community water systems (NTNCWSs) are potentially affected by the LCRMR, as are state staff responsible for implementing the Lead and Copper Rule in their state. (More information is provided by EPA's Safe Drinking Water Hotline, 1-800426-4791, and by the Office of Ground Water and Drinking Water web page at In 1997, in accordance with the Clean Air Act, the U.S. EPA signed a new national ambient air quality standard for particulate matter and ozone, which provides increased protection against a wide range of particulate-matter-related health hazards and could have a significant impact on the mining and mineral-processing industries. The existing standard regulates particles that are 10 microns in size or smaller and sets an average annual concentration limit of 50 micrograms per cubic meter and an average daily limit of 150 micrograms per cubic meter. Expected to take effect between 2000 and 2002 (Platt’s Metals Week, 1997e), the new standard for particles 2.5 microns or smaller is an annual mean of 15 micrograms per cubic meter and a 24-hour mean of 65 micrograms per cubic meter. In 1989, the Basel Convention on the Control of Transboundary Movements of Hazardous Wastes and their Disposal came into force and has since been ratified by more than 100 countries, including the United States. However, the U.S. has not passed legislation necessary to implement the Convention. An international Technical Working Group met four times between September 1995 and February 1997 to consider which materials should be classified as hazardous and, hence, affected by the various bans. The group elected to include copper scrap, copper slags, and copper oxide mill scale in the B list, the list of materials not covered by the Basel Convention as hazardous and, thus, not subject to any export ban. In February 1998, the Basel Convention held its Fourth Meeting of the Council of Parties in Kuching, Malaysia, and adopted the A list, wastes characterized as hazardous and therefore subject to regulation, and the B list, as developed by the Technical Working Group. However, material contained in the B list is not precluded from regulation, if it contains any of a core list of materials to be controlled

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because they exhibit hazardous characteristics. (More information can be found on the web site of the International Copper Study Group (ICSG) at

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Table 1: International Regulatory Practices for Copper in Drinking Water*

Country EU

Regulates copper in Drinking Water Yes

Max. copper Levels 2 mg/l

Guiding factors (Aesthetic, Mandatory Health) or Guideline M Aesthetic (smell, taste, color)



1 mg/l


Aesthetic (stain)



2 mg/l




1 mg/l


Health (WHO Guidelines) Aesthetic

EU 1998 Directive Historical levels



2 mg/l





2 mg/l




1 mg/l




Basis for Limits 1998 Cu directive 98/83/EC is currently in effect. NA

Sampling Basis NA

Planned changes to regulations Yes

New Regulations to be Introduced No

Random. Regional responsibility






EU 1998 Directive




EU 1998 Directive





Historical levels



2 mg/l


Aesthetic (taste and stain)

EU 1998 Directive

At tap after 12 hours standing



2 mg/l


EU 1998 Directive





0.2 mg/l at tap after flushing; 2 mg/l no flush


Aesthetic (taste, corrosion, stain) Aesthetic (corrosion, taste, stain and green hair) Health

Sampling sites: Water works and consumers tap. At consumers' tap, without and after flushing.


United Kingdom


2 mg/l



Tap samples




1 mg/l



Scientific research epidemiological studies, by experience, EU drinking water regulations, WHO guidelines WHO Guidelines, EU 1998 Directive NA




1 mg/l


Aesthetic (taste, stain)

Reviewing procedures Provincial responsibility

Based on new EU Drinking Water Directive† No



Staining @>1mg/l Recognize essentiality of copper.


Based on new EU Drinking Water Directive † No Based on new EU Drinking Water Directive † Based on new EU Drinking Water Directive † Based on new EU Drinking Water Directive † Based on new EU Drinking Water Directive † Based on new EU Drinking Water Directive† Based on new EU Drinking Water Directive†

This table was adapted from information provided at and should not be considered complete. It contains as much reliable information as could be gathered at the time that this report was written. † Under normal circumstances EU member states are expected to amend their national regulations within two years of formal promulgation of a new EU Directive.

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Regulates copper in Drinking Water Yes

Max. copper Levels 1 mg/l

Guiding factors (Aesthetic, Mandatory Health) or Guideline M Aesthetic

China Czech Republic

Yes Yes

1 mg/l 0.1 mg/l


NA Health



1.0 mg/l




1.5 mg/l


Aesthetic (smell, taste, stain) NA



1.0 mg/l


Aesthetic (stain)



1.0 mg/l


Aesthetic (stain, taste, corrosion) Health (Liver damage from long-term accumulation) Recom. Dietary Reqt: 20-80 µg/kg body weight/day

Mexico Australia


NA 2 mg/l


New Zealand


NA Aesthetic (1 mg/l) Aesthetic (staining, 1mg/l) Health parameters (projectile vomiting, 2mg/l)

Norway Peru

Yes Yes

Country Chile


0.3mg/l 1 mg/l



Planned changes to regulations No

New Regulations to be Introduced No

Yes Yes

2001 1999/2000



Per Indian Standards IS1622:1981 and IS3025 PartI: 1987 Service pop. > 100,000 : sampled every month





Sample analyzed by the Chemistry Dept. of Malaysia based on International Standard Procedures, Preservation and Protocol under ISO Guide 25 NA NA





Aesthetic: No Investigation of complaints Health: First sample of the day at the end of the spurunflushed



No No

Sampling Basis Min. one per year from source NA Single sample Note: when copper salts are needed for algae control, a limit of 1 mg/l is permitted for a transient period. WHO Guidelines NA

Basis for Limits 1985 WHO Aesthetic parameter NA Set in 1970's. Based on soviet study suggesting embryotoxic effects of copper

WHO 1984 Guidelines, Indian Council of Medical Research 1971 Lower than WHO provisional levels to prevent staining WHO Guidelines

NA NA Staining value selected from experience with complaints which start at about 2mg/l; Projectile vomiting in children starts at about 4 mg/l NA Copper sulfate used as algicide in doses of less than 3 mg/l per week during normal water treatment, paying attention that filtered levels are free of metallic particulate

No No

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Regulates copper in Drinking Water Yes

Max. copper Levels 1.3mg/l

Guiding factors (Aesthetic, Mandatory Health) or Guideline M Health (NOAEL)











Country United States

Basis for Limits LOAEL of 5.3mg/day adjusted to std. daily dose of 2l/d by adult. Uncertainty factor of 2 applied WHO recommendation s (upper limit) NA

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Sampling Basis Sample sites selected based on high risk for Pb contamination. Number depends on system size. NA

Planned changes to regulations No

New Regulations to be Introduced No







3 A BRIEF OVERVIEW OF THE PHYSICAL, CHEMICAL AND BIOLOGICAL PROPERTIES OF COPPER A brief review of essential physical, chemical and biological attributes of copper is presented here to provide a quick reference and the necessary background for analysis of environmental and biological chemodynamics of copper, the subject of the next chapter.

3.1 Physical and Chemical Attributes of Copper

Since early history (see Chapter 2) copper's unique combination of properties have made it one of the world's most important metals. These properties include among others: • • • • • •

appearance, malleability, low corrosion, alloying ability, high thermal conductivity, and high electrical conductivity.

Properties of metallic copper, such as electrical conductivity and fabricability, vary markedly with purity. Standard classifications have been defined according to processing method. For example, ASTM B5-74 is 99.90% pure and is the accepted basic standard for electrolyte copper wire bars.

General Chemical Properties

Copper is a transition metal (the first element of Group IB of the periodic table) with atomic mass of 63.54 daltons (Da), it has two stable isotopes, 63Cu and 65Cu, with natural abundances of 69.2 and 30.8%, respectively.* Copper generally occurs in nature in one of four oxidation states: copper(0), copper(I), copper(II), and copper(III); although trivalent copper is very rare. Along with silver and gold, it is classified as a noble metal and, like them, can be found in nature in an elemental form. Indeed, copper metal is fairly unreactive, due to its high nuclear charge, small size, and consequent high ionization potential. Having one more electron than nickel and a higher nuclear charge, copper has a smaller atom and more tightly bound electrons. Since it is positioned below hydrogen in the electromotive-force series, it will not displace hydrogen ions from dilute acid. Accordingly, copper will not dissolve in acid unless an oxidizing agent is present. Therefore, while it readily dissolves in nitric acid and hot concentrated sulfuric acid, it dissolves slowly in hydrochloric acid and dilute sulfuric acid, only when exposed to the atmosphere. It is also attacked by acetic acid, and other organic acids. Copper is stable in pure dry air at room temperature, but in damp air it forms a green patina of basic salts such as the carbonate (and, near the sea, the chloride). This tightly adhering coat protects the underlying * There are seven radioisotopes of copper, most with half-lives of seconds or minutes; the two with the longest half-lives, 67Cu (61.9h) and 64Cu (12.9h), and the two stable isotopes are used as tracers of copper metabolism.

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metal from further attack and is also prized for its appearance. Despite its low reactivity, copper forms a wide range of compounds, and its complexes, both of copper(II) and of copper(I), are amongst the strongest known for their oxidation states. Although copper is a soft metal, it can form harder alloys with metals that contribute more outer electrons, such as tin and zinc. Copper, with its high nuclear charge and small size, is found in combination with both oxygen and sulfur. Copper is the only member of the 3d series that, in nature, occurs in the (I) state (e.g. as Cu2O or Cu2S), as well as in the (II) state (e.g. as the basic carbonate Cu2CO3(OH)2). Copper can also exist in higher oxidation states, but these are rare. Currently an area of very intense investigation is that of mixed oxide ceramic superconductors which copper forms with other cations. These function at relatively high temperatures (up to 125 K for Ti2Ba2CaCu3O10). The oxide Yba2Cu3O7-x (nicknamed “1-2-3”) appears to contain copper in oxidation states (I), (II), and (III), but little is known little about how it functions.*


Copper(I), the cuprous ion, disproportionates rapidly in aqueous solution to form copper(II) and copper(0), the change being favored by the high hydration energy of Cu2+ compared with Cu+. The Cu+ ion is, however, stabilized by solvents such as CH3CN, which solvate it strongly; and by groups with which it forms insoluble solids, such as CuI and CuCN. It has been shown, however, that copper(I) complexes may be formed in seawater by photochemical processes and may persist for several hours (Moffett & Zika, 1987a). Cuprous compounds are generally colorless. Copper(I) forms a number of polymeric species; tetramers are particularly favored. Copper(I) also forms a variety of simple organometallic compounds: the explosively unstable ionic acetylicle Cu2C2; sigma-bonded aryls and alkyls, CuR; and complexes containing pi-bonded ethene and ethyne. The chloride combines with CO to form the 18-electron compound Cu(CO)3Cl but no other copper carbonyl compound has been made at ordinary temperatures. Copper(II)

Copper(II), the cupric ion, is the most important oxidation state of copper; indeed, it is the oxidation state generally encountered in water. Cupric ions are coordinated with six water molecules in solution; the arrangement of the water molecules is distorted in that four molecules are closely bound to the copper in a planar array and the other two are more loosely bound in polar position. Most cupric compounds and complexes are blue or green, and are frequently soluble in water. Cu2+(aq) is mildly hydrolysed in near-neutral solution, forming the dimer Cu2(OH)22+. Addition of alkali gives a precipitate of the hydroxide, which is somewhat soluble in excess of alkali, probably forming the Cu(OH)42- anion. Cu2+ ion forms stronger complexes than other doubly charged metal ions, and, because part of this enhanced stability is a result of ligand field effects, it is not surprising that nitrogen ligands, which produce relatively high fields, bind copper(II) even more strongly than do lower field oxygen donors. Copper(II) complexes are *

This compound of copper received extensive attention following publication of reports that a spinning disc of 1-2-3 causes a small (2%) decrease in the force of gravity. (Rossotti, 1998)

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almost always blue or green. With halide ions, copper(II) shows a rich diversity, except with iodide, which it oxidizes to I2.

3.2 Biological Functions of Copper

In general, copper’s biological functions involve electron transfer catalysis, by means of its two accessible oxidation states (given suitable ligands), which differ by one unit, and can tolerate a variety of geometries. Copper is widely used by microorganisms, plants, and animals as a component of many electron transfer enzymes. Detailed descriptions of these proteins and their functions have been published (Owen, 1982a; Linder, 1991a; Linder, 1991b). A dimeric dioxygen-bridged copper protein, haemocyanine, is used as an oxygen carrier in some invertebrates. Other such proteins include ascorbate oxidase, carboxypeptidase A, laccase, and uricase. Like many other functional bioinorganic proteins, their mode of action is incompletely understood. The unraveling of these processes, together with a deeper insight into reasons for the varied stereochemistry of Cu(II), and for the formation of so many copper dimers and cluster compounds, expected to be subjects of research for many years. It should be noted that copper is most often in biological systems as Cu2+; in fact, at least three distinct “types” of the bound cation can be found in copper containing enzymes, often in combination within a single protein (Owen, 1982a; Owen, 1982b): • • • 3.2.1

Type 1 refers to deep blue proteins, typically copper-containing oxidases, Type 2, characteristic of many multicopper oxidases, is not blue, but is detectable by electron paramagnetic resonance (EPR), Type 3, is neither blue, nor detectable by EPR. Biochemical Functions

A brief description (adapted from Turnlund, 1999) of the major copper-containing proteins found in the human organism follows; for details the reader should consult (Owen, 1982a; Linder, 1991a).

Copper-Containing Enzymes Found in Humans

Amine Oxidases Several important amine oxidases are cuproproteins. Relatively small amounts of these enzymes are found circulating in blood plasma, where they inactivate and catabolize physiologically active amines such as histidine, tyramine, and polyamines. They are found in tissues throughout the body. Their activity is elevated when connective tissue activation and deposition take place, in such conditions as liver fibrosis, congestive heart failure, and hyperthyroidism, and during childhood, and senescence.

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• •

Monoamine Oxidase. Involved in inactivation of catecholamines, monoamine oxidase reacts with substances such as serotonin, norepinephrine, tyramine, and dopamine, and is inhibited by tricyclic antidepressant drugs. Diamine Oxidase. A number of copper-dependent diamine oxidase enzymes are found in cells throughout the body. Diamine oxidase inactivates histamine, acting in the small intestine, where histamine stimulates acid secretion, and in allergic reactions throughout the body, where histamine is released in response to antigens. It also inactivates polyamines involved in cell proliferation, which suggests that diamine oxidase may play a role in limiting excessive growth. Diamine oxidase activity is highest in the small intestine. Activity is also high in the kidney, where diamine oxidase inactivates diamines filtered from the blood, and in maternal placenta, where it may inactivate amines produced by the fetus. Lysyl Oxidase. Lysyl oxidase, a unique amine oxidase, functions in the formation of connective tissue, including bone, blood vessels, vasculature, skin, lungs, and teeth. It acts on lysine and hydroxylysine side chains of collagen and elastin; it eliminates the lysine of newly formed, immature elastin and collagen, after which cross-links are formed. Concentrations are highest during development. Long-term estrogen treatment increases its activity, and malignant transformation decreases it. Peptidylglycine-a-Amidating Monooxygenase. A newly identified cuproenzyme, peptidylglycine-a-amidating monooxygenase, is involved in the synthesis of many bioactive peptides and may be influenced by copper deficiency.

Ferroxidases • Ceruloplasmin. Ceruloplasmin, also called ferroxidase 1, is an alpha-2 glycoprotein with a molecular weight of about 132 kDa. It contains six (possibly seven) atoms of copper per molecule, including all three types of copper(II) atoms described previously. Four copper atoms appear to be involved in the oxidation/ reduction reactions catalyzed by the enzyme. The role of the other atoms is not yet understood. Ceruloplasmin catalyzes ferrous iron oxidation and plays a role in the transfer of iron from storage to hemoglobin synthesis sites. It also oxidizes aromatic amines and phenols. 60 to 95% of copper in blood plasma is bound to ceruloplasmin, and this fraction appears to be relatively constant within an individual, while varying considerably among individuals. • Ferroxidase II. Ferroxidase II also catalyzes ferrous iron oxidation. It accounts for only about 5% of ferole in some animal species. Cytochrome c Oxidase Cytochrome c oxidase enzyme, present in the mitochondria of cells throughout the body, is the terminal link in the electron transport chain. It reduces O2 to form water and permits formation of adenosine triphosphate (ATP) in mitochondrial energy production. Cytochrome c oxidase is considered the single most important mammalian cell enzyme, because it is rate limiting in electron transport. It contains two or three copper atoms per molecule. The activity of this enzyme is high in brain, liver, and kidney tissues, highest in the heart.

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Dopamine beta-Hydroxylase Dopamine beta-hydroxylase catalyzes the conversion of dopamine to the neurotransmitter norepinephrine in the brain. Estimates of the copper content of dopamine beta-hydroxylase range from two to eight atoms per molecule, the most recent estimates being eight. Dopamine beta-hydroxylase concentration is two to three times higher in brain gray matter than white matter, and it is present in the adrenal gland, where it is required for epinephrine production. Superoxide Dismutase • Extracellular Superoxide Dismutase (EC-SOD). EC-SOD, a copper-containing enzyme, is present in high amounts in the lungs, thyroid, and uterus and in small amounts in blood plasma. It functions as a scavenger of superoxide radicals and protects against oxidative damage. • Copper/Zinc Superoxide Dismutase (SOD). Copper/zinc SOD, which contains two copper atoms per molecule, is present in most cells of the body, primarily within the cytosol. High concentrations are found in human brain, thyroid, liver, pituitary, erythrocytes, and kidney. SOD erythrocyte levels are high in alcoholics and individuals with Down's syndrome. It protects intracellular components from oxidative damage, converting the superoxide ion to hydrogen peroxide, and requires both zinc and copper for catalytic function. Tyrosinase Tyrosinase catalyzes conversion of tyrosine to dopamine, and oxidation of dopamine to dopaquinone, steps in melanin synthesis. It is present in the melanocytes of the eye and skin and is responsible for hair, skin, and eye color. Tyrosinase deficiency causes albinism.

Copper-Binding Proteins

Metallothionein (MT) MTs are small nonenzymatic proteins, rich in cysteine, that are responsible for binding copper. Each molecule can bind 11 or 12 copper atoms, as well as zinc and cadmium. They appear to play a role in metal storage by sequestering excess metal ions, thus preventing toxicity. MTs are found in many human tissues, with highest concentration in the liver, where metals accumulate in MT fractions. Their presence in small amounts in blood plasma suggests that they also play a role in copper transport, but if so, the role would be minor. Albumin Albumin, a protein with a molecular weight of 68,000, is the most prevalent protein in blood plasma and interstitial fluids. Albumin binds and transports copper and may also play a role in binding excess copper that would otherwise be toxic. Estimates of the fraction of copper in blood plasma bound to albumin range from 5 to 18%.

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Transcuprein Transcuprein, a recently isolated plasma protein with a molecular weight of about 270,000, binds copper and is found in humans. It has not yet been completely characterized and its functions are not clear, but it may play a role in copper transport. Considerably less serum copper is bound to transcuprein than to albumin. Blood-Clotting Factor V Blood-clotting factor V, a nonenzymatic component of the blood-clotting process, has recently been found to contain one atom of copper per molecule. Although this indicates that copper is required for blood clotting, impaired blood clotting is not among the reported manifestations of copper deficiency.

Low-Molecular-Weight Ligands

Amino acids and small peptides also carry a small fraction of the copper in the blood plasma. Estimates range from less than 1% to 4%. Histidine, glutamine, threonine, and cystine are examples of amino acids that bind copper in plasma, and at least one copper peptide complex, glycyl-histidine-lysine, has been isolated from human plasma. The role of these complexes is not known, but the copper carried by low-molecular-weight ligands is thought to exchange with nonceruloplasmin copper in the blood. The ligands may carry copper to cells. 3.2.2

Physiologic Functions

Much of copper’s physiologic activity is related to reactions catalyzed by cuproenzymes, some is related to copper deficiency. A brief review, based on Turnlund, 1999, is included here for convenient reference; more detailed information on the physiologic functions of copper can be found in (Davis & Mertz, 1987). •

Connective Tissue Formation. Copper, through the enzyme lysyl oxidase, is essential for cross-linking collagen and elastin, both required for formation of strong, flexible connective tissue. Thus, copper plays a role in bone formation, skeletal mineralization, and heart and vascular system connective tissue integrity. Lysyl oxidase activity declines during severe copper deficiency in weanling rats, and the resulting defects in connective tissue formation may be responsible for the multiple effects of copper deficiency on cardiac system integrity and bone formation. Copper depletion also results in modest changes in lysyl oxidase activity in the skin, which, owing to its large excess there, does not compromise function. Iron Metabolism. Several mechanisms have been proposed for the role of copper in iron metabolism and N-thropoiesis. Ceruloplasmin and ferroxidase oxidize ferrous iron, so that it can be transported from the intestitial lumen and storage sites to sites of erythropoiesis. This may explain why anemia develops with copper deficiency, through iron accumulates in the intestinal lumen and liver. Copper may also be required for formation of normal bone marrow cells, necessary for red blood cell formation.

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Central Nervous System. Copper plays more than one role in the central nervous system. It is required for formation and maintenance of myelin, a protective layer covering neurons, composed primarily of phospholipids. Phospholipid synthesis depends on cytochrome c oxidase activity, which may explain why copper deficiency leads to poor myelination, necrosis of nerve tissue, and neonatal ataxia in copper-deficient animals. The role of cuproenzymes in catecholamine metabolism (conversion of dopamine to norepinephrine by dopamine beta-hydroxylase and the degradation of serotonin, norepinephrine, tyrainine, and dopamine by monoamine oxidase) implies a function in normal neurotransmission. Melanin Pigment Formation. The role of copper in the pigmentation of skin, hair, and eyes is related to the requirement for tyrosinase in melanin synthesis. Depigmentation of hair and skin is observed with copper deficiency in several animal species and in Menkes' disease. Cardiac Function and Cholesterol Metabolism. The role of copper in cardiac function has been explored in a number of laboratory animal experiments. Cardiac myopathy and a variety of other conditions appear when weanling, but not older, rats are deprived of copper. Cardiac symptoms have not been reported in the few frankly copper-deficient humans, though links to heart irregularities in humans have been suggested. Blood cholesterol levels increase in animals fed copper-deficient diets, but results of studies on the effects of low-copper diets on blood cholesterol in humans are inconsistent; levels have increased in some and declined in others, and copper supplementation increased low-density lipoprotein (LDL) in a study in adult males. Other Functions. Other physiologic functions suggested for copper are not as well understood as those described above. These include roles in thermal regulation and glucose metabolism. Its known role in blood clotting through factor V has not yet been clearly associated with clinical manifestations of copper deficiency. Copper is known to be both prooxidant and antioxidant. Two key antioxidant enzymes, ceruloplasmin and superoxide dismutase, decrease in copper deficiency and may result in impaired antioxidant status. Recent evidence suggests a role for copper in immune function; both low- and high copper intake influenced immune function in laboratory animals. Some indices of immune function declined with copper depletion of humans, but were not reversed by higher copper intake. The effect of dietary copper on these functions is still the subject of research (Turnlund, 1999).

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Table 2: Copper Containing Proteins Found in Humans Protein Monoamine Oxidase

Site of Action Brain

Diamine Oxidases Lysyl Oxidase

Small intestine, kidney, and other tissues Extracellular matrix

Ferroxidase 1I

Blood plasma

Cytochrome c Oxidase Dopamine β-hydroxylase

Mitochondria of cells throughout the body Brain – adrenal gland


Melanocytes of the eye and skin

Metallothionein Albumin Transcuprein Blood-Clotting Factor V Extracellular Superoxide Dismutase Copper Superoxide Dismutase

Liver (and GI tract tissues) Blood plasma Blood plasma Blood plasma Extracellular matrix, especially lung, thyroid, and uterus Most cells in body

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Function Inactivates serotonin, norepinephrine, tyramine, and dopamine Inactivates histamine, diamines, and polyamines Eliminates lysine and hydrolysine side chains of collagen and elastin Catalyzes oxidation of ferrous iron Permits formation of adenosine triphosphate (ATP) Catalyzes conversion of dopamine to norepinephrine Catalyzes conversion of tyrosine to dopamine and oxidation of dopamine to dopaquinone Binds copper (and other metals) Binds copper Binds copper Required for blood clotting Scavenges superoxide radicals


Table 3: Selected Industrial Copper Compounds and their Properties Name


Copper Abietinate (Cupric Abietinate) Copper Acetate


Copper Acetate, Basic Copper Acetoarsenite

Cu(C2H3O2)2-CuO-6H2O C4H6O16Cu4As6

Copper Ammonium Sulfate Copper Arsenate, Basic (Cuprous Arsenate) Copper Arsenide

CuSO4-4NH3-H2O Cu(CuOH)AsO4

Copper Arsenite (Cupric Arsenite, Sheele's Mineral)


Copper Boride (Cupric Boride) Copper Carbonate Hydroxide (Cupric Carbonate)


Copper Chloride (Cupric Chloride)


Copper-Y-Chloroaceto Acetanilide Copper-8-Cunilate Copper Cyanide (Cupric Cyanide)


Cu(C2H3O2)2 H2O




Copper Diazo Amino Benzene Copper Dichlorobenzoate

Physical/Chemical Properties Green scales

Toxicity Studies

Greenish-blue, fine powder or small crystals; mp: 115°, bp: 240°, d: 1.882, (anhy): 1.93

Oral LD50 (rat) = 710 mg/kg

Emerald green powder

Acute tox data: Oral LD50 (rat) = 22 mg/kg; LD50 oral (mammal) = 18 mg/kg

Black crystals; mp: decomp, d: 7.56 Yellowish-green powder; mp: decomp. Yellow crystals; d: 8.116 Green powder; mp: decomp @ 200°, d: 4 Yellowishbrown hygroscopic powder; mp: 498°, d: 3.054

Yellowish-green powder; mp: decomp before melting Orange crystals; mp: 270° (decomp)

Copper Dimethyl Dithiocarbamate

Acute tox data: Oral LD50 (rat) = 159 mg/kg; oral LD50 (birds) = 810 mg/kg Acute tox data: Oral LD50 (rat) = 140 mg/kg; oral LD50 (human) = 200 mg/kg

Acute tox data: ip LD50 (rat) = 50 mg/kg


Copper Gluconate (Cupric Gluconate)


Light blue, fine crystalline powder

Copper Hydride


Copper Hydroselenite


Red-brown crystals; mp: decomp @ 60°, d: 6.38 Bluish-green, tiny prisms

HIGH via oral and inhal routes. Used as a fungicide. Also a trace mineral added to animal feed. See copper compounds acetanilide and chlorides. See copper compounds HIGH via ip route. See cyanides and copper compounds See copper compounds

Acute tox data: ip LD50 (rat) = 25 mg/kg

Copper Fluoride (Cupric Fluoride) Copper Fluoroacetic Acid

Summary Toxicity Statement

Monoclinic blue crystal; d: 2.9 Acute tox data: Oral LD50 (rat) = 10 mg/kg

A toxic material. See copper compounds. Used as a fungicide. HIGH via ip route. See carbamates and copper compounds See fluorides and copper compounds. HIGH via oral and inhal routes. See fluorides U. See copper compounds. A nutrient and/or dietary supplement food additive. Also a trace mineral added to animal feed See copper compounds and hydrides See selenium and copper compounds.

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Physical/Chemical Properties Blue gelatinous or amorphous powder

Toxicity Studies

Copper Hydroxide (Cupric Hydroxide)


Copper Naphthenate


Solid; flash p: 100°F, d: 1.055

Copper Nitrate (Cupric Nitrate)


Copper Nitride


Copper Nitrodithioacetate Copper Oleate (Cupric Oleate)


Blue, deliquescent crystals; mp: 114.5°, d: 2.047 Dark-green powder; mp: decomp @ 300°, d: 5.84 @ 25°/4° Solid Brown powder or greenish-blue mass

Acute tox data: Oral LD50 (mouse)- 110 mg/kg Oral LD50 (rat) = 940 mg/kg

Copper Oxalate (Cupric Oxalate) Copper Oxide, Black (Cupric Oxide, Paramelaconite) Copper Oxychloride (Brunswick Green, Cupric Oxychloride)

CuC2H4-1/2 H2O

Copper-2,4-Pentanedione Derivative (Acetylacetonate of Copper) Copper Perchlorate


Blue crystals; mp: >230°, bp: subl.


Crystalline mp: 60°.

Copper Peroxide


Copper-3-Phenyl Salicylate


Brown or brownish-black crystals Crystalline; mp: 145°.

Copper Phosphate (Cupric Phosphate)


Bluish-green powder

Copper Phosphide. Syn: Cupricphosphide.


mp: decomp, d: 6.67.

Copper Propargylate Copper Propionyl Acetate Copper Pyrophosphate



Light bluish-green powder Fine black powder; bp: decomp @ 1026°, d: 6.4. Emerald green to greenish-black powder; mp: -3H20 @ 140°.


Yellow-green powder Mp: decomp @ 210°.

Copper Resinate Copper Ricinoleate

Green powder Green plastic solid Softening p: 64°.

Copper Sebacate

Cu(C20H29O2)2 Cu2(CO2(CH2)7CHCHCH2CHOH(CH2)5) Cu(CH2)8C2O4

Copper Silicate


Greenish crystals


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MOD via oral route

See copper compounds and nitrides

Acute tox data: Oral LD50 (rat) = 700 mg/kg; oral LD50 (human) = 200 mg/ kg

Acute tox data: ip LD50 (rat) = 29 mg/kg

Acute tox data: Oral LD50 (rat) = 520 mg/kg

Solid. Crystals.

Copper-8-Quinolinolate. Syn: Copper-8-HyDroxyquinoline

Summary Toxicity Statement HIGH via oral and inhal routes. A trace mineral added to animal feeds. Used as a fungicide. [109] See copper compounds HIGH via oral and inhal routes

ip LD50 (mouse) = 67 mg/ kg. [3]

See copper compounds Used as a fungicide. See copper compounds and oleic acid. A recog carcinogen See oxalates and copper compounds. Used as fungicide. Also a trace mineral added to animal feeds HIGH to MOD via oral and inhal routes

HIGH via ip route. See perchlorates and copper compounds. See copper compounds and peroxides. MOD via oral and inhal routes. See copper compounds A trace mineral added to animal feeds. See copper compounds and phosphates. See phosphides and copper compounds. Mixed with KNO3 or KC1 O3, can explode. See copper compounds. See copper compounds. U. A trace mineral added to animal feeds. See also copper compounds and phosphates HIGH via ip route. An exper (±) neo via oral and sc routes. A fungicide. See copper compounds. See copper compounds See copper compounds. See copper compounds and silicates.




Copper Silicide


Copper Stearate Syn: Cupric Stearate Copper Suboxide Copper Subsulfate Syn: Cupric Sulfate, Basic.


Copper Sulfate. Syns: Blue Vitriol, Blue Stone, Roman Vitriol

CuSO4 - 5H2O

Copper Sulfate, Ammoniated. Syn: Cupric Sulfate, Ammoniaied Copper Sulfide. Syn: Cupric Suffide


Copper Tellurite


Copper Tetrazol Copper Thiocyanate. Syn: Cuprous Thiocyanate


Copper Trichlorophenate


Copper Xanthate. Syn: Copper Ethyl XanthoGenate. Copper Zinc Chromate


Copper Zinc Sulfate

Cu4O 4CuO-SO3


Variable in composition.

Physical/Chemical Properties White metallic crystals

Toxicity Studies

Light blue amorphous powder Olive green crystals Light blue powder

mp: 1250.

Blue crystals or blue, crystalline granules or powder mp: -4H20 @ 110°, d: 2.284. Dark blue crystals Black powder or crystal, mp: transition @ 103°, bp: decomp @ 220°, d: 4.6 Green solid.

White to yellowish powder, mp: 1084°, d: 2.85. A crystalline solid Yellow precipitate mp: decomp

mp: 850º, d: 7.53.

mp: decomp Oral LD50 (human) = 200 mg/ kg Oral LD50 (rat) = 960 mg/ kg; ip LD50 (mouse) = 33 mg./kg

Summary Toxicity Statement See copper compounds and silanes.

HIGH via oral route. See copper compounds and sulfates HIGH via ip; MOD via oral and inhal routes

See copper compounds and sulfides

See tellurium compounds and copper compounds. See copper compounds See copper compounds and thiocyanates. See chlorinated phenols and copper compounds

HIGH. A recog carcinogen Toxic. See copper and zinc compounds. Used as a fungicide

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4 ENVIRONMENTAL RELEASES OF COPPER Human economic activities, involving the production and usage of copper and copper compounds, as well as the consumption of materials (including food) that contain amounts of copper, result in the re-distribution of copper in different environmental media. Information on the magnitudes and trends of environmental copper releases is summarized in this section; the review in ATSDR, 1990 has provided the basis for this summary. Additional information on environmental releases as well as more recent information on evolving methods and databases for assessing copper exposures can be found in the forthcoming report “A Framework and Data Sources for the Assessment of Exposures to Copper” (Georgopoulos et al., 2002). More details are provided on the website of the Center for Exposure and Risk Modeling (CERM), in the section devoted to copper (

4.1 Atmospheric Releases

Copper is emitted into the air from both natural and anthropogenic sources. Since copper is a component of the earth's crust, the primary natural source of copper is windblown dust. Other natural sources of emission, in order of importance are: volcanoes, decaying vegetation, forest fires and sea spray (Davies & Bennett, 1985). Similarly, anthropogenic emission sources include: nonferrous metal production, wood production, iron and steel production, waste incineration, industrial applications, coal combustion, nonferrous metal mining, oil and gasoline combustion, and phosphate fertilizer manufacture. In 1980, it was estimated that only 0.04% of copper released to the environment is to air (Perwak et al., 1980). The EPA conducted a detailed study of copper emissions into the atmosphere to estimate exposure (Weant, 1985). The sources of emissions and the estimated quantities of copper emitted in 103 kg/yr were: primary copper smelters, 43-6000 (2100, most probable value); copper and iron ore processing, 480-660; iron and steel production, 112-240; combustion sources, 45-360; municipal incinerators, 3.3-270; secondary copper smelters, 160; copper sulfate production, 45; gray iron foundries, 7.9; primary lead smelting, 5.5-65; primary zinc smelting, 24-340; ferroalloy production, 1.9- 3.2; brass and bronze production, 1.8-36; and carbon black production, 13. Using the most probable emission value for primary copper smelters and the range for other sources, estimated United States copper emissions are 2,959,000-4,300,000 kg annually. Daily stack emission rates have been reported for three coal-burning power plants on a kg/day/1000 MW basis (Que et al., 1982). They are 0.3-0.7 and 2.00 kg/day/1000 MW for those using low-sulfur western coal and highsulfur eastern coal, respectively. Emission factors in grams of copper released to the atmosphere per ton of product have been estimated for various industries (Nriagu & Pacyna, 1988). These factors would enable estimation of an industry's copper emissions from its production volume. Missing from these emission estimates is fugitive dust arising from drilling, blasting, loading, and transporting operations associated with copper mining. The only control of fugitive dust is the manual use of water sprays. The amount of copper and other pollutants originating from a waste site in wind blown dust is of some concern. In one study, the amount of airborne copper and other heavy metals deposited near a large refuse dump that received municipal and industrial waste and sewage sludge was determined by first measuring the amount of the metal accumulated in moss bags. The deposition rate was then determined and compared with that for an agricultural control area. The mean copper deposition rates in the two areas were about the

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same; the maximum deposition rate was twice as much near the dump as in the control area (Lodenius & Braunschweiler, 1986). Only in a few cases has the form of copper released into the air been determined. In general, metals released into the atmosphere will be in particulate matter in the form of an oxide, sulfate, or carbonate. Combustion processes are reported to release copper into the atmosphere as the oxide, elemental copper, and adsorbed copper. Cupric oxide has been identified in emissions from steel manufacturing and in fly ash from oil- fired power plants and open-hearth steel mills (Perwak et al., 1980). Copper associated with fine particles (< 1 µm) tends to result from combustion and other high-temperature sources, while that associated with large particles (> 10 µm) is likely to originate from wind blown soil and dust (Schroeder et al., 1987).

4.2 Releases to Wastewater

As this section illustrates, there is extensive historical information on copper releases to wastewater. An estimated 28,848 metric tons of copper entered waterways in the United States in 1976 (Perwak et al., 1980)*. This figure represents 2.4% of the identified releases of copper to the environment. Much of this copper is associated with particulate matter. Copper is a natural constituent of soil and will be transported into streams and waterways in runoff due either to natural weathering or to disturbed soil. Sixty-eight percent of releases to water is estimated from this source. Copper sulfate use represents 13% of release to water, and urban runoff contributes 2% (Perwak et al., 1980). In the absence of specific industrial sources, runoff is the major factor contributing to elevated copper levels in river water (Nolte, 1988). In the EPA-sponsored National Urban Runoff Program, in which 86 samples of runoff from 19 cities throughout the United States were analyzed, copper was found in 96% of samples, at concentrations of 1-100 µg/l (ppb) (Cole et al., 1984). Of the 71 priority pollutants analyzed for, copper, along with lead and zinc, was the most frequently detected. The geometric mean copper concentration in runoff water was 18.7 µg/l. Domestic wastewater is the major anthropogenic source of copper in waterways (Nriagu & Pacyna, 1988). Studies in Cincinnati and St. Louis showed discharges of copper into sewer systems from residential areas to be significant, with an average loading of 42 mg/day/person (Perwak et al., 1980). Concentrations of copper in influents to 239 wastewater treatment plants (12,351 observations) were 0.0001-36.5 ppb and the median value was ~0.4 ppb (Minear et al., 1981a; Minear et al., 1981b). Inputs into Narraganset Bay, Rhode Island, in decreasing order of importance, are: sewage effluent, rivers, urban runoff, and atmospheric fallout (Santschi et al., 1984; Mills & Quinn, 1984). Ninety percent of both dissolved and particulate copper was from sewage treatment plant effluent discharged into the Providence River. The range of removal efficiencies reported for pilot and full scale sewage treatment plants suggests that removal depends strongly on plant operation or influent characteristics. The best data on typical publicly-owned treatment works (POTWs) using secondary treatment are that 5590% of copper is removed in these plants with a median and mean removal of 82% (Perwak et al., 1980). By contrast, those plants using only primary treatment had a 37% median removal efficiency. A more recent study focused on heavy metal removal in three POTWs that received primarily municipal sewage and which used activated sludge as a secondary treatment. The study *

It should be noted that ATSDR, 1990 erroneously reported this estimate as 28,848 million tons of copper.

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looked at removals in both the primary and secondary treatment stage. The average removal of soluble copper and total copper after secondary treatment was 49-82% and 83-90%, respectively. The average copper concentration in the final effluent was 17-102 ppb (Aulenbach et al., 1987; Stephenson & Lester, 1987). Releases from these facilities contribute about 8% of all copper released to water (Perwak et al., 1980). Discharges to water from active mining and milling are small, and most of the western operations do not release any water; water is a scarce resource and is recycled (Perwak et al., 1980). Runoff from abandoned mines is estimated to contribute 314 million tons annually (Perwak et al., 1980). These discharges are primarily insoluble silicates and sulfides and readily settle out. Wastewater generated from mining operations comes from seepage, runoff from tailing piles, or from utility water used for mine operation. The amount of wastewater generated ranges from 0-300 liters of water/metric ton of ore mined for open pit copper mines and 8-4000 liters of water/metric ton of ore mined underground. Copper concentrations in wastewater from a selected open pit and underground copper mine were 1.05 ppm and 0.87 ppm, respectively. Discharges from electroplating operations are either directly to water or indirectly via POTWs. Releases from copper-containing products may be substantial but difficult to predict. Corrosion of copper in plumbing or construction may result in direct discharges or runoff into waterways. Copper and brass production releases relatively little copper to water. Data regarding copper concentrations in wastewater associated with selected concentrating, smelting, and refining operations can be found in (PEDCo Environmental Inc., 1980). Results of an EPA industrial effluent survey show that mean and maximum levels of copper in treated wastewater from six industries exceeded 1 and 10 ppm, respectively. These industries and their mean and maximum discharges in ppm are: inorganic chemicals manufacturing ( 3000 Da), comprised of microparticles and macromolecules too small to settle under gravitational force. It is generally agreed that the toxicity of dissolved copper with respect to aquatic biota is dependent upon the concentration of free Cu2+ ion. In addition, evidence from studies in which pH, and therefore the relative proportions of dissolved copper species differed, suggests that inorganic complexes of copper, such as hydroxy and carbonato complexes may also be involved in copper toxicity (Allen & Hansen, 1996). Dissolved copper bound to organic ligands however, are not thought to be involved in the toxicity of copper. At constant pH, the ratio of the concentrations of free copper ion and inorganic complexes is usually constant (typically around 2.5%), because the inorganic ions with which copper forms complexes are usually in stoichiometric excess or are buffered, and are therefore relatively constant. Organic ligands form strong complexes with copper but are generally present in concentrations only slightly higher than copper. Therefore, their capacity to buffer the concentration of free copper ion is limited, resulting in a nonlinear relationship between dissolved copper and free copper ion. An important implication of this nonlinearity is that although the concentration of both dissolved copper and organic matter in a body of water are increased by the introduction of effluent, the concentration of Cu2+ may actually decrease. A quantitative framework for characterizing the bioavailability of copper to aquatic biota is provided by the Biotic Ligand Model or BLM (DiToro et al., 2000). The capacity of organic ligands in natural marine water to bind copper can range from 0.5 to 35 µg/l, depending upon the amount of ligand present, and its binding constant. Discharge of effluent enriched in both copper and organic ligands could result in either increased or decreased copper ion concentrations, depending upon the extent to which the additional organic ligand in the effluent can buffer the copper. 5.3.2

Fluxes of Copper in the Hydrosphere

Copper present in the hydrosphere comes from several types of sources. Lewis (Lewis, 1995b) summarized them as follows: •

Minerals in soil and soil parent material (e.g. weathered rock) that form sediments and suspended particles in water. Copper is a common element and occurs naturally in rocks and soils either as native copper or as a mineral. Weathering, physical disintegration and chemical decomposition of exposed parent material forms soils and causes the release of soluble copper and copper-containing compounds to water. River transport provides a means of delivering this, as well as particles containing copper, to lakes and coastal marine waters. Although parent material is an important source of micronutrients for soils, copper mostly occurs as a minor component except in areas of mineralization. Extraction of copper from soil parent material into a dissolved state. This process can occur by water percolating through soil or from particles after they are introduced into an aquatic system. It can occur when an oxidized form of copper is on the surface of the

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• •

particles, but not when it is in the interior matrix, less accessible to oxidation and dissolution. Biological particles, including both living and dead organic material and products of organism decay. Organisms acquire copper to satisfy their metabolic requirements. In doing so they remove copper from the water or sediments where they live. When organisms die, they form copper-containing particles which ultimately contribute to the mineral and organic load of the sediments. Suspended particles can scavenge (e.g. plankton) or chemically bind (e.g. organic matter) copper from the water and initially immobilize it in bottom sediments. However, this biological removal from the water is subject to remobilization by chemical or biological activity within sediments. Hydrothermal systems in which heated or chemically altered water are found. Hydrothermal sites occur infrequently. They are sulfide-rich areas that have been responsible for the formation of porphyry copper deposits over geologic time and are commercially exploited today. Metal-rich sediment associated with hot brine from the Atlantis II deep in the Red Sea is reported to contain 500 ppm of copper depletion in the water around active hydrothermal sites. This is believed due to precipitation of copper-containing sulfide minerals. Boyle (Boyle, 1979) points out that excess copper found at hydrothermal locations must come primarily from a seawater source rather than just leaching from rocks. Copper input from sediments introduced into the overlying water column from benthic sediments (bottom sediments). Geochemical conditions affecting this input vary widely. As a result, there are inherent difficulties in providing accurate estimates of input. One estimate (Nriagu, 1979d) of input of “regenerated” or “refluxed” copper for the entire ocean is 4.7 x 108 kg/yr. Anthropogenic inputs, either directly into the water or leached after deposition on land. These include industrial and municipal effluents as well as antifouling coatings, pesticide residues, manure and sludges. A major source of aerosol copper is anthropogenic, from industries like power plants that use fossil fuels. Other sources include windblown dust and particles injected into the atmosphere by volcanic activity. Aerosol deposition is most pronounced in the Northern Hemisphere, a result of the greater concentration of industrial activity there compared with the Southern Hemisphere (Soudine, 1989).

Using a simplified conceptual model, Nriagu (Nriagu, 1979b) calculated copper fallout into the oceans, assuming a dry deposition velocity of 0.4 cm/sec, an average atmospheric copper concentration of 0.5 ng/m3, and an oceanic surface area of 3.6 E18 cm2. The calculated dry deposition rate (6.34 E9 g) was then multiplied by two to convert it into a total deposition (including rainfall) of 13 E9 g over the oceans. This estimate suggests that about 20% of the copper released each year into the atmosphere is deposited in the oceans. (Weiss et al., 1975) determined the post 1945 rate of copper deposition in the Greenland glacier to be 19 to 39 ng/cm2 year. If these data are extrapolated to the world oceans, the resulting annual atmospheric deposition of copper is found to be 68 to 140 E9 g; these figures, in fact, exceed the total annual emission of copper to the atmosphere. Nriagu’s value for the atmospheric fallout of copper over the oceans, however, is roughly twice the 5 E9 g/year given in a National Academy of Sciences

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(1975) report. Clearly, all these estimates should only be interpreted as preliminary attempts to quantify attributes of a very complex system; the specific values of such estimates reported in this and other chapters of the present document should therefore also be seen from the perspective of an exploratory analysis. The following is a list of estimates of annual global copper fluxes through the aquatic environment (Martin & Windom, 1991): Input to ocean margins (both natural and anthropogenic copper). • • • • •

Soluble atmospheric flux 15 x 106 kg/yr Riverine Dissolved 58 x 106 kg/yr Particulate 1500 x 106 kg/yr Biological removal in margins 3 x 106 kg/yr

Input to open ocean • • • •

Output from margins to open ocean 145 x 106 kg/yr (Estimated total retention in coastal margins 1428 x 106 kg/yr) Atmospheric input to open ocean (natural and anthropogenic) 7 x 106 kg/yr Total removal in open ocean 220 x 1106 kg/yr

The following residence times in years: • •

Ocean margins 6 x 102 years Deep ocean 1.5 x 103 years

were calculated using the average composition of deep-sea clay and a deep-sea clay accumulation rate of 1.1 x 1012 kg/yr. Several comments need to be made about these estimates of copper fluxes: •

Estimates of riverine fluxes from land are based on particulate and dissolved metal concentrations and river flow rate.

Estimates of removal of riverine sediment inputs in the ocean's margins range from 80% to 95%. This would include particulate copper that remained in the sediment as well as any dissolved copper that became associated with the sediment in the river estuary.

The atmospheric inputs to the oceans in Martin and Windom (Martin & Windom, 1991) are from the 1989 World Meteorological Organization report by the Joint Group of Experts on the Scientific Aspects of Marine Pollution (Joint Group of Experts on the Scientific Aspects of Marine Pollution (GESAMP), 1989). This report gives an estimate of 52 x 106 kg/yr for the maximum total atmospheric flux of copper to the world's oceans.

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The soluble flux (dissolvable in seawater) global value for copper is given as 30 x 106 kg/yr or twice the input into coastal waters. Nriagu (Nriagu, 1979d) estimated the total copper reservoir in lakes and streams using an average copper concentration of 2 µg/l. According to Nriagu, this value corresponds to the average concentration in small Norwegian lakes (Henriksen & Wright, 1978) and in the Great Lakes (unpublished results by the author). By comparison, the copper burden of glaciers is 3.3 E 12 g, assuming an average copper concentration in ice of 0.2 µg/kg (Silver et al., 1975). From the freshwater biomass value of 57 E12 g and mean copper content of 5 µg/g the copper burden is calculated to be about 3 E8 g. This comprises about 0.5% of the total copper in lakes and rivers. The total net productivity for lakes and rivers accounts for about 1.0 E15 g of dry matter. If the mean copper concentration in living biomass is also assumed to be 5 µg/g, the amount of copper cycling each year through freshwater biota is estimated to be 5 E9 g; this represents about 10% of the freshwater copper pool. It follows that the residence time for copper in the freshwater biota is 22 days. Nriagu (Nriagu, 1979d) was not able to assess the sinks for copper in freshwater ecosystems. Basically, lakes and rivers are transient features, and it seems unlikely that freshwater sediments play an important role in the global copper economy. Gibbs’ (Gibbs, 1977) value (6.3 E12 g/year) for copper in river runoff was adopted by Nriagu (Nriagu, 1979d). According to Gibbs, only 1% (6.1 E10 g) of the copper transported by rivers to the oceans is in soluble form. Of the solid copper load in rivers, it was shown that 85% was associated with particulate crystalline phases, 5.7% was bound to metal hydroxide coatings on particles, 4.5% was associated with organic material, and only 3.5% was adsorbed onto suspended particulates. Gibbs (1977) also compared the concentration of copper in the transported material to the average crustal copper abundance and found a high ratio (9 to 13), pointing to continental depletion of copper. This observation is consistent with Bowen’s (Bowen, 1966) data, showing that the average copper content of soils (20 µg/g) is less than the average crustal value. The ratio of average copper concentration in the crust to that in pelagic sediments ranges from about 3 (Atlantic clays) to over 20 (North Pacific clays), further indicating continental copper depletion. High geochemical copper mobility at earth surface conditions presumably has enabled erosional removal of copper from the continents. The oceanic copper pool (about 2 E14 g) was estimated to be less than that stored in sediment pore waters (5 E15 g). All reported oceanic copper profiles show a steady increase in concentration down to the seafloor. This suggests substantial regeneration of copper from seafloor sediments, where the copper concentration in the pore waters has been found to be about 15 µg/kg (Boyle, 1979). Boyle also estimates the flux of copper from the sediments to the bottom waters to be 0.13 µg/cm2-year, a value that corresponds to 4.7 E11 g/year for the entire ocean. Although copper regenerated from the sediments amounts to only 7.5% of total riverine input, it is about an order of magnitude higher than copper delivered to the oceans in soluble form. Thus sediments must be regarded as an important source of available copper in the deep oceans. COPPER: Environmental Dynamics and Human Exposure Issues Page 56


The annual input into oceans of dissolved copper derived from antifouling paints has been estimated at 2.1 E10 g (National Academy of Sciences, 1975). The same report also estimates industrial discharges of copper into freshwater environments to be 1.4 E10 g/year. The amounts of industrial wastes and sewage sludge dumped into the ocean by the United States have been given as 3.4 E12 and 5.0 E12 g/year, respectively. Worldwide waste discharges into the oceans may be assumed to be three times those of the United States, or, 25 E12 g/year. If the average copper content of sewage sludge and industrial wastes is taken as 670 µg/g, copper input associated with waste discharge is 1.7 E 10 g/year, a value comparable with ombrogenic input of copper to the oceans. The deep ocean residence time for copper was calculated by Nriagu (Nriagu, 1979a) to be only 1500 years, considerably shorter than times reported in the literature, such as 20,000 (Goldberg, 1965), and 4100 years (Bewers & Yeats, 1977). The major cause of this discrepancy has to be the use of higher average oceanic concentrations in previous estimates. The total amounts of copper in living marine biomasses were estimated from biomass values of 2 E14 and 3 E15g for plants and consumers, respectively. On the basis of the compilation in the two volumes edited by Nriagu (Nriagu, 1979d) the average copper contents of marine plants and organisms were estimated to be 3.5 and 1.0 µg/g dry weight, respectively. Thus the copper burdens of living plants and organisms in the oceans are 7 E8 and 3 E9 g, respectively. The particulate organic carbon content of the oceans has been given as 3 E16 g. Taking the C-O-N-SP ratio in the particulate organic matter as 106-106-16-2-1 gives a value of 7 E16 g for oceanic mass particulate organic matter. If the copper content of particulate organic matter is assumed to be 7 µg/g, the copper burden in this reservoir is 50 E10 g. From the dissolved organic pool of 1.6 E 18 g and the C-O-N-S-P ratio given above, the mass of oceanic dissolved organic matter is estimated to be 3.5 E18 g. Since little information is currently available on dissolved organic matter copper concentration, the size of this copper pool cannot be estimated. Total primary oceanic production may be assumed to be 6 E16g/year dry matter. If the same copper concentration as in living marine plants (i.e., 3.5 µg/g) is assumed, annual consumption of copper by marine biota is 21 E10 g. Thus the residence time for copper in marine biota is (3.7 E9)/(2 1 E10) years, or about 1 week. The rapid rate of turnover is not surprising since copper is essential to all life forms. Oceanic sinks for copper remain essentially intractable. To a first approximation, the flux of copper from ocean to sediment should equal the sum of inputs from the atmosphere (1.3 E10 g), waste disposal (1.7 E10 g), and the river-suspended load (624 E10 g), that is, about 6.3 E12 g/year. The flux rate for copper and other trace metals obtained using the marine sedimentation rate was only 35% of the rate obtained on the basis of the stream supply. 5.3.3

Chemistry of Copper in the Hydrosphere

The chemistry of copper affects its distribution and final fate after release into aquatic ecosystems, and its availability and potential toxicity to biota. The physicochemical form of copper is of major importance in determining the nature of its interaction with chemicals and particles and its role in the sedimentary cycle.

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Copper occurs in one of four oxidation states with copper(I) and copper(II) being the most common. Copper(I) is found primarily in anaerobic conditions and is readily oxidized to copper(II). Ionic copper is chemically labile, reacting with a variety of inorganic and organic ligands, as well as with many types of particulates. As a result, the element is found in soluble, colloidal and particulate forms. Assessment of potential biological effects of anthropogenic copper must consider metal concentration, speciation, and the changes in its chemistry that will occur after the metal is introduced into the environment. Although mass budgets of copper may be of interest in determining elemental loading, the values must not be applied directly to the determination of biological effects. The biological availability of copper is affected by natural processes, particularly complexation/chelation and sorption/desorption (Harrison, 1985). The picture is further complicated by biological processes such as changes in metal speciation of organically bound metals caused by bacterial degradation. These processes must be taken into account when considering the nature of copper in natural environments and when attempting to apply geochemical models to predict the effects of anthropogenic copper. Often, organisms are used to “assay” water quality and the biological effect or availability of metals in aquatic systems. Cairns (Cairns, 1990a; Cairns, 1990b) reviews the development of biomonitoring in aquatic ecosystems, commenting on the importance of multispecies, rather than single species, testing.

Inorganic and Organic Ligands

Copper forms complexes with hard bases (ammonia, carbonate, chloride, hydroxide, nitrate and sulfate). It reacts strongly with sulfur (to form insoluble sulfides) and with sulfur-containing ligands. Labile copper species include ions, ion pairs, readily dissociable inorganic and organic complexes, and easily exchangeable copper sorbed on colloidal inorganic or organic matter. Copper forms stable cuproorganic complexes with a number of organic anions, amino acids, amino sugars, alcohol, urea, etc.). It may also bind to high-molecular-weight organic material and occur on or in colloids. Much of the high-molecular-weight material is termed “humic substance,” a diverse group of refractory organics derived from the degradation of particulate organic materials (primarily plant material) in terrestrial and aquatic environments. Since labile dissolved copper readily associates with a variety of organic and inorganic substances, its chemistry is related to the behavior of dissolved and particulate inorganic and organic constituents of water. In seawater and in many freshwater systems, copper speciation is now recognized to be dominated by complexation with naturally-occurring organic ligands. The extent of complexation can vary widely. Examples of the importance of naturally-occurring organic ligands in metal speciation are numerous and include (Lewis, 1995a): •

A study of the Tamar Estuary (U.K.) in which Van den Berg et al. (Van den Berg et al., 1990) found very low cupric ion activities as a result of the complexation of copper by organic ligands. Van den Berg (Van den Berg, 1992) comments that copper occurs fully complexed by organic material in the North Sea.

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A report by Gardner and Ravenscroft (Gardner & Ravenscroft, 1991) that in the Tweed Estuary in England there was a substantial excess of complexing capacity over total dissolved copper across the salinity range that was measured. They also report changes in the copper affinity of ligands in different regions of the estuary.

In North Pacific waters, Coale and Bruland (Coale & Bruland, 1988) found over 99% of the total dissolved copper associated with organic ligands at depths less than 200 m.

Oganesyan and Babayan (Oganesyan & Babayan, 1987) found that in the surface waters of the Sevan Lake Basin of Armenia, “approximately 90% of the copper was complexed, 40% of it in compounds with a molecular weight of 9,000 or more” (translation).

Bugenyi and Lutalo-Bosa (Bugenyi & Lutalo Bosa, 1990) examined possible effects of excess copper from a copper mining region in the Western Rift Valley of East Africa. In a saline lake, they found that detrimental effects were reduced by both organic and inorganic agents.

Organic metal-complexing ligands are produced as products and byproducts of metabolism as well as of the breakdown of biological material. The chemical properties of the ligands can change as a result of changes in the nature of the organisms producing them. A variety of naturally-occurring agents collectively classified simply as “dissolved organic matter” (DOM) include humic substances that have demonstrated capacity to affect both copper toxicity and copper uptake by aquatic organisms. Humic substances which complex copper and other cations include both humic and fulvic acids. Recent work suggests that humic acids (especially the higher-molecular-weight fractions) have an enhanced ability to bind metal ions at higher pH as a result of their high surface potential. Particles can provide a major source of copper in rivers, lakes and oceans, a source which can vary seasonally and over longer periods. Since many types of particles can adsorb metallic ions, the concentration of copper in particulate form has been used as an indicator of the pollution status of a body of water. Copper reactions with particles are controlled by processes including • • • •

sorption, desorption, complexation (including chelation) and coprecipitation.

Since organisms are particles, this list can also include processes associated with copper uptake by organisms. From a physical/chemical standpoint, the surface of a particle can be complicated and capable of interacting with a number of metallic ions as well as with other components in water or sediment. The dynamics of the processes occurring on the surface can also change, for example, when the particle carries a coating of organic matter. A great deal is still to be learned about the factors that affect the uptake and exchange of metallic ions by particles. The lack of knowledge is unfortunate since particles not only play major roles in metal transport but can also serve as metal sinks. Chemical models do exist that attempt to address some of the problems

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associated with the scavenging of metal ions by marine particulate matter (Clegg & Sarmiento, 1989). The movement of particles in aquatic environments is determined by their size and density and by the speed and turbulence of water currents. As a result, particulate copper may be deposited in bed-load sediments in the immediate vicinity of a source or carried into adjacent environments. Settling particles can account for a major flux of metals to the sediments in both freshwater and saltwater environments. Chemical factors, such as dissolved oxygen, or physical factors, such as the overturning of lakes in winter, will affect sedimentation rates and the fate of the copper on or in the particles.

Effect of acidic deposition

Acidic deposition not only causes corrosion but reduces the ability of upland soils and lake sediments to retain atmospherically deposited trace metals. Although the degree of acid rain effect is open to debate, certainly at low levels of acidification there is evidence a decrease of pH over a range of approximately >5 to 2.5 increases leaching of metal from soils and other substrates, which increases soluble metal concentration in runoff water.

Behavior of copper in estuaries

Changes occur in both the concentration and chemical nature of riverine copper when it is introduced into salt water in estuaries. From a physical perspective, the speed of a river decreases when it enters an estuary. This allows copper-containing particles to settle to the bottom. From a chemical perspective, river water acquires major ions when it mixes with salt water in an estuary. These ions can have an important effect on the chemistry of copper. They react with both particulate and what is termed dissolved (< 0.4 pm) material (e.g. humic substances) in rivers. The flocculation of humic substances that occurs in estuaries is a result of this. Copper in a filterable state is transformed into particulates, which often settle to the bottom. This process results in the loss of dissolved metal and metal-containing material within estuaries (e.g. Sholkovitz, 1980). The net effect is an increase in particulate copper and a decrease in dissolved copper within estuaries. Sedimentation of both inorganic and organic particles retains riverine copper in coastal sediments. Estuaries and coastal waters thus act as trace metal sinks or “traps” for copper (Martin & Windom, 1991). 5.3.4

Drainage from Urban Areas

Elevated levels of copper have been reported in drainage from urban areas and roadways. As a result, urban catchment areas and detention ponds can be enriched with copper as well as with other metallic elements. Much of the copper will be associated with fine grained sediments. Mesuere and Fish (Mesuere & Fish, 1989), for example, note that copper was the dominant metallic element in a detention pond system in a parking lot near Portland, Oregon. They report that “Copper was found to be deposited in the pond sediments in a small but highly concentrated

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plume (up to 130 mg kg-1), extending axially from the runoff inlet pipe.” (Note: With the banning of asbestos in automotive brake linings, copper and brass are extensively used in new formulations. Copper is also an additive in some lubricating oils and diesel fuels.) Detention ponds can be a useful management practice for controlling runoff from parking lot areas, although water from detention ponds and runoff from storm drains is a source of anthropogenic metal. 5.3.5

Copper in Groundwater

Copper enters surface and groundwater from natural and anthropogenic sources. The latter include leachate from landfills, chemical waste sites, detention ponds and deep-well injection. The chemistry of leachate and the degree of complexation of copper differs between the various types of sources and also among sites. This provides a range of copper forms and species and makes it difficult to model the chemical reactions and transport of copper in the aquatic environment. The predictive ability of groundwater flow and transport models is limited by uncertainties in estimating some of the field parameters. Recent work includes models to simulate the effect of sediments in highway detention ponds on metal uptake and introduction into ground water. Evidence suggests that detention ponds can operate for many years before they become saturated and create the potential for groundwater contamination. 5.3.6

Mathematical Modeling of Metal Speciation

Estimates of the geochemical and biological impacts of copper require an understanding of metal speciation. Because of the complexity of natural environments, no single generalization about the forms of copper that might be expected will be accurate. With certain assumptions, equilibrium speciation models can help to resolve ambiguities concerning metal partitioning in natural environments (Campbell et al., 1988). The results of mathematical models of metal partitioning can be difficult to apply to natural systems, however, for a number of reasons, including problems with multiple substrate systems and localized chemical variability. A number of speciation models have been constructed using available information about pH, inorganic cations and anions and organic ligands. Some of the models include particulates; others relate to metals in sediments; most provide an indication of the fate of labile metal species, which does relate to part of the biologically available metal load, although the models do not include lipid-soluble metal species, which may be biologically important. The copper(I) ion is unstable in aqueous solution, tending to disproportionate to copper(II) and copper metal unless a stabilizing ligand is present (Callahan et al., 1979). The only cuprous compounds stable in water are insoluble ones such as Cu2S, CuCN, and CuF. In its copper(II) state, copper forms coordination compounds or complexes with both inorganic and organic ligands. Ammonia and chloride ions are examples of species that form stable ligands with copper. Copper also forms stable complexes with organic ligands such as humic acids, binding to -NH2 and -SH groups and, to a lesser extent, to -OH groups. Natural waters contain varying amounts of inorganic and organic species; this affects the complexing and binding capacity of the water and the types of complexes formed. In seawater, organic matter is generally the most important complexing agent (Coale & Bruland, 1988). The formation of ligands may affect other

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physicochemical processes such as adsorption, precipitation, and oxidation-reduction in water (Callahan et al., 1979). The major species of soluble copper found in freshwater, seawater, and a combination of the two over a range of pHs is Cu2+, Cu(HCO3)+, and Cu(OH)2 (Long & Angino, 1977). At the pH values and carbonate concentrations characteristic of natural waters, most dissolved copper(II) exists as carbonate complexes rather than as free (hydrated) cupric ions (Stiff, 1971a; Stiff, 1971b). Dissolved copper concentration depends on factors such as pH, the oxidation-reduction potential of the water, and the presence of competing cations (Ca2+, Fe2+, Mg2+, etc.), anions of insoluble cupric salts (OH-, S2-, PO43-, Co32-), and organic and inorganic complexing agents. If the concentration of a particular anion is high enough to exceed the solubility of a copper salt, precipitation of that salt will occur. The most significant precipitate formed in natural waters is malachite [copper 2(OH) 2CO3]; other important precipitates are copper (OH)2 (and ultimately CuO), and azurite [copper 3(OH) 2(CO3)2]. In anaerobic waters, copper 2S, copper 2O, and metallic copper forms and settles out (Callahan et al., 1979). The combined processes of complexation, adsorption, and precipitation control the level of free copper(II). The chemical conditions in most natural water are such that, even at relatively large copper concentrations, these processes will reduce the free copper(II) concentration to extremely low values. As a result of all the aforementioned physico-chemical processes, copper in water may be dissolved or associated with colloidal or particulate matter. Copper in particulate form includes precipitates, insoluble organic complexes, and copper adsorbed to clay and other mineral solids. In a survey of nine rivers in the United Kingdom, 43-88% of the copper was in the particulate fraction (Stiff, 1971b). A study using suspended solids from the Flint River in Michigan found that the fraction of adsorbed copper increased sharply with pH, reaching a maximum at a pH of 5.5-7.5 (McIlroy et al., 1986). The colloidal fraction may include hydroxides and complexes with amino acids. The soluble fraction is usually defined as that which will pass through a 0.45 mm filter; it includes free copper and soluble complexes, as well as fine particulates and colloids. The importance of copper's association with inorganic and organic ligands will vary depending on the pH and concentration of competing ligands in the body of water. In river water from the northwestern United States that had a relatively high pH (7.0-8.5) and alkalinity (24-219 ppm as CaCO3), inorganic species like CO32- and OH- were the most important ligands at high copper concentrations. However, other species were important at low copper concentrations. On the other hand, samples from lakes and rivers in southern Maine with a relatively low pH (4.6-6.3) and alkalinity (1-30 ppm as CaCO3) were largely associated with organic matter. After a period of rain in southeastern New Hampshire, inorganic constituents contributed more to the copper binding in lakes and rivers than dissolved organic matter did (Truitt & Weber, 1981a; Truitt & Weber, 1981b). Runoff induced by the rain had added to the inorganic load of the rivers and lakes, as was evidenced by their pH (5.7-7.4) and alkalinity (1.7-43.4 ppm as CaCO3). A green precipitate, confirmed to be malachite [copper 2(OH) 2CO3] formed in river water in Exeter; this water had the highest pH and alkalinity. A computer simulation of the copper species in pond water and artesian well water that fed the pond predicted that 98% of the copper in the artesian

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well water would be bound to organic matter, whereas 88% and 63% of the copper in pond water would be bound to organics in spring and fall, respectively (Giesy et al., 1983). 5.3.7

Copper in Seawater

Seawater samples obtained in a transect of uppermost Narragansett Bay in August 1980 were analyzed for dissolved, particulate, and organically-bound copper to investigate the geochemistry of copper-organic complexes (Mills & Quinn, 1984). Narragansett Bay is a partly mixed estuary in Massachusetts and Rhode Island that receives organic matter and metals from rivers, municipal and industrial effluents, and runoff. Dissolved copper represented 60% of the total copper and ranged from 16.4 µg/kg in the Providence River to 0.23 µg/kg in Rhode Island Sound. Analysis of the data indicated that ~75% of this copper is removed within the Providence River. Particulate copper concentrations ranged from 2.42-0.06 µg/kg and generally comprised 40% of the total copper. 14% to 70% (0.12-2.30 µg/kg) of the dissolved copper was complexed with organic matter. Organic ligands may contain a variety of binding sites and the strength of the resulting copper complexes will vary accordingly (see, e.g., Meyer et al., 1998). Over 99.7% of the total dissolved copper in surface ocean water from the northeast Pacific was associated with organic ligands (Coale & Bruland, 1988). The dominant organic complex, limited to surface water, was a strong ligand of biogenic origin. A second, weaker class of organic ligand was of geologic origin. An independent study showed that the copper binds to humic material at a number of sites; the binding strength of the sites varied by 2 orders of magnitude (Giesy et al., 1986). The humic material in the study was derived from nine surface waters in the southeastern United States. Soluble copper in water discharged from a nuclear power station was primarily complexed with organic matter in the 1000-100,000 molecular weight range (Harrison et al., 1980). 10% to 75% of the discharged copper was in particulate form. High pH decreases copper adsorption (Kester et al., 1975). It is important to specify the pH when attempting to predict adsorption-desorption in geochemical systems. In aqueous systems, such as the mouth of rivers, the adsorbed metals will desorb as the salinity rises, if the pH is kept constant. Both salinity gradient values and pH gradient values are needed to accurately predict the behavior of the adsorbed metals as they pass over the gradients. 5.3.8

Copper in Sediments

Many organic and inorganic pollutants have been released into both freshwater and marine environments from industrial and sewage treatment facilities, as well as through atmospheric cycling of particulate matter from power plants and vehicular emissions. During the past two decades there has been an increased interest in understanding the characteristics and processes associated with certain heavy metals emission and transport into, and fate in, aquatic systems. This type of information is fundamental to examining the effects of various discharges on the aquatic environment, their exposure to humans and, ultimately, their impact on human health.

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The following sections evaluate copper release into, and its fate in, sediment within the hydrosphere, taking into account geographical and temporal trends when possible. They will also characterize copper’s effects upon marine, freshwater and estuarine sediments. Major influences on the aquatic environment, natural and anthropogenic, will be treated separately.

Marine Sediment

Marine sediment consists of both particulate organic and inorganic matter which has accumulated on the sea floor. The sediment may originate from multiple sources, and then, after transport by wind and rain, is deposited, precipitated from solution, or secreted by organisms. Layers of sediment on the sea floor are affected by various physical, biological and chemical processes occurring throughout the hydrosphere. Sediments are composed of numerous materials and can be classified into five categories based upon their source: terrigenous, biological, cosmic, volcanic and chemical. Since the terrigenous and biological sources are by far the most prevalent, the studies summarized focused upon marine sediment arising from them. Terrigenous components of sediment are produced from erosion and weathering of rocks. The eroded material is predominantly transported to the oceans by rivers, streams and wind, and in the process, are continuously ground into finer particle sizes. In order to classify these sedimentary materials and establish working definitions for them, their size ranges have been classified and organized in Table 5. The only other major category of sediment formation originates from biological processes. For example, the skeletons of aquatic plants and animals, comprised largely of calcium carbonate and silica, are shed during reproduction or settle after death, becoming incorporated into the sedimentary matrix. In certain shallow tropical marine environments, various algal species secrete a considerable amount of calcium carbonate. The distribution of biological components within the marine sediment matrix depends upon the aquatic species present, their skeletal history, and the depth of the water column. Most sediment components are deposited near their point of introduction. The sand component is deposited close to the shoreline, with coarse and medium sands remaining close to the beach area and finer sands being swept seaward. Silt and clay materials are among the finest sedimentary components and are carried further out to sea; the concentration of terragenous material decreasing in a seaward direction. Examination of an east-west sediment profile across the Pacific basin showed sediment categories distributed symmetrically. Coarse varieties of sand are found in the surf along beach areas, while finer sand and mud comprise the sediment further offshore. The continental rise consists of a mixture of sediments of terrigenous and biological origin. Sediments of a biologic origin are usually found to a depth of approximately 4.8 kilometers, after which a clay-like sediment predominates. The north-south sediment profile contains different types of biologic sediment, but closely parallels the east-west profile. Core profiles of various sediments reveal a layered composition. Deep ocean core samples exhibit uniform composition throughout the sample, except for sediment close to the water-sediment interface. Samples taken from shallower water have a mixture of sediment types.

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Estuarine Sediment

Copper in the estuarine environment was examined separately due to the special nature of an estuarine system. An estuary is a partially enclosed body of water connected to the open ocean. Within this water body, marine water is diluted with freshwater. Estuary systems are bordered predominantly by mudflats or marshes containing a rich diversity of animals and plants. The estuarine environment is not only an important feeding and breeding ground for numerous aquatic species, but is also a prime location for residential and commercial developments. Thus, estuaries receive pollutants dumped into rivers upstream. Sediment types may vary throughout an estuarine environment. Clay particles are also known to clump within this environment when mixed with seawater.

Freshwater Sediments

Concentration profiles in freshwater sediment are often used to trace the history of metal accumulation from anthropogenic sources via long-range hydrologic and atmospheric transport. Unlike marine sediments, lake sediments are not as reliable in determining contamination trends because they are more subject to acidification which may demobilize metals. Copper binds primarily to organic matter in estuarine sediment, unless the sediment is organically poor. A study evaluated the importance of different nonlithogenic components of aerobic estuarine sediment to copper by determining copper's adsorptivity to model sedimentary phases from artificial seawater (Davies-Colley et al., 1984). These phases included hydrous iron and manganese oxides, clay, aluminosilicates, and organic matter. The binding affinities varied by over a factor of 10,000 and were in the following order: hydrous manganese oxide > organic matter > hydrous iron oxide > aluminosilicates > clay (montmorillonite). The partition coefficients at pH 7 for the more strongly bound phases (manganese oxide, iron oxide, and estuarine humic material), were 6300, 1300, and 2500, respectively. The affinity increased somewhat with pH but did not vary appreciably when salinity was reduced. Considering the compositional characteristics of estuarine sediment, the results indicate that copper binds predominantly to organic matter (humic material) and iron oxides. Manganese oxide contributes only ~1% to the binding because of its low concentration in sediment; the other phases are generally unimportant. These findings concur with results of selective extraction experiments and the association of copper with humic material (Raspor et al., 1984a; Raspor et al., 1984b). Experiments performed in synthetic seawater and water from Biscayne Bay, Florida, showed that a reaction occurred, the rate of which was first-order in Cl- and second-order in H2O2. The chloride ion is thought to be required for forming stable CuClOH-. Experiments showed that as much as 15% of copper in seawater was present as copper(I). Additionally, sunlight increases the percentage of free copper(II). The photochemical reduction mechanism is supported by the observation that the copper(I) concentration is highest in the surface layer of seawater and that the hydrogen peroxide concentration increases in parallel to that of copper(I) (Moffett & Zika, 1987b; Moffett & Zika, 1987c). In addition the percentage of free copper(II) is highest on the surface.

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Once copper(I) is formed, its lifetime is determined by its rate of oxidation to copper(II). After Biscayne Bay water was exposed to sunlight for five hours, the copper(I) formed was oxidized to copper(II); the half- life of the copper(I) was twelve hours. Dissolved oxygen is primarily responsible for this reaction. Since the oxidation of copper(I) by O2 in distilled water occurs in 64 for mineral soils and >273 for organic soils. Of the eight heavy metals in the study, only Pb and Sb had higher partition coefficients than copper. Most of the copper in Columbia River estuary sediment and soil was correlated with inorganic carbon (e.g., carbonate), but not with the amount of extractable Fe or the organic carbon content of the sediment (Fuhrer, 1986). The amount of ammonium acetate- and DTPA-extractable copper, in wetland soil/sediment, resulting from atmospheric deposition from smelters in Sudbury, Ontario, showed the same pattern as total copper, despite random variations in soil pH, redox potential, and organic carbon (Taylor & Crowder, 1983b; Taylor & Crowder, 1983c). Thus, in this case, soil characteristics were not the dominant factors determining extractability and availability; the form of copper deposited was. The median concentrations of total copper, ammonium acetate-extractable copper, and DTPA-extractable copper at 25 sample sites were 371, 49, and 98 ppm, respectively. Within the estuarine environment, anaerobic sediments are known to be the main reservoir of trace metals. Under anaerobic conditions, cupric salts will reduce to cuprous salts. The precipitation of cuprous sulfide and the formation of copper bisulfide and/or polysulfide complexes determine copper's behavior in these sediments (Davies-Colley et al., 1985). In the more common case where the free sulfide concentration is low due to the controlling coexistence of iron oxide and sulfide, anaerobic sediment acts as a sink for copper. However, in the unusual situation where the free sulfide concentration is high, soluble cuprous sulfide complexes may form, and the copper concentration in sediment pore water may then be high. In sediment, copper is generally associated with mineral matter or tightly bound to organic material. As is common when a metal is associated with organic matter, copper is generally associated with fine, as opposed to coarse, sediment. Badri and Aston (Badri & Aston, 1985) studied the association of heavy metals in three estuarine sediments with different geochemical phases. The phases were identified by their extractability with different chemicals and termed easily or freely leachable and exchangeable, oxidizable-organic (bound to organic matter), acidreducible (Mn and Fe oxides and possibly carbonates), and resistant (lithogenic). In the three

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sediments, the non-lithogenic fraction accounted for ~14-18% of the total copper, and the easily exchangeable component was 5% of the total copper. Sediment samples taken from western Lake Ontario were similarly analyzed by a series of sequential extractions (Poulton et al., 1988) with regard to the compositional associations of copper. The mean (SD) percentages of copper in the various fractions were: exchangeable, 0 (0); carbonate, 0.1 (0.3); iron or manganese oxidebound, 0.2 (0.3); organic-bound, 40 (11); and residual, 60 (8). Another study found that 10-20% of the copper in Lake Ontario sediment samples was bound to humic acids, and virtually all the copper was bound to organic matter (Nriagu & Coker, 1980). The concentration of copper associated with humic acids was 21-40 times greater than in the sediment as a whole. To determine the factors affecting copper solubility in soil, Hermann and Neumann-Mahlkau (Hermann & Neumann-Mahlkau, 1985) performed a study in the industrial Ruhr district of West Germany which has a high groundwater table (10-80 cm from the surface) and a history of heavy metal pollution. Groundwater samples were taken from six locations and two horizons of soil, an upper oxidizing loam and a lower reducing loam. Total copper concentrations were high in the upper soil horizons and low in the lower horizons. Copper showed a pronounced solubility only in the oxidizing environment; in the reducing environment, solubility was low, possibly due to the formation of sulfides. The form of copper at polluted and unpolluted sites may affect its leachability, particularly by acid rain. The leaching of heavy metals by simulated acid rain (pH 2.8-4.2) was measured by applying rainwater to columns containing humus layers from sites in a Swedish spruce forest both near to and far from a brass mill (Strain et al., 1984). Leaching of copper increased considerably when water with a pH 290 nm and can undergo charge transfer reactions where the copper(II) is reduced and the ligand oxidized. Photochemically-generated reducing agents such as O2- and H2O2 could also reduce copper(II) to copper(I).

Fish Bioaccumulation

The bioconcentration factor (BCF) of copper in fish obtained in field studies is 10-100, indicating a low potential for bioconcentration. The BCF is higher in molluscs, especially oysters, where it may reach 30,000 (Perwak et al., 1980). This may be due to the fact that they are filter feeders, and copper concentrations are higher in particulates than in water. However, there is abundant evidence that there is no biomagnification of copper in the food chain (Perwak et al., 1980). A study was conducted with white suckers and bullheads, both bottom-feeding fish, in two acidic Adirondack, NY, lakes (Heit & Klusek, 1985). These lakes were known to have received elevated loadings of copper, but the suckers and bullhead had average copper levels of only 0.85 and 1.2 ppm (dry weight) in their muscle tissue. The biomagnification ratio (the concentration of copper in the fish to that in potential food) was 1000 wells and 600 surface sites were sampled, the median copper levels in groundwater and surface water were 5.0 and 3.0 µg/l, respectively (Page, 1981). The respective 90th percentile and maximum levels were 64.0 and 2783.0 µg/l for groundwater and 9.0 and 261.0 µg/l for surface water. The pattern of contamination in surface water correlates with light hydrocarbons, while that in groundwater correlates with heavy metals. This indicates that the sources of contamination in the case of surface water and groundwater are probably different. The nature of the sites with elevated levels of copper was not indicated. Experimental data demonstrate that leaching of copper is minimal. The copper concentration in some bodies of water evidently varies with season. In one small pond in Massachusetts, the concentration varied from Fe>Pb>Ni and the factors for Thessaloniki followed the order Cr>Zn>Fe>copper>Ni>Pb. Since all of these values were much higher than unity, the result demonstrated that effluent discharges have significantly impacted sediment quality near both cities. Sediment partitioning experiments indicated that copper was predominant near both cities in the organic phase (Athens, 51%; Thessaloniki, 61%), followed by the residual phase (i.e., non-exchangeable, non-carbonate, non-organic, non-reducible). Copper extraction from sediment was determined to be independent of acid concentration, with most copper extraction from sediment occurring within the first 10-15 minutes after contact with HCl. The aforementioned sediment studies were short-term in duration, lasting either days or months, and did not allow for temporal analysis of copper contamination. A year-long study was conducted through the Latvian Academy of Sciences in the Gulf of Riga to determine the impact of a waste treatment facility on the level of heavy metals in the area’s ecosystem (Seisuma et al., 1993). Water, sediment, plankton and mollusks were sampled near the sewage discharge area. Copper concentrations ranged between 0.3 and 8.8 mg/kg dry weight in the estuarine discharge area during sampling. The highest concentration of copper in water occurred at the central discharge point, but sediment concentrations were highest in the areas north and west of the central discharge location. Thus, it appears that waste waters washed contaminated sediment away from the central discharge location and transported metals, including copper, in a northerly and westerly direction. Ocean sampling in the area revealed copper concentrations ranging between 0.6 and 4.4 µg/l. Transport of metals did not appear to affect concentrations in the mollusk population studied, while uptake by plankton was directly related to metal concentrations in water but not sediment. An additional analysis of activated sludge from the treatment plant determined that healthy activated sludge actively accumulates heavy metals. Copper concentrations within the sludge ranged between 2052-5980 mg/kg dry weight. When the sludge is deactivated it not only fails to absorb more metals, but allows desorption to occur. To examine the impact of sewage discharge on sediment contamination along the North American continent, temporal trends of metal concentrations in sediment were studied near a wastewater outfall in California (Phillips & Hershelman, 1996) between 1985 and 1992. This outfall discharged approximately 260 million gallons per day (MGD) of treated municipal and industrial wastewater. While average effluent flow volume increased by 12% during this sampling period, the copper emission rate was determined to have been 84 kg/day in the 19851986 sampling year, but steadily declined to 31 kg/day during the 1991-1992 sampling year. Sampling of the sediment near the outfall found copper concentrations ranging from 20.5 to 51.6 mg/kg during the 1985-1986 sampling period and then decreasing to a range of 14.4 to 21.4

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mg/kg of sediment. When effluent copper concentration was normalized by suspended solids concentrations, it also showed a steady decline, with 1,860 mg copper/kg SS discharged between 1985-1986 and only 800 mg copper/kg SS discharged between 1991 and 1992. Sediment and effluent concentrations were found to be significantly correlated, indicating the impact of wastewater contamination on the sedimentary matrix. Another study which examined heavy metals accumulation in sediment was conducted to determine the impact of untreated sewage, industrial waste and landfill leachate discharge into Halifax Harbor, Nova Scotia (Buckley et al., 1995). Since the early twentieth century, copper concentrations have increased from 26 µg/g in the harbor sediment to 88 µg/g. A factor analysis of various geochemical data determined that sediment contamination from waste discharges was associated primarily with both total, and organic-bound forms of, lead, zinc and copper. Contamination attributed to leachate from landfills and from modification of contaminants by chemical reactions in the sediment have been attributed to acid-labile forms of zinc, nickel and copper. Characterization of the sediment found it to exist mostly of fine-grained particles. A study of heavy metals distribution in the Eastern Mediterranean (Ergin et al., 1996) suggested that aluminosilicates and organic matter are effective carriers of metals, facilitating transport to sediments. Copper concentration in surface sediment in the Gulf of Iskenderun ranged from 9 and 33 µg/g. However, in order to compensate for dilution caused by carbonates, total concentrations for various metals were recalculated on a carbonate-free basis. Upon recalculation, copper concentrations ranged between 14 and 52 µg/g, similar to those found in average crustal rocks (crustal rock concentrations not given). This study also noted that copper concentrations increased significantly with decreasing grain size in sediments composed of clay or mud. This is due to the larger surface area of the small particles and their strong adsorptive properties. The study also noted a strong correlation (r=0.84) between copper and magnesium concentrations. A significant correlation was also found between sediment copper and zinc concentration (r=0.87). This appeared to be due to uptake of these elements as micronutrients by microorganisms. The major source of copper in this Eastern Mediterranean region was primarily weathering of crustal rocks. A study supporting the hypothesis that transport caused copper concentrations to increase with distance offshore was conducted in Europe. Research performed in the Netherlands (Nolting et al., 1996) showed that copper concentrations increased from the Lena Delta seaward. Sediment copper concentration found in open waters off the coast varied between 2.1 and 20 µg/g, while in that same region, estuarine concentrations ranged between 26 and 219 µg/g, increasing seaward. The metals measured in the study (Ni, Cd, Pb, Zn) were found in highest concentration in river sediments, while offshore concentrations were not significantly higher than background (not given). Depth profiles of sediment cores did not reveal much variation in the concentration of any trace metal, including copper. While the aforementioned transport studies were of rather short duration, weeks to several months, a longitudinal study was conducted over a seven-month period off the coast of Sydney Australia. Its purpose was to examine the distribution of contaminants in sediment found in three

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spatially discrete zones from the shoreline (Schneider & Davey, 1995). Sampling was conducted from shore 2 km seaward, between 2 and 4 km from shore and, finally, beyond 4 km from shore. The turbulent waters in this region are storm-dominated with wave heights recorded up to 12 meters. Bottom currents were measured up to 1 m/s in the region, whereas surface currents could exceed a speed of 1.5 m/s. The lowest concentration of copper was measured in sediment 2 km from shore in waters up to 50 meters in depth. Similarly low levels were also measured in the sampling region 4 km offshore in waters that exceeded 80 m in depth. Highest concentrations of copper were measured in the 2-4 km region offshore in sediment at depths between 50 and 80 meters (actual concentrations not provided, only contour modeling profiles). A stepwise regression model was developed to isolate the anthropogenic contribution to sediment contamination from natural geochemical processes. The high copper levels in the sediment 2-4 km offshore were determined to originate from the harbor area, due to the fast-moving currents in the region. Harbor copper contribution to total offshore sediment copper was calculated at 16%; the remaining copper contamination was attributed to deep sea sewage outfalls from urbanized coastal areas. The previously discussed studies attempted to address the spatial distribution of copper contamination from both sewage and industrial discharges. While extremely important, this information does not adequately address the issue of copper accumulation throughout the depth of the sediment matrix. Thus, information is needed on the sedimentary profile of copper accumulation. For example, temporal and spatial variations in trace metals within bottom sediments were examined in the Sea of Japan (Tkalin et al., 1996). Sediment core analysis was performed in 2 cm increments from surface to 28 cm in two locations around the Sea of Japan. Analysis of the Amursky Bay sediment core found fairly uniform copper concentrations ranging between 5.9 and 8.3 ppm throughout the core sample. The average copper core concentration was 6.9±0.7 ppm. A Ussuriysky core sample (taken in Peter the Great Bay area of the Sea of Japan), while having a slightly higher copper concentration averaging 10.3±1.7 ppm, was found to also maintain a relatively uniform concentration throughout the 30 cm depth of the core sample. Fluxes for copper in the Amursky Bay and Ussuriysky Bay were calculated at 43 mg m-2 year-1 and 18 mg m-2 year-1, respectively. Sedimentation rates for trace metals were estimated for the Amursky Bay and Ussuriysky Bay as 0.17 g cm-2 year-1 and 0.12 g cm-2 year-1 respectively, indicating not much variation between the two bays. In contrast to other core studies, a Croatian study of trace metal accumulation in sediment (Mihelcic et al., 1996) found the highest concentrations of copper within the first 10 centimeters of sediment within a saltwater lake (11.3-13.2 µg/gm dry weight). Copper concentrations at depths between 15 and 40 centimeters of sediment were 5-6 times lower than surface concentrations. The transport and fate of mine tailings in sediments were studied in a coastal fjord in British Columbia (Odhiambo et al., 1996). Core samples collected from Alice Arm and Upper Observatory Inlet were analyzed for metals, including copper, to establish sediment accumulation rates. Accumulation rates of copper in Alice Arm ranged between 1.4 and 2.0 g/cm2 per year, whereas accumulation rates in the inlet varied between 0.17 and 0.76 g/cm2 per

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year. Thus, sediment, including metals from local mining operations, become trapped in the Arm after entering the waterway. Sediment copper concentrations near the mines, one located at the head of Alice Arm and another located 32 km from the head of Alice Arm on the banks of the Kitsault River, ranged between 56.3 and 70 µg/g; copper concentrations averaged 650 µg/g in the vicinity of a smelter located about 25 km from the head of Alice Arm. Crustal copper concentration in the mining area was determined to be 55 µg/g. Core samples taken outside of Alice Arm waterway in the Upper Observatory inlet in British Columbia, showed enrichment of copper, Cd, Pb and Zn in the sediment. Concentration of these metals decreased with distance from the mine. The best tracer for mine tailing dispersion was copper since, copper background levels being very low, even low levels in the sediment could be detected. A study to evaluate the effect of organic matter decomposition on diagenetic remobilization of copper was undertaken in the Kalix River estuary in northern Sweden (Widerlund, 1996). Diagenetic changes occur when certain substances in sediment undergo chemical reaction and are transformed into rock after being buried under subsequent sediment deposits. This estuarine basin was 10-15 meters deep and contained sediments which are deposited at a rate of approximately 1.0 cm yr-1. Early diagenetic copper demobilization was found to be controlled entirely by decomposition of a highly reactive organic matter fraction in the surface layer of the sediment. Core profiles taken from the estuary system at a depth up to 36 cm revealed copper concentrations between 5.7 and 42 nmol/gm. Sediment traps in the river measured copper concentrations between 456 and 740 nmol/gm. Flux calculations indicated that approximately 3 percent of all copper deposited into the sediment layer is released back into the water column throughout the year. A European study conducted in the Mediterranean (Scoullos et al., 1996) reported copper concentrations in sediment ranging between 18 and 52 µg/g. Contrary to the results of other studies mentioned earlier, metal contamination within core samples was seen at a surface sample; however little information was available on sediments and core sampling since the primary focus was on water sampling and water levels. Along the Asian continent, heavy metal fluxes in Thailand were estimated to approximate the movement of copper between the water column and sediments (Cheevaporn et al., 1995). Sedimentation rates for the Bang Pakong River estuary showed sediments to be accumulating at a rate of 0.19-0.28 gm cm-2 yr-1. Sedimentary fluxes (0.1-16.8 µg cm-2 yr-1) of copper and other metals studied were higher than the diffusive fluxes of these metals (0.01- 4.8 µg cm-2 yr-1). The mean concentration of copper in porewaters ranged from 22-100 ppb, and did not appear to decrease with sample depth. Metal concentrations in the porewaters, however, did appear to be higher than in waters overlying sediments. In this system, between 70% and 91% percent of diagenetic contribution to metal flux was attributed to copper. Thus, for copper, diagenesis contributes significantly to metal enrichment of surface sediments. An evaluation of surficial sediments from Lake Michigan and the Virginian province (coastal Virginia) was performed to assess proposed sediment quality criteria for copper and other metals (Leonard et al., 1996). The concentration of acid-volatile sulfide, simultaneously extracted

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metals, total metals and total organic carbon were measured in core samples taken from these areas. Ninety-one percent of collected sediments contained measurable quantities of acid volatile sulfide. More than seventy percent of the marine sediments had concentrations of acid-volatile sulfide exceeding the concentration of the simultaneously extracted total metals, indicating little bioavailable metal in the sediment. Freshwater samples contained more simultaneously extracted total metals than acid-volatile sulfide, indicating more bioavailable metal in the freshwater sediment region. Heavy metal removal (transport) was studied in the Changjiang estuary in China (Shen & Liu, 1995). The annual transport of copper from the Changjiang River to the East China Sea is approximately 8000 tons. Measurements of particulate and ion forms of copper indicated that concentrations of the ion form of heavy metals, including copper, were quite low (0.48 µg/L for copper) compared to the particle form (11.7 µg/L for copper). River water entering the East China Sea changes both the salinity and pH of the water and leads to coagulation of the ionic form and particulate form of the heavy metals. Very large particles result from coagulation and enter the sediment. Most metals (e.g., zinc, cadmium) are removed in the salinity range of 10 to 23. The maximum movement of this copper complex from water into the sediment was 22%. With changing salinity, the horizontal distribution of total copper indicated a 51% decrease from estuary to sea.

Fresh Water Sediments

Fresh water systems and copper sediment contamination is the last category to be examined in this chapter. Concentration profiles in this aquatic environment are often used to trace the history of metal accumulation by the sediment from anthropogenic sources via long-range hydrologic and atmospheric transport. As opposed to the marine environment, lake sediments are not as reliable in determining contamination trends since they are more subject to acidification which may cause demobilization of metals. One of the most extensively studied systems is Lake Erie, one of the lower Great Lakes of North America. Trace element profiles of the sediment to a depth of approximately 50 centimeters were recently obtained to determine relationships between sediment metal concentrations and the concentration in pore water of the Central Basin of Lake Erie (Azcue et al., 1996). The study determined that the lowest copper concentration (approximately 30 µg/g) in the sediment existed between 30 and 50 centimeters below the watersediment interface. The concentration of copper steadily increased from a depth of 30 centimeters through a depth of 18 centimeters below the sediment surface, coinciding with heavy anthropogenic inputs during the 1960s (approximately 75 µg/g). Concentrations of copper then decreased steadily through the column from 18 cm to the sediment surface (approximately 60 µg/g). Sediment concentrations of dissolved trace elements below the water-sediment interface are approximately one order of magnitude greater than concentrations measured in the lake water. Copper and other element concentrations measured within the first several centimeters of sediment were determined not to be representative of anthropogenic deposition since this upper layer is subject to a variety of biogenic processes and the effects of hydrologic transport.

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Heavy metals (Zn, As, Pb, copper and Cd) present in sediments were also studied in two tributaries of the Chesapeake Bay (Gupta & Karuppiah, 1996). This was done to determine the contributions of both point source and non-point source pollution to the estuary. Metal concentrations were found to be higher in water near a sewage treatment facility and a poultry farm than in sediment just downstream from a sewage treatment facility alone. Sediments in deeper water also had lower copper concentration than sediment in shallower water, presumably resulting from hydrologic transport and dilution. Metal accumulation of cadmium, chromium, copper, iron, nickel, lead and zinc in bottom sediments of wet ponds was investigated in south and central Florida (Yousef et al., 1996). Accumulation rates for certain heavy metals, specifically copper, correlated with ratios of pond surface area to drainage basin area. The mean copper accumulation rate for the 13 sampling sites was 1.3 kg/ha yr±1.3 kg/ha yr. The content of copper was determined as 23.5 µg metal/g dry sediment in the accumulated top sediments. Copper concentration declined exponentially with depth. An equation was developed for predicting copper accumulation rates in kg/ha yr as follows: copper=0.013*(DA/PA)1.08 + 0.35, where DA=total drainage area (ha) and PA=pond surface area (ha). Copper accumulation rates appear to level off when the PA:DA ratio ≥2%. An evaluation of copper, lead and zinc contained in inter-city canal sediments was conducted in the Netherlands (Bijlsma et al., 1996). The study was designed to characterize the sediment, which typically had a thickness of 20-30 cm, resting upon the sandy bottom of the canals. Copper concentration in the sediment top layer ranged between 45 and 281 mg/kg (mean 166 mg/kg) and ranged between 150 and 530 mg/kg (mean 350 mg/kg) in the sediment bottom layer. Mass balance calculations determined a copper accumulation of 164 kg/yr. While the previous fresh water studies were of relatively short duration, days to weeks, several studies of longer duration were conducted in the fresh water environment to establish temporal sediment trend of metals. The release of trace metals (iron, manganese, lead, copper cadmium zinc and chromium) from sediments was studied during a two-year period along a 110 km segment of the Seine River in France surrounding a sewage treatment facility (Garban et al., 1996). Copper concentrations in the sediment increased noticeably several kilometers downstream from a sewage treatment facility (8% in 1991 and 87% in 1992) and steadily decreased from approximately 24 kilometers downstream from the sewage treatment outlet through the end of the sampling area. Zinc concentrations, while consistently higher than copper, showed an overlapping concentration profile, suggesting similar aquatic behavior. Copper concentrations in the sediment were similar from year one (215 mg/kg) to year 2 (216 mg/kg). In addition, an increase in copper sediment concentration was accompanied by an increase in PAH and PCB concentrations. Seasonal variation in river metal contamination was studied in sediments near the discharge from an abandoned copper mine (Herr & Gray, 1996). Over the course of a one-year period, copper drainage from the mine averaged 11 kg/day. Significant positive correlations were found between copper and iron, suggesting co-precipitation of the two metals. Average copper sediment concentrations were lower in the spring (414.5±229 µg/gm) than in the summer

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(587±379 µg/gm), but gradually continued to increase over time. Mean surface copper decreased approximately 15% during the study period. Transport of copper, lead, zinc and cadmium was observed during flood events in an Alpine river in Germany (Hamm et al., 1996) and it was determined that flooding has a major impact on total copper transport. During both normal flood and high flood events during the sampling period, approximately one quarter of annual copper transport occurred. Copper concentrations in flood waters reached a maximum of 400 µg/l at the highest stage of flooding. Pollution status of the Indus River in Pakistan was examined through the study of fish, sediment and water heavy metal concentration (Ag, As, Cd, Cr, copper, Fe, Mn, Ni, Pb, Zn, Hg) (Tariq et al., 1996). This is one of the world’s largest rivers in terms of drainage basin, discharge and sediment load. Concentration analysis of copper and other heavy metals indicated an increase in concentration at locations downstream. Sediment concentrations for copper ranged between 2.6 and 33.2 µg/g. Fish provided a uniform distribution of metals along the sampling route, whereas this association could not be established for water or sediment. There was an inverse relationship between trace metal and nutrient content in the fish. There was a positive correlation between trace metal concentration in fish and sediment. Maximum macronutrient concentration in sediment was found where the trace metal concentrations in fish were not at maximum levels. It was also determined that trace metal concentrations followed a decreasing order as follows: sediment>fish>water. A statistical analysis was undertaken in the analysis of river sediments to identify areas of lead, copper, iron and chromium enrichment caused by anthropogenic activities by examination of concentration ratios of pairs of metals (Murray, 1996). Sediment contamination was identified by comparing both trace metal-to-conservative metal concentrations and absolute metal concentrations in the sediment along the Rouge River in Michigan. Comparison of the ratio of an element, such as copper, to another element of lower sediment variability, such as iron, can determine if metal enrichment has occurred. Since these low variability metals are always detected at higher initial concentrations, they are less affected by anthropogenic activities, while trace metals are more susceptible to changes from anthropogenic activity. The average copper concentration along the river as measured from 16 sampling locations was 153±16 ppm. The use of a copper-to-iron ratio reduced the effect of sediment grain size and organic content on metal concentrations and was seen as a conservative estimate of sediment contamination. Ratios can also be used to separate metal contamination due to natural and anthropogenic sources. A statistical approach was also developed to delineate sediment zones according to trace metal distributions (Poulton et al., 1996). Analysis included ratio matching, cluster analysis and principal component analysis. The use of ratio matching was employed to reduce the effect of sediment dilution by other, inert, materials. Sediment samples of similar origin were defined by ratio matching and cluster analysis. Average total copper in the sediments of Hamilton Harbor on Lake Ontario was 89±36.8 µg/gm. Extractable copper for the same sediment was measured to be 24.2±11 µg/gm. Zones were delineated as follows: Group 1 - center of harbor, deep water undergoing good mixing, fine sediment; Group 2 - located close to shore and had higher sand

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content and noticeably lower trace metal concentration, high wave action; Group 3 - similar to Group 2, but high cadmium concentrations, region experiencing longshore drifts; Group 4 located in a channel having abnormally high copper and other metal concentrations; Group 5 very close to shore with an extremely high sand content (99.5%) and very low metal contamination, extremely shallow water, large amount of mineral separation, high wave activity. Copper and trace metal contamination was directly related to the presence of clay and silt and inversely related to the presence of sand. It is not possible to provide a clear picture of copper accumulation in sediments found in the marine, estuarine or freshwater environments based upon the studies examined for the monograph from the Copper Source Book and other sources of information. Perhaps the most noticeable aspect of each sediment study reviewed in the document was that copper was seldom the major focus of a study. Its concentration in the sediment was usually reported without adequate discussion. Studies of sediment contamination primarily focus on elements, such as lead, arsenic, mercury and zinc, or examine the essential nutrient composition of sediment in terms of the requirements of a host of aquatic organisms. In these studies, a variety of methods were used to sample for heavy metal or trace metal contamination in the aquatic environment without any noticeable attempt at standardization of both sampling and analysis. For example, several of the papers focusing on the marine environment attempted to examine the impact of industrial and municipal waste outfalls on heavy metal contamination in water and sediment. However, no standardized method or consistent set of criteria for sediment sampling was used in the vicinity of outfalls. An attempt should be made to define zonal areas at discrete distances from a wastewater outfall in both seaward and down-current directions to test copper-specific hypotheses on flux, transport and deposition in sediment. This would enable researchers to conduct sampling at defined distances from outfalls and to better compare sediment contamination and contaminated sediment transport among numerous sampling locations near outfalls. Environmental parameters such as water current direction and flow rate, salinity and degree of turbidity should also be consistently reported in the literature when conducting these type of studies. Many of the studies do not mention the type of sediment under investigation, nor do they provide adequate detail on sediment particle size fractions. Core sampling in the aquatic environment needs a set of criteria, like those mentioned above, to allow for inter-comparison of results from different studies. A review of the literature focusing on temporal trends of heavy metal (e.g., lead) accumulation in sediments indicates an increase in contamination beginning in the late 19th century. Then, after the Second World War, a rapid increase in heavy metal contamination (e.g lead, arsenic, cadmium, mercury) was reported, reaching a plateau between the 1960s and 1970s when stricter regulations on waste discharges were implemented. These observations, however, must be carefully interpreted since there does not appear to be a standardized core sampling methodology to compare data obtained by different studies. Standardization is necessary to ensure that reported elevations or reductions in sediment metal concentrations (e.g., copper) are not artifacts of a poorly-developed sampling strategy. Several of the aforementioned studies only examined sediment in the first 2-5 cm from

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the surface and this data may be erroneously compared to studies using deeper core samples. Since turbulent water redistributes contaminated sediment over vast distances, it would be more appropriate to take deeper samples to determine the concentration of more permanent (sequestered) copper in the region, as opposed copper found in the superficial upper layer. Even when deeper core sampling was performed, there was no established standard depth for taking core samples. Some core sampling is performed between 2 and 10 centimeters below the surface, whereas other cores are taken to a depth of 50 cm. Appropriate comparison of core profiles can only be conducted if the profiles are of similar depth. Only when this standardization is implemented can sediment layering characteristics and depth of copper contamination be fully understood. Finally, while many papers state that levels of copper in the sediment do not significantly exceed background levels for that area or those levels determined in crustal samples, background values are frequently left unreported for that region. Sufficient research has not been conducted to determine aquatic background levels of copper in either pristine environments, or areas accepting urban and industrial effluents.

6.3 Copper Concentrations in Soils and Terrestrial Biota

Copper occurs naturally at levels of ~50 µg/g in the earth's crust, which includes soil and parent rock (Perwak et al., 1980). In agriculturally productive soils, copper ranges from 1-50 µg/g, while in soil derived from mineralized material, copper levels may be much higher (Perwak et al., 1980). Copper concentrations in soil samples collected throughout the United States yielded a geometric mean of 18 µg/g (Fuhrer, 1986). Samples were taken at a depth of 8 inches to avoid anthropogenic contamination; 2/3 of the samples contained copper concentrations between 8.0 and 40 µg/g. These copper levels are supported by a review of soil copper concentrations that reported a median concentration of 30 µg/g (dry weight) and a range of 2-250 µg/g (Davies & Bennett, 1985). Copper concentrations in soil may be much higher in the vicinity of a source. Concentrations in the top 5 cm of soil near the boundary of a secondary copper smelter were 2480 ± 585 µg/g (Davies & Bennett, 1985). Maximum wetland soil/sediment copper concentrations were 6912 µg/g in the immediate vicinity of a Sudbury, Ontario smelter, but the concentrations decreased logarithmically with increasing distance from the smelter (Taylor & Crowder, 1983a). Results suggest that copper in the soil from the study area was primarily associated with particulate emissions from the smelter. In a study in which the copper concentrations of 340 soil samples were presented in terms of land-use types, the average copper concentrations reported were 25 µg/g in agricultural land, 50 µg/g in suburban/residential land, 100 µg/g in mixed industrial/residential areas, and 175 µg/g in industrial/inner urban areas (Haines, 1984). From an analysis of the spatial distribution of the copper, it was concluded that most of the contamination was a result of airborne deposition from industrial sources. Soils from Lemhi, Twin Falls, and the Idaho National Engineering Laboratory in southern Idaho had geometric mean copper concentrations of 13.4-20.4 µg/g dry weight (Rope et al., 1988).

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Elevated concentrations of copper in soils and dust appears to be localized to areas of industrial activity, particularly in areas of copper smelting. The effect of land-use on copper concentrations in top soil was investigated in Richmond-upon-Thames (low industrial activity), and Wolverhampton (high industrial activity), in the U.K. (Kelly et al., 1996). The concentrations of copper in Wolverhampton were approximately 2-fold higher than the concentrations in soil from areas with similar land uses. However, there were order of magnitude differences in residential soil concentrations within each of the two areas, probably due to the practice of disposing of fireplace ash in gardens. The deposition rate of copper in the vicinity of a copper refinery in Prescott, UK before and after installation of emissions controls was reported in a study of copper concentrations and ecotoxicity (Turner et al., 1993). Copper deposition appears largely localized to areas in the immediate vicinity of the plant, where the deposition rate (averaged from the years 1970-1990) was more than 150-fold greater than urban background in the UK. At a distance of 3 km from the site, the deposition rate dropped to only two-fold UK urban background. The high rate of copper deposition resulted in elevated concentrations of copper in soil, and although there was considerable spatial variability in copper concentrations (281 - 2106 ppm in 1991), largely due to variation in tree canopies, copper concentrations were generally elevated down to a depth of 1 m. Copper concentrations in soil near a copper refinery, in which emissions controls were installed in 1979, were found to decrease over the period 1988 to 1991 from an average concentration of 2363 ppm to 1522 ppm (Turner et al., 1993). The high concentration of copper in soils near the site resulted in growth reduction in sycamore trees, as evidenced by tree ring widths, compared to a reference site 3 km away. Subsequent reduction in soil copper concentrations were accompanied by reduction in extent of growth inhibition at the site. Several other studies demonstrate the localized effect of industrial activity on copper concentrations in soil. Copper concentrations in the top soil were studied from the Kastela Bay region of Croatia (Milos et al., 1993). This is a densely populated 120 sq km coastal region, containing industrial and urban areas. Copper concentrations ranged from 10 m distant. The influence of treated dock piling wood appears to be localized

COPPER: Environmental Dynamics and Human Exposure Issues Page 105


to a radius of less than 10 m, and even within this area, elevated concentrations appear to have no effect on oysters. The ability of microalgae from the Sea of Japan to monitor heavy metal concentrations in marine waters was investigated (Zolotukhina et al., 1993). Two microalgae species, chosen as biomonitors on the basis of laboratory studies, indicated that copper concentrations at three sampling stations increased over the period 1989-91. Copper and other trace element accumulation in arctic marine ecosystems was investigated by examining tissue levels in three species of birds at Hornoya, Norway (Wenzel & Gabrielsen, 1995). The average copper concentration in bird feathers, liver, kidney, and muscle were similar for all three species, as well as in juvenile birds of one of the species, despite differences in the feeding habits of these birds. The results indicate that copper concentrations are well regulated in birds living in the arctic. Seasonal variations in copper and other heavy metal concentrations in bivalves in the Sea of Japan, near Vladivostok was investigated (Shulkin & Kavun, 1995). Copper concentrations increased between February and April, and subsequently decreased as temperatures increased. Increases in copper concentration were attributed to increased filtration, and the subsequent decrease was attributed to increased metabolic activity and physiological regulation concomitant with increased ambient temperature.


As a part of the National Contaminant Biomonitoring Program of the U.S. Fish and Wildlife Program, eight species of freshwater fish were collected at 112 stations in the United States in 1978-1979 and 1980-1981 (Lowe et al., 1985). The geometric mean concentrations of copper in µg/g (wet weight) for the two years were 0.86 and 0.68; the 85th percentiles were 1.14 and 0.90 and the ranges were 0.29-38.75 and 0.25-24.10, respectively. The highest concentration, 38.75 and 24.10 µg/g, during both collecting periods was in white perch from the Susquehanna River and the second highest concentration, 19.3 µg/g, was found in white perch from the Delaware River near Trenton, NJ. Copper residues in muscle of 268 fish specimens were analyzed over a five-year period in several surface water systems in eastern Tennessee (Blevins & Pancorbo, 1986). The mean residue levels in the muscle of different species of fish from nine stations ranged from 0.12-0.86 µg/g (wet weight). Maximum levels ranged from 0.14-2.2 µg/g. Mean and median copper concentrations of 127 samples of edible fish from Chesapeake Bay and its tributaries were 1.66 and 0.36 µg/g in 1978, and 1.85 and 0.61 µg/g in 1979 (Eisenberg & Topping, 1986). Copper levels were increased in the livers and to a lesser degree, the gonads, compared with the flesh. The copper content of muscle tissue of several species of fish collected from metal-contaminated lakes near Sudbury, Ontario, ranged from 0.5-1.4 µg/g (dry weight). No major pattern in variation was evident among species or among the study lakes (Bradley &

COPPER: Environmental Dynamics and Human Exposure Issues Page 106


Morris, 1986). The copper concentration in the livers, however, ranged from 5-185 µg/g (dry weight) and differed significantly among species and among lakes. Unlike muscle tissue, liver tissue is a good indicator of copper availability, although the data indicate that there are other factor(s) influencing copper availability and bioaccumulation in these fish.

Aquatic Invertebrates

Copper concentration in the soft tissue of mussels and oysters collected as part of the U.S. Mussel Watch Program in 1976-78 ranged from 4-10 µg/g (dry weight) for mussels and 25-600 µg/g for oysters (Goldberg, 1986). Copper concentrations in mussels collected from eleven sites near Monterey Bay, CA, were 4.63-8.93 µg/g (dry weight) (Martin & Castle, 1984). Perwack et al. (Perwak et al., 1980) reported similar results for mussels (3.9-8.5 µg/g) and clams (8.4-171 µg/g).


Although copper concentrations in plants vary widely, they usually range from 1-50 ppm (dry weight) (Davies & Bennett, 1985; Perwak et al., 1980). Adding lime to the soil to increase pH to 7 or 8 reduces copper availability to plants. 6.4.1

Sewage Sludge

In an EPA-sponsored study to determine metal concentration in sewage sludge, copper concentrations in primary sludge at seven POTWs were reported to be 3.0-77.4 µg/g, with a median concentration of 20.5 µg/g. The plant with the highest copper concentrations received wastes from plating industries, foundries, and coking plants. In a comprehensive survey of heavy metals in sewage sludge, 30 sludges from 23 American cities were analyzed (Mumma et al., 1984). Copper concentration in the sludges ranged from 126-7729 µg/g (dry weight), with a median value of 991 µg/g. The proposed limit for copper in sludge spread on agricultural land is 1000 µg/g (Mumma et al., 1984). For comparison, the concentration of copper in cow's manure is ~5 µg/g (Mumma et al., 1984). Additional information for copper levels in sludge can be found in a recent ICA report (Landner et al., 2000).

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7 ASSESSING HUMAN EXPOSURE, DOSE AND RISK 7.1 Exposure Potential and Pathways

In discussing exposure to copper, the important question is whether individuals are exposed to readily available copper, which in general, means free (hydrated) copper(II), and perhaps some weakly complexed or adsorbed forms of copper. Available data indicate in general that copper in natural water, sediment, and soil is not in a labile form. Potential for high exposure of the general population to copper may exist where people consume large amounts of tap water that has picked up copper from the distribution system. This is most likely to occur where the water is soft and is not allowed to run and flush out the system. In such cases, copper concentrations frequently exceeds 1 mg/l, a large fraction of which may be free cupric ion, resulting in exposure by ingestion and dermal contact. Based on available data, copper at National Priorities List (NPL) sites is not expected to be hazardous to people living close to the sites or to clean-up workers. Data suggest that the copper at these sites would not in general be in a labile form and should not leach into groundwater. People living near copper smelters and refineries and workers in these and other industries may be exposed to high levels of copper in dust by inhalation and ingestion. In some industries, workers may be exposed to fumes or very fine dust that may be more hazardous than more typical dust. A key issue in exposure and health risk assessment is the need to evaluate exposure/intake from all possible sources, routes and pathways. For example, the main route of exposure to copper is oral, and food and water, the predominant sources. Inhalation exposure from polluted air may occur in the workplace, especially in mining and agricultural work where copper salts are used as pesticides. The effect of airborne copper may be of particular importance because of direct effects on the lung. However, for most practical purposes, the oral route is the only one of health significance in terms of general population exposures to copper. Food may account for over 90% of copper intake in adults where water has low copper content (< 0. 1 mg/l). If water copper content is high, 1-2 mg/l, it may account for close to 50% of total intake. In the case of infants consuming copper-supplemented artificial formula, the contribution of water may be less than 10%, whereas, if the formula is not fortified with copper, water may contribute over 50% of total copper intake, especially when water copper content is in the high range of 1-2 mg/l.

7.2 Environmental Exposures 7.2.1


Everyone is exposed to copper in atmospheric dust. Estimates of atmospheric copper concentrations from representative source categories yielded a maximum annual concentration of 30 mg/m3 (USEPA, 1987; ATSDR, 1990). If a person is assumed to inhale 20 m3 of air/day, this would amount to an average daily intake of 600 mg of copper. For the reported range of annual

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atmospheric copper concentrations, 5-200 ng/m3, the average daily intake by inhalation would range from 0.1-4.0 mg. At a maximum reported ambient air concentration, 100 mg/m3 for a 24hour period at a location within one-half mile of a major source, the average daily intake would rise to 2000 mg. However, these estimates assume that all the copper is attached to particles of inhalable size, which is usually not the case.


The mean copper content of cigarette tobacco was 24.7 ppm, with a standard deviation of 10.8 ppm (Mussalo-Rauhamaa et al., 1986). However, only 0.2% of this copper passes into mainstream smoke. 7.2.2


Ingestion exposures are dominated by copper in drinking water and food and are considered in more detail under a following section. The average daily dietary intake of copper from food is 2 mg/day. It is less likely that high dermal exposures will result from bathing in this tap water because the distribution system will flush itself out as the water is drawn. Airborne dust containing copper may also contribute to ingestion exposures for people living near copper smelters and refineries and for workers in these and other industries. 7.2.3


A less likely situation where exposure to high levels of free copper(II) may occur is from swimming in water that has been recently treated with a copper-containing algicide. In natural waters, the level of free copper would be expected to decrease rapidly. Soluble cupric salts are used extensively in agriculture and in water treatment. Workers engaged in the production and application of these chemicals, as well as industrial workers, such as those in the plating industry, may come into dermal contact with these copper-containing compounds.

7.3 Occupational Exposures

A National Occupational Exposure Survey (NOES) conducted by NIOSH from 1981- 1983 estimated that potentially 505,982 workers, including 42,557 women, are occupationally exposed to copper in the United States. Interpreting the significance of the NOES is particularly difficult for copper because we are all exposed to copper; furthermore, the survey does not indicate the level or form of copper to which the worker may be exposed. The NOES estimate is provisional because all of the data for trade name products that may contain copper have not been analyzed. Of the potential exposures, 1073 are to pure copper, while in the other cases, the molecular form of copper was unspecified. Additionally, according to the NOES, 125,045 workers, including COPPER: Environmental Dynamics and Human Exposure Issues Page 110


38,075 women, are potentially exposed to copper sulfate. The NOES was based on field surveys of 4490 facilities and was designed as a nationwide survey based on a statistical sample of virtually all workplace environments in the United States where eight or more persons are employed in all standard industrial codes (SIC), except mining and agriculture. The exclusion of mining and agriculture is significant for estimating exposure to copper because of their high potential for exposure. Studies of workers that include quantitative measurements of ambient and personal exposure (i.e., breathing zone) have been performed on U.S. copper smelter employees. In a 1995 study (reported in Nriagu, 1979b), Wagner described conditions of exposure in fourteen U.S. copper smelter facilities that were the subject of industrial hygiene surveys conducted by NIOSH. Measurements of airborne concentrations of As, Pb, Zn, Cd, Mo, copper, and sulfur dioxide revealed that only the last two occurred consistently at relatively high levels. For copper, the industry-wide average for both area and personal samples exceeded 1 mg/l for dust and fumes in several plant locations. The current federal standard for occupational exposure to copper restricts in-plant exposure to concentration of 1 mg/m3 for dust and 0.1 mg/m3 for fume emissions. However, it has been found that most of the element typically occurred as a “non-respirable” dust, and that this dust consists mainly of a relatively inert copper sulfide. Urinary levels of copper were determined in the smelter worker studies, but Wagner indicates that methods of collection and analysis were not well controlled for potentially confounding factors. These reservations notwithstanding, a mean urinary copper content of 79 µg/l determined from 206 individual samples was well above the 9 to 18 µg range suggested as normal for 24-hr urinary excretion in human beings. Nonetheless, evidence of copper toxicity, acute or chronic, was not revealed in the smelter survey despite excessive ambient, personal, and urinary levels of copper. Perhaps this relates to the absence in these facilities of free metallic copper and copper oxide, the chemical forms most often associated with the so-called “metal fume fever” of copper and other types of metal exposures. It should be noted that arsenic contaminants in copper ore have been considered a likely cause of the increased risk in pulmonary carcinoma in smelter worker populations and in non-occupationally exposed residents of at least one copper-mining city. In general, in epidemiologic studies, as in the U.S. copper smelter studies, health hazards related to this metal have been found to be of low order. However, over time various questions have been raised regarding unrecognized, occupationally related, copper-induced illness. Of considerable importance in this regard was the discovery of serious to fatal pulmonary and hepatic abnormalities in a group of Portuguese vineyard sprayers exposed to a copper sulfate-containing fungicide for 3 to 45 years. Such evidence concerning long-term copper toxicity could imply that it would be prudent to reserve conclusions about the health-hazard implications of prolonged exposure until critical longitudinal studies of worker health become available.

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7.4 Dietary Exposures to Copper: Drinking Water and Food 7.4.1

Copper in Food

Estimates of dietary intake by Americans prior to 1970 were considerably higher than current intake estimates. This however mostly reflects marked improvements in analytical techniques for measuring copper and awareness of the importance of avoiding copper contamination of analytical samples. The usual diet was thought to contain 2 to 5 mg of copper, but studies, including one study of 132 diet composites, now show that few diets contain over 2 mg per day. As with all nutrients, copper intake can vary widely, depending on food choices. Diets in countries where more whole-grain products, legumes, and organ meats are eaten, contain more copper (Turnlund, 1999). The amount of copper in the daily diet sufficient for health is in the range of 20-60 µg/Cu/kg/day for most children and adults. The needs of newborn infants are 40-150 µg/Cu/kg/day; premature babies may need 150-400 µg/Cu/kg/day and malnourished children may need over 500 µg/Cu/kg/day for recovery. The question of how much copper is too much is more difficult to answer. The upper safe limit of around 200 µg/Cu/kg/day is safe for most adults and children, but formula-fed children may exceed this and the needs of preterm and marasmic infants are much higher (Ralph & McArdle, 2001). The richest sources of dietary copper contain from 0.3 to over 2 mg/100 g (50 to >300 nmol/g). These include shellfish, nuts, seeds (including cocoa powder), legumes, the bran and germ portions of grains, liver, and organ meats. Most grain products; most products containing chocolate, fruits and vegetables, such as dried fruits, mushrooms, tomatoes, bananas, grapes, and potatoes; and most meats have intermediate amounts of copper, from 0.1 to 0.3 mg/100 g (20-50 nmol/g). Other fruits and vegetables, chicken, many fish, and dairy products contain relatively low concentrations (1000 ppb (160) (450), 24% of samples >1000 ppb ó5-530 [ó5]

at tap, standing water

Maessen et al., 1985

running water standing water

Maessen et al., 1985 Maessen et al., 1985

raw water

Meranger et al., 1979

ó5-100 [ó5]

treated water

Meranger et al., 1979

ó5-220 [20]

distributed water

Meranger et al., 1979

ó5-80 [ó5]

raw water

Meranger et al., 1979

ó5-100 [ó5]

treated water

Meranger et al., 1979

ó5-560 [40]

distributed water

Meranger et al., 1979

ó5-110 [ó5]

raw water

Meranger et al., 1979

ó5-70 [ó5]

treated water

Meranger et al., 1979

10-260 [75]

distributed water

Meranger et al., 1979

New Jersey


1063 samples, 90th percentile 64.0 ppb, maximum 2783 ppb, groundwater may or may not be used for drinking water

United States

(4.2) [4.0]

53,862 occurrences

New Jersey


Surface, marine

E. Arctic Ocean


Surface, marine

Atlantic Ocean


Pond Lakes

Massachusetts Canada