Metal Biogeochemistry in 

Surface-Water Systems 

A Review of 

Principles and Concepts

U.S. GEOLOGICAL SURVEY CIRCULAR 1013
Cover. Symbols of the elements discussed in this report are shown. The triangular 

background represents three-phase partitioning of the elements in aquatic systems the 

soluble phase, the suspended sediment phase, and the bottom sediment phase.
Metal Biogeochemistry in 

Surface-Water Systems 

A Review of 

Principles and Concepts

By JOHN F. ELDER

U.S. GEOLOGICAL SURVEY CIRULAR 1013
DEPARTMENT OF THE INTERIOR 

DONALD PAUL MODEL, Secretary

U.S. GEOLOGICAL SURVEY 

Dallas L. Peck, Director

UNITED STATES GOVERNMENT PRINTING OFFICE : 1988

Free on application to the Books and Open-File Reports Section, 

U.S. Geological Survey, Federal Center, Box 25425, Denver, CO 80225

Library of Congress Cataloging in Publication Data

Elder, John F.

Metal biogeochemistry in surface-water systems.

(U.S. Geological Survey circular; 1013)

Bibliography: p.

Supt. of Docs, no.: I 19.4/2:1013

1. Metals Environmental aspects. 2. Trace elements in

water Environmental aspects. 3. Biogeochemistry.

I. Title. II. Series. 

QH545.M45E43 1988 574.5'263 87-600257
CONTENTS

Abstract 1 

Introduction 1

Purpose and scope 2 

Terminology and elements included 2 

Acknowledgments 3

Sources of metals in the aquatic environment 3 

Atmospheric deposition 3 

Industrial and municipal point sources 5 

Drainage from mining areas 5 

Nonpoint sources 6 

Other sources 7

Occurrence and distribution of metals in aquatic systems 7 

In water 7 

In sediments 7

Partitioning of metals among different phases 8 

General characteristics of partitioning 8 

Analysis of metal species 9 

Factors that control metal partitioning 11 

Hydrogen-ion activity (pH) 11

Effect of pH on solubility 11 

Effect of pH on adsorption and desorption 12 

Effect of pH on organometallic interactions 12 

Effect of pH on biological uptake 13 

Implications of acidic atmospheric deposition 14 

Oxidizing or reducing conditions 14 

Combined Eh-pH effect 15 

Other effects of redox changes 15 

Adsorption to hydrous oxides of iron and manganese 16 

Concentration of organic matter 17

Effect of organic complexation on solubility of metals 17 

Other factors affecting attainment of equilibrium 17 

Effect of organic complexation on adsorption of metals 18 

Effect of organics on other reactions incorporating metals 19 

Overall effect on metal speciation 19 

Biological activity 19

Methylation and other bacterially mediated

transformations 19

Algal activity extracellular sequestering agents 19 

Enhancement of iron-manganese oxide formation 20 

Other biological effects on partitioning 20 

Hydraulic energy flow and turbulence 21 

Riverine hydrologic effects 21 

Lacustrine hydrologic effects 21 

Nature of substrate material 22

Contents III
Sediment grain size 23

General effects of grain size 23 

Grain-size effects in natural waters 23 

Bimodal distribution 24 

Effects of grain size on metal transport 25 

Temperature, light, salinity, and water hardness 25 

Effects of metals on aquatic biota 26

Bioavailability and bioaccumulation 26 

Multielement effects 26

Nonuniform bioaccumulation among and within individuals 27 

Effects of speciation on bioaccumulation 28 

Toxicity 28

Use of biota as pollution indicators 28 

Effects of metals on phytoplankton 29 

Toxicity as a function of bioaccumulation 29 

Effects of metal interactions 30 

Transport of trace metals in rivers and lakes 30 

Forms of transport in rivers 30 

Transport in lakes 31 

Overview 31 

References cited 33

FIGURES

1. Map of Schuylkill River, Pa., main stem and graph showing relation of 

chromium concentrations to river mile in bed sediments of the main stem 

of the Schuylkill River 4

2. Diagram showing metal partitioning in aquatic systems among three major 

phases and some environmental conditions that favor each phase 9

3. Activity ratio diagram showing speciation of Cu(II) as a function of pH 11

4. Plots of adsorption of cadmium on amorphous ferric hydroxide, 

as a function of pH and concentration of total iron 12

5. Graph showing adsorption of various metals on humic acids, 

as a function of pH 13

6. Graph showing removal of dissolved copper and lead by Nostoc muscorum, 

as a function of pH 14

7. Stability field diagram for solid and dissolved forms of iron, as a function of

Eh and pH as 25° C and 1 atmosphere pressure 15 

8-10. Graphs showing:

8. Charge variation on surfaces of some common suspended solids, 

as a function of pH 17

9. Free cupric-ion activity as a function of total copper concentrations 

and concentration of total iron, in cultures of Anabaena cylindrica 20

10. Poly chlorinated biphenyl concentration in water, as a function 

of discharge in the Hudson River, N.Y. 22

11. Diagram showing iron cycle in a lake having an anoxic hypolimnion, 

showing various transport processes 23

IV Metal Biogeochemistry in Surface-Water Systems
12-15. Graphs showing:

12. Total surface area of spherical particles whose total mass equals 

1,000 grams and density equals 1 gram per cubic centimeter, 

as a function of the number of particles 24

13. Nickel concentrations as a function of surface area in bottom sediments 

of the Ottawa and Rideau Rivers near Ottawa, Canada 25

14. Concentration of zinc in sediments of the Saddle River, N.J., 

as a function of grain size 26

15. Metal bioaccumulation by an organism,

as a function of weight of the organism 28

TABLES

Principal elements discussed in this report and their abundances in the

Earth's crust 2

Published reports of metal concentrations in sediments in surface

waters in North America 8

Summary of three sequential partial extraction procedures for

separating metal fractions in particulate material 10

METRIC CONVERSION FACTORS

For readers who wish to convert from the metric system of units to the inch-pound system of units, the conversion factors are listed below:

Multiply By To obtain

meter (m) 3.281

kilometer (km) 0.6214

millimeter (mm) 0.03937

micrometer (/mi) 3.937 x 10 ~ 5

gram (g) 0.03527

kilogram (kg) 2.205

megagram (Mg) 1.102

milligram (mg) 0.00003527

microgram (/tg) 3.527 X 10~ 8

square centimeter (cm2) 0.1550

square meter per gram (m2/g) 305.0

cubic meter per second (m3 /s) 35.31

gram per cubic centimeter 6.243 x 10~ s

(g/cm3)

milligram per liter (mg/L) 1.0

microgram per liter (/tg/L) 1.0

microgram per gram (/tg/g) 1.0

milligram per kilogram (mg/kg) 1.0

microgram per kilogram 0*g/kg) 1.0

degree Celsius (° C) 1.8x(° C + 32)

foot (ft)

mile (mi)

inch (in.)

inch (in.)

ounce (oz)

pound (Ib)

ton, short (t)

ounce (oz)

ounce (oz)

square inch (in2)

square foot per ounce (ft2/oz)

cubic foot per second (ft3 /s)

pound per cubic foot

(lb/ft3)

parts per million (ppm)

parts per billion (ppb)

parts per million (ppm)

parts per million (ppm)

parts per billion (ppb)

degree Fahrenheit (° F)

Conversion Factors
Metal Biogeochemistry in Surface-Water Systems 

A Review of Principles and Concepts

By John F. Elder

Abstract

Metals are ubiquitous in natural surface-water 

systems, both as dissolved constituents and as particulate 

constituents. Although concentrations of many metals are 

generally very low (hence the common term "trace metals"), 

their effects on the water quality and the biota of surfacewater systems are likely to be substantial. Biogeochemical 

partitioning of metals results in a diversity of forms, including hydrated or "free" ions, colloids, precipitates, adsorbed phases, and various coordination complexes with 

dissolved organic and inorganic ligands.

Much research has been dedicated to answering 

questions about the complexities of metal behavior and effects in aquatic systems. Voluminous literature on the subject has been produced. This paper synthesizes the findings 

of aquatic metal studies and describes some general concepts that emerge from such a synthesis. Emphasis is on 

sources, occurrence, partitioning, transport, and biological 

interactions of metals in freshwater systems of North 

America. Biological interactions, in this case, refer to 

bioavailability, effects of metals on ecological characteristics 

and functions of aquatic systems, and roles of biota in controlling metal partitioning.

This discussion is devoted primarily to the elements 

aluminum, arsenic, cadmium, chromium, copper, iron, lead, 

manganese, mercury, nickel, and zinc and secondarily to 

cobalt, molybdenum, selenium, silver, and vanadium. 

Sources of these elements are both natural and anthropogenic. Significant anthropogenic sources are atmospheric 

deposition, discharges of municipal and industrial wastes, 

mine drainage, and urban and agricultural runoff.

Biogeochemical partitioning of metals is controlled 

by various characteristics of the water and sediments in 

which the metals are found. Among the most important controlling factors are pH, oxidation-reduction potential, 

hydrologic features, sediment grain size, and the existence 

and nature of clay minerals, organic matter, and hydrous 

oxides of manganese and iron. Partitioning is also controlled 

by biological processes that provide mechanisms for 

detoxification of metals and for enhanced uptake of nutritive 

metals. Partitioning is important largely because availability to biota is highly variable among different phases. Hence,

accumulation in biological tissues and toxicity of an element 

are dependent not only on total concentration of the element but also on the factors that control partitioning.

INTRODUCTION

During the past several decades, there has been a 

growing awareness of widespread and often serious environmental problems that have resulted from pollution by 

metals. Problems posed by elevated concentrations of toxic 

metals in the aquatic environment include hazards to human 

health and varying degrees of toxicity to most plant and 

animal species. Some species may be so severely affected 

that they are eliminated from contaminated ecosystems. Food 

web structures of entire communities thus may be significantly altered. There also is the possibility of accumulation 

of metals in biological tissues (bioaccumulation) through 

various levels of the food chain and of latent effects resulting 

from gradual bioaccumulations.

Stimulated by this awareness of the metal contamination problem, a great deal of research in the water resources 

area has been directed at investigations of metals in aquatic 

systems. Most of the research has dealt with metals of known 

or suspected toxicity, although there has also been a considerable amount of interest in the nutritive value of metals. 

Some elements that are toxic at high concentrations may have 

nutritive value for aquatic biota in unpolluted systems containing very low concentrations of those elements.

Aquatic metal investigations are numerous and 

diverse. The diversity of research sites, objectives, and study 

designs that have been applied in these investigations has produced an almost equal variety in results. Therefore, it is difficult to review aquatic trace metal literature and to extract 

general rules that meaningfully apply to a variety of aquatic 

systems. Perhaps the most unquestionable generality is that 

the biogeochemistry of metals in natural water systems is 

extremely complex. A knowledge of trace metal chemistry 

is, in fact, only a beginning to acquiring some understanding

Introduction 1
of the complexity of trace metals in natural waters. Biological 

and physical processes exert an important influence over the 

forms in which the elements occur and the pathways that lead 

to those forms. It has become apparent that biota not only 

are affected by the presence of toxic metals but also have 

a great deal of control over the concentrations and toxicities 

of those same metals in the aquatic environment.

Considerably more research is needed to improve our 

understanding of the complexities of aquatic metal biogeochemistry. As in any scientific pursuit, however, the investigators undertaking such studies need to be familiar with 

what has already been done. The primary purpose of this 

report is to help fill that need.

Purpose and Scope

This report provides a synthesis of the current (1986) 

state of our understanding of metals in surface-water systems. 

It is intended to aid the reader in identifying important concepts and principles that have emerged from the diverse collection of published research in this area. It also identifies 

and discusses some findings that are controversial those 

where the results or conclusions of some studies differ from 

others and offers possible explanations for the apparent 

inconsistencies.

Recent publications have already provided many syntheses of aquatic metal biogeochemistry research. Notable 

among these are books by Jarvis (1983), Forstner and Wittmann (1981), and Salomons and Forstner (1984). These 

monographs provide a comprehensive coverage of examples 

that document principles developed over the past quarter century of research. Shorter papers, such as that of Komarovskiy and Polishchuk (1981), provide information on more 

specific topics. Thorough discussions of general water 

chemistry concepts, including principles of trace metal 

behavior in natural waters, were written by Stumm and 

Morgan (1981) and Hem (1985). Kuwabara and Helliker 

(1988) discussed various concepts of trace contaminant 

transport in streams and provided extensive reference citations. Leland and Kuwabara (1986) produced a recent summary of current knowledge of toxicology of trace metals. 

Review papers published annually in the Journal of Water 

Pollution Control Federation are good sources of 

bibliographic information. Three papers in the 1985 review 

issue (Lathrop and Davis, 1985; Pritchard, 1985; Roush and 

others, 1985) contained numerous citations of literature 

dealing with metals in surface waters.

This paper is not intended to provide a complete 

bibliography of trace metal research, which would only 

duplicate other readily available reviews. Instead, I hope that 

this discussion will complement other publications that provide thorough annotated bibliographies. The focus is on concepts rather than specific studies. Of course, the concepts 

have evolved through analysis and reanalysis of diverse study 

results, and pertinent references are cited in the subsequent 

discussion of these concepts.

Nearly all the publications referenced in this paper 

appeared in readily available technical journals, most of them 

within two decades prior to this writing. Recent studies, 

although not necessarily more significant than earlier studies, 

are more likely to characterize the current status of 

knowledge. Reports having limited distribution were 

generally avoided unless judged especially significant.

Emphasis in this paper is on freshwater systems, both 

lotic and lentic. Considerably more work has been done on 

marine and estuarine systems than on freshwater environments. While studies of marine and estuarine systems 

are by no means ignored in this review, reference to them 

is limited to cases whose results contribute substantially to 

our understanding of metals in freshwater, as well as in 

marine systems.

Terminology and Elements Included

The terms "trace metals," "toxic metals," "trace 

elements," "heavy metals," and simply "metals" are often 

used interchangeably. I acknowledge that such usage is rather 

unfortunate because it leads us to ignore the slight distinctions 

among all of the sets of elements that these terms represent. 

However, maintaining total accuracy in terminology becomes 

cumbersome, especially because some of the elements under 

study, such as arsenic and selenium, are nonmetals. For 

simplicity, the term "metals" will be used in this review 

to include a group of elements that are of greatest interest 

and concern in aquatic systems. A list of these elements and 

their abundances in the Earth's upper crust is given in table 

1. Some other elements not on the list are considered occasionally in this report, primarily for comparison with 

elements of primary or secondary focus.

Table 1. Principal elements discussed in this report and 

their abundances in the Earth's crust

Abundance in 

Earth's upper crust 

Element (micrograms per gram)

Aluminum. ..............

Arsenic .................

Cadmium ...............

Chromium ..............

Cobalt ..................

Copper .................

Iron ....................

Lead ...................

Manganese ..............

Mercury ................

Molybdenum .............

Nickel ..................

Selenium ................

Silver ..................

Vanadium ...............

Zinc ....................

....... >84,700

....... 22.5

....... 2 .14

....... >35

....... '10

....... >25

....... >35,000

....... >15

....... >600

....... 2 .11

....... 23.8

....... >20

....... 2 12

....... 2.07

....... '60

...... '52

 Taylor and McLennan, 1981. 

2Turekian and Wedepohl, 1961.

2 Metal Biogeochemistry in Surface-Water Systems
Acknowledgments

Special contributions from a number of people helped 

to make this report possible.

Consultation with various researchers helped to 

establish the directions and goals of the study. Consultants 

from outside the U.S. Geological Survey included Dr. Wim 

Salomons of the Delft Hydraulics Laboratory in the 

Netherlands, Dr. Anders W. Andren of the University of 

Wisconsin, Dr. Thomas L. Crisman of the University of 

Florida, Dr. Patrick L. Brezonik of the University of Minnesota, and Dr. William H. Funk of Washington State 

University. U.S. Geological Survey personnel who contributed useful suggestions and reference information included K.E. Bencala, Edward Callender, J.A. Davis, G.L. 

Feder, V.C. Kennedy, J.S. Kuwabara, H.V. Leland, S.N. 

Luoma, G.E. Mallard, D.M. McKnight, and J.M. McNeal. 

A.J. Horowitz and J.S. Kuwabara provided critical review 

of the manuscript.

Special acknowledgment is also given to J.J. Collins 

of the U.S. Geological Survey in Tallahassee, Fla., for 

assistance in collection and preparation of much of the 

background literature used for this report.

SOURCES OF METALS IN THE 

AQUATIC ENVIRONMENT

Metals are common in the Earth's crust, as well as 

in products and wastes of human activity. Hence, the sources 

of metals in the aquatic environment are both natural and 

anthropogenic.

Aluminum and iron are among the most common 

elements in the Earth's crust (table 1). Erosion of materials 

from the watershed substrate supplies large quantities of these 

elements as background levels in natural waters. In the case 

of aluminum, anthropogenic sources are generally assumed 

to supply insignificant quantities of the metal, relative to 

amounts that are already in the system due to natural processes. This characteristic makes aluminum a useful reference 

metal in assessing the degree of pollution of other elements 

whose sources are primarily anthropogenic (Bruland and 

others, 1974; Forstner and Wittmann, 1981; Thomas and 

others, 1984). Concentrations of elements having anthropogenic sources may be normalized to concentrations of 

aluminum to identify changes over time due to new sources. 

Considerable care must be used in applying this procedure, 

however, especially when there are pH changes that may 

mobilize significant amounts of aluminum. Titanium, which 

also may be used as a reference metal (Forstner and Wittmann, 1981), has the advantage of being less subject to 

mobilization under low pH conditions.

This discussion is devoted primarily to the elements 

aluminum, arsenic, cadmium, chromium, copper, iron, lead, 

manganese, mercury, nickel, and zinc and secondarily to 

cobalt, molybdenum, selenium, silver, and vanadium. 

Sources of these elements are both natural and anthropogenic.

Metals of anthropogenic origin may be transmitted 

to surface waters through various pathways, including atmospheric deposition, localized industrial and municipal 

discharges (point sources) of waste waters or byproducts, 

drainage from mining areas (which may be either point or 

nonpoint sources), and urban and agricultural surface-water 

runoff (nonpoint sources). Documentation of some of these 

types of input is limited, but for other types, it is quite 

abundant.

Long-term records of contaminant inputs to lakes are 

commonly contained in sediment cores (Feltz, 1980). A 

number of investigators have taken advantage of this occurrence to describe trends of metal inputs to lakes (Nriagu and 

others, 1979; Dillon and Evans, 1982; Davis and others, 

1983; Evans and others, 1983; Mudroch, 1983; Ochsenbien 

and others, 1983; Ouellet and Jones, 1983; Thomas, 1983; 

Thomas and others, 1984). Typically, the cores contain concentration peaks at or near the surface in recent sediments. 

This condition is nearly always interpreted as a reflection 

of increasing anthropogenic inputs in recent decades.

Atmospheric Deposition

Numerous investigations have shown that atmospheric deposition is an important source of metals to lakes. 

Most notable of the airborne contaminants is lead, due to 

its widespread discharge into the atmosphere from exhausts 

of gasoline engines. The atmospheric pathway is most likely 

to be significant for large, shallow lakes and for reservoirs 

near urban areas. Three factors contribute to the importance 

of atmospheric input in such systems: (1) proximity to 

sources of atmospheric pollutants, (2) relatively high ratios 

of lake-surface area to lake volume, and (3) relatively low 

ratios of watershed area to lake-surface area.

Sedimentary records from lakes in southern Ontario, 

Canada, were used by Dillon and Evans (1982) to examine 

the importance of various sources of lead in the lakes. The 

elevated lead levels in the upper layers of the sediments of 

all of the lakes indicated the effects of anthropogenic lead 

inputs in recent years. However, little variation among lakes 

was found, regardless of factors such as size of watershed, 

ratio of watershed to lake-surface area, or lake residence 

time. It was concluded that the uniform input rate could be 

explained only if the primary entry pathway was atmospheric. 

Where spatial differences did occur, they were attributed to 

redistribution of sediments. Later work by the same group 

(Evans and others, 1983) gave similar results for zinc and 

cadmium.

The importance of atmospheric deposition as an input pathway for lead, zinc, and cadmium was corroborated 

in a similar study by Thomas and others (1984). Enrichment 

factors for metals were calculated by dividing surface sediment concentrations by background and normalizing to 

aluminum concentrations. Very little enrichment of copper 

was found. Arsenic enrichment was attributed to agricultural 

use in the basin. However, substantial enrichments of lead,

Sources of Metals in the Aquatic Environment 3
zinc, and cadmium were observed, and these enrichments 

approximately corresponded to atmospheric fluxes previously 

determined for the area. Atmospheric flux of mercury has 

been shown to be significant in coastal waters (Windom and 

others, 1975), and atmospheric transport may also play an 

important role in redistribution of mercury to inland waters.

Lakes that are in geologically similar watersheds and 

in the same general vicinity are not necessarily similar in 

atmospheric metal inputs. The differences may be due to local 

anthropogenic activities. These effects may be sorted out by 

correcting for bedrock (background) contributions to the 

metal loading (Davis and Galloway, 1981).

The Great Lakes are subject to considerable influence 

from atmospheric deposition. Because of their size, their 

watershed area to lake-surface area ratios are relatively low,

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Figure 1. Chromium concentrations in river bed sediments 

of the Schuylkill River, Pa. A, Location of the main stem of 

the Schuylkill River, fi, Relation of bed-sediment chromium 

concentrations to river mile (from Stamer and others, 1985, 

p. A20).

4 Metal Biogeochemistry in Surface-Water Systems

and approximately 50 percent of the water entering the lakes 

falls as precipitation directly on their surfaces. Schmidt and 

Andren (1984) examined the results of numerous studies that 

produced atmospheric loading estimates for metals in the 

Great Lakes and found that there was a great deal of 

discrepancy. Such discrepancy may have been partially attributable to real spatial and temporal differences, but 

whatever the cause, it precluded the possibility of precise 

determinations of atmospheric deposition of the metals. 

However, rough estimates did allow the authors to conclude 

that more than half of the zinc and 90 percent of the lead 

entered the lakes through atmospheric pathways.

One of the studies on the Great Lakes (Sievering and 

others, 1981) indicated the importance of dry deposition as 

a source of metals to the lakes. At least 40 percent of the 

total lead input, 20 percent of the total zinc input, and 15 

percent of the total iron input were attributed to dry deposition alone. These results suggest some urgency for further 

examination of dry deposition and wet deposition separately. 

Studies of dry deposition have been hampered by difficulties 

in analyzing and interpreting dry deposition data (Schmidt 

and Andren, 1984).

Atmospheric deposition may cause enrichment at the 

surface microlayer of natural waters (Elzerman, 1981; Barnes 

and others, 1982). Dissolved and paniculate matter may accumulate in the surface microlayer [0.01- to 1.0-mm 

(millimeter) thickness] at higher concentrations than in the 

adjacent bulk water. All material directly deposited on water 

bodies from the atmosphere must obviously contact this layer. 

In doing so, the material may be retained there, at least temporarily causing surface microlayer enrichment. Surface 

enrichment of metals, often retained by surface tension or 

organic slicks, has been documented in seawater (Piotrowicz 

and others, 1972; Hunter, 1980; Hardy and others, 1985), 

but little is known about surface enrichment in freshwater 

systems. Metal enrichment in this unique environment 

(Hardy, 1982) may induce effects that are distinct from those 

in the bulk solution.

Other investigations in various parts of the country 

have shown quite consistently the importance of atmospheric 

deposition as a source of metals in lakes, especially as a 

source of lead. Rainwater in Florida contained elevated concentrations of lead near the urban areas of the southern part 

of the State (Irwin and Kirkland, 1980). An oligotrophic lake 

in northern Wisconsin had appreciable lead accumulation in 

its sediments, mostly attributable to atmospheric inputs 

(Talbot and Andren, 1984). In a subalpine catchment in the 

northern Cascade Range in Washington, more than 50 percent of lead, copper, arsenic, and other trace metals could 

be accounted for by atmospheric deposition (Dethier, 1979).

The widespread occurrence of lead emissions in the 

atmosphere near aquatic systems is an important consideration in selecting sampling sites for monitoring studies. Care 

should be taken to avoid locating sampling sites near highway 

bridges or sites of airborne emissions from industry. Unlike
discharges directly into the streamflow, atmospheric transport 

may affect metal concentrations both upstream and downstream of the source.

Industrial and Municipal Point Sources

Because anthropogenic sources account for many occurrences of elevated metal concentrations in natural waters, 

investigations of the biogeochemistry of metals are often 

carried out in systems where such sources are common. 

However, very few of these studies have actually identified 

the sources of the metals. The studies tend to be oriented 

toward descriptions of behavior of the elements once they 

are in the system, rather than toward identification of their 

sources. In the few cases where source monitoring has been 

conducted, results show quite consistently that, for both water 

and sediments, concentrations are appreciably higher near 

the point source and they diminish rapidly as distance from 

the source increases.

The most extensive documentation of industrial and 

municipal point-source contamination has been from studies 

of the Rhine River in Europe (Salomons and Forstner, 1984). 

Metal "mapping" in the sediments showed marked increases 

in cadmium, lead, zinc, mercury, copper, and chromium in 

a downstream direction. Cadmium exhibited the most 

dramatic increases, reaching concentrations more than 100 

times natural background values. Virtually all of these increases were associated with industrial activity along the 

river.

Stamer and others (1985) found elevated concentrations of chromium and other metals in the sediments of the 

Schuylkill River, Pa., just downstream of effluent discharges. 

Figure 1 shows the relation of chromium concentrations to 

river mile. 1 Reading, at river mile 76, is an industrial area 

that seems to be the major site of chromium input to the river. 

Metal concentrations gradually decreased with distance 

farther downstream. Profiles for other metals, particularly 

copper and nickel, were similar, although their concentrations upstream of Reading were higher than those of 

chromium. Both natural and anthropogenic sources of copper 

and nickel appeared to be significant.

Wiener and others (1984) reported that Upper 

Mississippi River concentrations of several metals, particularly cadmium, lead, and mercury, tended to be highest 

at or near the Twin Cities metropolitan area. However, 

various other inputs in the system were indicated by scattered 

concentration peaks both upstream and downstream of the 

Twin Cities.

A number of investigations in the Great Lakes indicated that industrial and municipal effluents are important 

sources of metal inflow to the lakes. Examination of mercury concentrations in Lake Ontario sediments by Thomas 

(1972) showed increased accumulation relating to industrial 

pollution that dated back to the turn of the century. Although

'River mile = distance of 1 mile, measured upstream from the river 

mouth, along a line located in the established course of the stream.

there was a tendency for greater concentrations to occur in 

fine-grained sediments near the center of the lake, high concentrations [up to 2,000 j«,g/kg (micrograms per kilogram)] 

were found in nearshore areas. These nearshore concentrations indicate that the principal inflow was from the Niagara 

River. A more generalized study (Fitchko and Hutchinson, 

1975) provided data for 10 metals from the mouths of 116 

tributaries to the Great Lakes system. Most of the tributaries 

that drained urban-industrial areas had elevated metal concentrations, while most of the relatively undisturbed 

tributaries did not. Also, concentrations tended to be higher 

in tributaries of the lower Great Lakes (Erie, Ontario, and 

Michigan) than in the upper Great Lakes (Superior and 

Huron), correlating to proximity of urban industrial activity. 

Nriagu and others (1979) calculated annual inputs and sediment retention of cadmium, copper, lead, and zinc in Lake 

Erie. The sedimentary records showed evidence of high 

anthropogenic inputs. The authors pointed out that sewage 

discharges, in particular, were an important source of metals 

to the lake.

In a study of Chesapeake Bay, Helz (1976) found that 

the relative importance of natural and anthropogenic sources 

varied among different metals and among different parts of 

the bay. Waste water was the primary source of cadmium, 

and atmospheric deposition was the primary source of lead, 

regardless of location in the bay. However, for many other 

metals, including manganese, iron, cobalt, nickel, and zinc, 

industrial discharges were the primary contributors in nearshore areas, and natural sources were more important for 

the bay as a whole. Because only part of the input was 

removed by sedimentation, anthropogenic inputs were judged 

to have significant input on water quality.

In a large city, one might expect to find substantial 

impact of anthropogenic inputs. Evaluating the sources of 

metals in New York City wastewaters, which ultimately 

reach New York Harbor, Klein and others (1974) reported 

that nonindustrial inputs, especially wastewater from residential areas, are more important than industrial discharges for 

metals such as copper, zinc, and cadmium. For nickel and 

chromium, the electroplating industry was the major contributor, but residential wastewater was still very significant. 

These findings complicate the problem of how to regulate 

sources of contamination. Obviously, regulation of one 

perceived point-source (in this case, the electroplating industry) might reduce total pollution but would not eliminate 

the problem.

Drainage from Mining Areas

Acid mine drainage is a source of extremely high concentrations of metals at some sites around the country. Not 

only do the mines constitute a rich source of metals, but also 

the oxidation of sulfides in pyrite and in other ores releases 

sulfate, producing low pH in drainage waters (Wentz, 1974). 

The acidity tends to solubilize the metals and to greatly increase their mobility (Moran and Wentz, 1974).

Sources of Metals in the Aquatic Environment 5
The Coeur d'Alene River in Idaho, the south fork 

of which receives drainage from a mining district, is an 

example of a river that is severely impacted by mine drainage 

(Funk and others, 1975). Dissolved metal concentrations 

range as high as 23 mg/L (milligrams per liter) zinc, 0.35 

mg/L cadmium, and 2.8 mg/L lead. Complete inhibition of 

growth of the green alga Selenastrum capricornutum occurs 

at 0.12 mg/L zinc and 0.080 mg/L cadmium (Bartlett and 

others, 1974). As expected, no populations of 5. capricornutum were found in the waters of the Coeur d'Alene River.

Another North American system that has been 

severely affected by mining is the Belle Fourche River and 

its tributary Whitewood Creek near Lead, S. Dak. Between 

1880 and 1977, about 100 million Mg (megagrams) of mine 

tailings containing arsenopyrite and other metallic sulfides 

were discharged into the creek (Marron, 1987). Arsenic concentrations in stream sediments are as high as 5,400 ppm 

(parts per million) 25 km (kilometers) downstream of the 

town of Lead. The contaminated flood-plain sediments will 

remain for centuries as a source of metals to adjacent streams. 

A study of this system is being conducted by the U.S. 

Geological Survey.

A lead-zinc-copper mine that discharges tailings into 

a British Columbia lake has caused substantial increases in 

metal concentrations in the lake during the mine's 18 years 

of operation (Austin and others, 1985). Changes in water 

chemistry have affected community structure of periphyton 

and phytoplankton. Metal-sensitive species have dramatically 

decreased in abundance and have been replaced by more 

tolerant species. Species diversity also has decreased.

Mining activity may alter speciation and transport of 

metals in drainage streams, not only by releasing metals 

directly into the drainage waters, but also by causing other 

perturbations that affect the metals. Modification of pH is 

one important example of such perturbations. Others are increased sedimentation due to construction and land disturbance and increased input of organic material (Smith and 

others, 1983). Cabrera and others (1984) observed that 

simultaneous release of organics and metals from old mine 

spoil heaps resulted in high metal concentrations downstream.

The effects of coal mining on water quality were examined by Wirries and McDonnell (1983). Data from inactive coal mine drainages in eight States showed that concentrations of aluminum, manganese, nickel, and zinc were 

found to exceed Federal effluent quality guidelines in 40 to 

60 percent of the samples collected. High concentrations of 

a number of metals in streams that drain coal mining areas 

tend to be associated with sulfate concentrations greater than 

2,000 mg/L (Toler, 1982).

Drainage waters from areas where specific precious 

metals are mined are likely to contain high concentrations 

of various other metals. Tin and tungsten mines in Tasmania 

produced severe pollution of two drainage streams (Tyler 

and Buckney, 1973). Contaminants, including sulfuric acid, 

cadmium, copper, iron, lead, and manganese, were at such

high levels that the waters were unusable for domestic or 

agricultural use. In South Africa, gold and uranium mining 

activity produced enrichment of lead, cadmium, mercury, 

arsenic, and selenium (Wittman and Forstner, 1977). Mining 

of barium can release several toxic metals, principally lead 

and zinc (Smith and others, 1983).

Nonpoint Sources

In areas remote from urban-industrial development, 

nonpoint sources are usually the predominant contributors 

of metals in aquatic systems (McElroy and others, 1975). 

Surface-water runoff represents a major component of the 

water budget of many surface-water systems, especially those 

without channelized tributaries (Winter, 1981). It is thus a 

major potential contributor to contaminant loading, especially 

in agricultural or silvicultural areas. Nutrient loading from 

point sources has been demonstrated and quite well 

documented, but similar studies of metal loading from nonatmospheric, nonpoint sources have been very limited.

One of the major nonpoint sources for water bodies 

near urban development is storm water runoff from manmade impervious surfaces. Studies that show elevated contaminant concentrations near densely populated areas lead 

to the conclusion that urban storm water runoff is an important source of the contaminant (Ellaway and others, 1982; 

Renwick and Edenborn, 1983). Runoff from basins in south 

Florida that included a highway and commercial development was compared to atmospheric deposition by Miller and 

Mattraw (1982). Runoff loads of zinc were more than half 

of the atmospheric loads. Runoff loads of lead equaled or 

even exceeded those from atmospheric deposition.

Erosion has been a major contributor of nutrients and 

metals to Lake Erie (Kemp and others, 1976). Large quantities of fine-grained sediments were released by erosion and 

could have been transported long distances in the lake (up 

to 150 km). Surface enrichment factors of cadmium, copper, 

mercury, lead, and zinc in the lake sediments were significant and reflect the importance of anthropogenic inputs that 

reached the lake largely by way of erosion.

Another area where erosion is the primary contributor 

of nutrients and metals is the Lake Tahoe basin (Leonard 

and others, 1979). The biological impact of such inflow in 

this system is eutrophication rather than accumulation of toxic 

levels of contaminants. Earlier work had shown that primary 

productivity in the lake was limited not only by macronutrients but also by some metals, including iron and zinc, 

that may be transported in large quantities to the lake during 

snowmelt runoff (Goldman, 1964).

The Mississippi River drains more than 40 percent 

of the land area of the contiguous United States. Analyses 

of its sediments thus provide indicators of broad regional contamination. Trefry and others (1985) used sediment analyses 

from the Mississippi River Delta to determine trends in lead 

contamination and effects of regulations limiting lead additives in gasoline. Some sediment core data, taken from near

6 Metal Biogeochemistry in Surface-Water Systems
the river mouth, showed peak lead concentrations in sediments deposited in approximately 1970, corresponding quite 

closely with the maximum annual consumption of lead in 

gasoline. The results implied that lead in gasoline had been 

an important source of lead pollution on a nationwide scale 

and that reduction of this source was quite effective in reducing lead inputs to the environment.

Other Sources

Ground-water flow is likely to be an important 

hydrologic control on contiguous surface-water systems 

(Winter, 1983). However, its role in supplying metals to 

those same systems is sharply limited because most metals 

are readily sorbed to soils and aquifer materials. Even 

relatively mobile metallic elements, such as sodium, 

potassium, and lithium, are sorbed to sedimentary materials 

(Bencala and others, 1983, 1984). However, there are appreciable differences in distribution coefficients, depending 

on the metal and substrate. Removal of the metals during 

ground-water transport generally results in negligible metal 

input by this pathway.

Some work has been done to investigate the importance of volcanoes as a source of metals in aquatic 

systems. Kitano and others (1983) described metal input and 

speciation in rivers near Mount Ontake, Japan, following its 

1979 eruption. Volcanic material contained various metals 

incorporated in sulfides and silicates. Within the river 

sediments, diagenesis yielded transformation to various other 

forms. The 1980 eruption of Mount St. Helens did not produce appreciable increases in dissolved concentrations of 

most metals in the Columbia River estuary (Riedel and 

others, 1984). Of seven metals measured, only manganese 

and copper showed slightly elevated concentrations due to 

leaching from the volcanic material.

OCCURRENCE AND DISTRIBUTION OF 

METALS IN AQUATIC SYSTEMS 

In Water

Reconnaissance studies that describe occurrence of 

trace metals over regions are generally limited to areas no 

larger than States or small groups of States. On a national 

scale, the most complete work of this nature was done by 

Durum and associates (Durum and Haffty, 1963; Durum and 

others, 1971). The 1971 report gives data for 720 samples 

collected in 1970 from sites throughout all 50 States and 

Puerto Rico. All samples were filtered, acidified, and 

analyzed for arsenic, cadmium, hexavalent chromium, 

cobalt, lead, mercury, and zinc. The filtration, through a 

0.45-/im (micrometer) membrane filter, creates a commonly 

used operational definition of the dissolved fraction.

Data on the occurrence of metals in surface and 

ground waters are available for each State and territory of 

the United States in annual reports published by the U.S. 

Geological Survey. Each of these reports is available as a

U.S. Geological Survey Water-Data Report XX-nn (where 

"XX" is the State abbreviation and "nn" is the year). Waterquality and physical data from more than 14,000 sites around 

the country are included in these reports. Most of the trace 

metals are reported as their concentrations, in micrograms 

per liter, in the dissolved form. In some cases, results of 

analyses of "total recoverable" and "total in bottom 

material" are also included.

In Sediments

No complete nationwide reconnaissance of metals in 

sediments of freshwater systems has been reported. Local 

studies that give results of monitoring in North America are 

available, but comparisons among them are not practical 

because of the differences in sampling and analytical techniques. One of the most significant problems in interstudy 

comparisons is the effect of grain size. In some studies, only 

selected size fractions are collected, whereas most studies 

do not include any size separation at all, and results are given 

as total concentrations.

Table 2 lists studies that included determinations of 

metal concentrations in sediments. Analytical procedures 

varied; hence, some data represent total metal concentrations, 

whereas other data represent certain extractable fractions. 

The studies listed are limited to those that were conducted 

in freshwater systems or estuaries in the United States and 

Canada. The information in the table indicates the locality 

of each study, the metals monitored, and whether or not 

grain-size fractions were separated.

Although nationwide monitoring information on 

stream sediments is not available, such information is 

available for soils and other surficial material in a report by 

Shacklette and Boerngen (1984). Analytical data were presented for 47 elements analyzed in substrate samples from 

more than 1,300 sites uniformly distributed throughout the 

conterminous United States. For most elements, analyses 

were performed by emission spectrography and yielded total 

concentration data. A separate map for each element was produced and was accompanied by frequency distribution data.

The soil data showed some patterns of geographical 

distribution of elements on a nationwide scale. Aluminum 

and the alkali-alkaline earth metals are considerably more 

abundant in the western part of the country than in the eastern 

areas. Soils of the Pacific Northwest have relatively high concentrations of aluminum, cobalt, iron, scandium, and 

vanadium, whereas in the Rocky Mountain region, copper, 

lead, and zinc are relatively abundant. Mercury concentrations tend to be high in soils of the Gulf Coast and the upper 

Atlantic Coast. Among individual States, Florida had the 

most striking geochemical pattern, showing low concentrations of most of the elements included in the study.

Additional soils survey work by the U.S. Geological 

Survey included an extensive nationwide reconnaissance entitled National Uranium Resources Evaluation (NURE). ApOccurrence and Distribution of Metals in Aquatic Systems
Table 2. Published reports of metal concentrations in sediments in surface waters of North America

[If any attempt was made to separate grain-size fractions, it is so indicated. Studies whose primary focus was on processes governing metal concentrations,

rather than on the concentrations themselves, are not included]

Elements monitored

Location

State 

or Country Al As Cd Co Cr Cu Fe Pb Mn Hg Ni Zn Others

Grain-size 

separations Reference Date

Lakes Wis.

Lake George N.Y.

Great Lakes tributaries _______

Reservoirs Okla.

Lake Erie _______

Streams Maine

Narragansett Bay R.I.

Willamette River Oreg.

Chesapeake Bay Md.

Bayou Chico estuary Fla.

Iowa River Iowa

Hudson River N.Y.

Lake Michigan Mich.

Lakes Canada

Lake Erie _______

Potomac River Va.

Great Bear Lake Canada

Ontario lakes Canada

Ontario lakes Canada

Rivers Del.

Eastern North America lakes 

Saw Mill River N.Y.

Lake Ontario _______

Indian River Fla.

Fontana Lake N.C.

Apalachicola River Fla.

Chesapeake Bay _______

Schuylkill River Pa.

Waterways La.

 X - X - X - X - X - X - Iskandar and Keeney 1974

 X - X - X - X - X - Schoettle and Friedman 1974

 X-X-X-X- X-X-X-X-X- Ag Fitchko and Hutchinson 1975

 X - X - X - Pita and Hyne 1975

X X X X X X X X Si, K, Na, Mg Kemp and others 1976

 X X X X X X Chork 1977

X X X X X X X X XX V Goldberg and others 1977

 X X X X X X X Rickert and others 1977

X X X X X X X X X X Ag, V Goldberg and others 1978

 x X X X X X Rb, Ti, Zr Pilotte and others 1978

 x X Ba, Rb, Sr, Zr X Tsai and others 1978

 X X X X X X Williams and others 1978

 XXX X X X Filipek and Owen 1979

 X X X X X Jackson 1979

 X X X X Nriagu and others 1979

 XX X X X X X X X X Feltz 1980

 X X X X X X X Ra, Th Moore and Sutherland 1981

 X Dillon and Evans 1982

 X X Evans and others 1983

 X X X X X X X Hoffman and Biggs 1983

X X X X Ouellet and Jones 1983

 XXXXXXXXXX X X Rogers 1983

 X X X X X X X Thomas 1983

 X Trefry and others 1983

X X X X X X Mg Abernathy and others 1984

XXXXXXXXXXXX Ag, Mo, Sn, Si X Elder and Mattmraw 1984

 X X X X Di Giulio and Scanlon 1985

 X X X X X X X Be X Stamer and others 1985

X X X X X X X White and Tittlebaum 1985

proximately one million samples (at least one from every 

2-degree quadrangle) were analyzed. Uranium was the 

primary target element in the study, but information on a 

number of other elements was also produced. NURE reports 

are available at the U.S. Geological Survey library in Reston, 

Va.

PARTITIONING OF METALS AMONG

DIFFERENT PHASES

General Characteristics of Partitioning

Partitioning is defined as the process by which any 

solute, including contaminants, may transfer among the different phases and biological components of a surface-water 

system. Transfer pathways are nearly always bidirectional; 

hence, a substance that resides in any particular phase rarely 

stays permanently in that phase and may move to a different 

phase as conditions change.

Figure 2 illustrates diagrammatically the general 

nature of partitioning of trace metals among different phases 

in aquatic systems. Three broad abiotic phase categories are 

depicted at the apexes of the "triangle." Biological components (plankton, nekton, and benthos) are also involved 

in metal partitioning and are shown as components in the 

phase diagram.

Some of the conditions that influence the directions 

and magnitudes of metal transfers among phases are shown 

on the perimeter of the partitioning triangle (fig. 2). The soluble phase is favored under conditions of low pH, low Eh 

(reducing environment), low particulate loads (both organic 

and inorganic), and high concentrations of dissolved organic 

matter. If pH and particulate loads increase, while hydraulic 

energy of the system is high enough to sustain suspension 

of the particles, partitioning is likely to shift toward the 

suspended phase. Finally, high pH and Eh, combined with 

elevated concentrations of organic matter in the bottom 

sediments, are conditions that favor partitioning in the bottom sediment phase.

In most surface-water systems, biota are abundant and 

diverse and constitute important partitioning components 

themselves. High productivity among planktonic (freefloating) and nektonic (swimming) organisms usually implies 

uptake of contaminants by these organisms from both sol8 Metal Biogeochemistry in Surface-Water Systems
uble and suspended phases. An abundant benthic community is likely to incorporate and accumulate contaminants from 

water or ingestion of sediments or planktonic organisms.

Metals in the soluble phase tend to approach an 

equilibrium condition where all dissolved forms coexist in 

accordance with the dissociation constant of each metalligand association. Such a steady state is rarely reached in 

natural waters, however, due to the continually fluctuating 

nature of controlling factors such as pH, redox potential, and 

temperature.

One form of the metal within the dissolved phase is 

the "free ion," so named because it is not associated with 

dissolved ligands. However, it should not be considered a 

"bare metal ion" (for example, Cu2+ ), because such an ion 

cannot exist as a separate entity in water (Manahan, 1972, 

p. 19). Instead, the free ion is a hydrated form having the

absorption, or assimilation into cells. For many metals, 

biological uptake by algae involves initial adsorption followed 

by actual cellular uptake or absorption (Crist and others, 

1982). Absorption is accompanied by discharge of the metals 

in waste products of the same organisms, again resulting in 

continual input and output.

Metals within the bottom-material paniculate phase 

are also associated with inorganic particles, organic particles, 

and biota. The benthic fauna and aquatic vegetation that are 

involved are different species than those which are in overlying waters, but the mechanisms of incorporation are virtually the same, involving both adsorption and absorption. 

The inorganic and organic particles include many of those 

that may be in the suspended phase, but, because of their 

density, the particles settle into the bottom material.

general formula M(H2O)/ + , where M= metal and « = its Analysis of Metal Species

ionic charge. Other dissolved species of the metal are those 

involving association with one or more ligands, either inorganic or organic, which form soluble complexes with the 

metal.

Within the suspended particulate phase, metals may 

be associated with inorganic particles, organic particles, or 

biota. The dominant mechanism of such association is adsorption, or linkage to the surfaces of solid particles due to 

charge attractions between the surfaces and the dissolved 

ions. Adsorption reactions are readily reversible, and the 

system thus approaches, but rarely reaches, an equilibrium 

state involving adsorbed and desorbed forms. For biota, 

another dominant mechanism of incorporation of metals is

Most research having to do with metals in aquatic 

systems addresses one or more facets of the partitioning 

"triangle" (fig. 2). A considerable amount of attention has 

been centered directly on rates and significance of various 

pathways of interaction among different metal species. Any 

aquatic trace metal study, even if not process oriented, must 

involve examination of the results, if not the mechanisms, 

of partitioning. At the very least, separate analyses are done 

for water and bottom material, although the "water" phase 

usually contains some fraction of the suspended material. In 

addition, many studies require focus on one or more species. 

This places special demands on methodology for sample 

treatment and analyses.

High pH

High particulate

organic matter

concentration 

High suspended

sediment load 

High hydraulic energy

SUSPENDED

SEDIMENT

PHASE

BOTTOM

SEDIMENT

PHASE

Low pH

Low Eh

Low particulate 

organic matter 

concentration

Low suspended 

sediment load

High dissolved 

organic matter 

concentration

High pH 

Hjgh Eh 

High concentration

of organic matter

in sediments 

High benthos diversity

Figure 2. Metal partitioning in aquatic systems among three major phases (inside 

the triangle) and some environmental conditions that favor each phase (outside 

the triangle). Biological components (plankton, nekton, and benthos) are positioned 

according to the phases with which they are most closely associated.

Partitioning of Metals Among Different Phases 9
There has been a considerable amount of work done 

during the past three decades to develop sequential partial 

extraction procedures that allow reliable measurement of 

speciation (or fraction analysis) of a metal. Each step of a 

sequential partial extraction scheme produces an extract that 

theoretically contains only one metal fraction from the 

original sample, and analyses may then be performed on each 

fraction separately. A thorough history of the development 

of this methodology was provided in discussions by Chao 

(1984),Horowitz (1985),and Horowitz (in press).

Although most sequential partial extraction procedures that have been described in the recent literature are 

similar, there is no standardized procedure. In general, most 

procedures are constructed on the premise that the metal fractions of primary importance in natural water systems are as 

follows:

1. the dissolved fraction,

2. the exchangeable fraction, only loosely bound to colloidal 

or particulate material,

3. the carbonate fraction, associated with sediment carbonates,

4. the iron and manganese oxide fraction, adsorbed to surfaces of iron-manganese hydrous oxide particles or 

coatings,

5. the organic fraction, bound to various forms of organic 

matter, and

6. the crystalline fraction, held within the crystal structure 

of minerals, not readily removed and not available to 

biota.

Table 3 summarizes three partial extraction procedures for separating metal fractions in sediments. The approaches of Tessier and others (1979) and Popp and Laquer 

(1980) are similar, the principal difference being that the 

latter does not include steps that separate out the carbonate 

and organic fractions. The alkaline acetate treatment is used 

in both procedures to extract the exchangeable fraction, and 

the citrate-dithionite combination is used in both to extract 

the metals bound to hydrous iron-manganese oxides. The use 

of hydrogen peroxide and dilute nitric acid to extract the 

organically bound fraction was omitted by Popp and Laquer 

(1980) because the river (the Rio Grande) where their investigation was carried out carries a very small organic fraction in the sediment phase. However, the hydrogen-peroxide 

extraction step has been used quite commonly in other studies 

(Gupta and Chen, 1975; Kitano and others, 1980).

The approach of Kitano and others (1980) reflects 

a somewhat different fractionation scheme (table 3). Dilute 

hydrochloric acid, rather than acetate, is used in the first step, 

giving a slightly more vigorous attack that extracts ionexchangeable metals, some sulfides, and most carbonates. 

Hence, there is no true separation of exchangeable and 

carbonate-bound fractions. The remaining sulfides are exTable 3. Summary of three sequential partial extraction procedures for separating metal fractions in particulate material 

[Extraction steps proceed from top to bottom, the residue from each step being used for the subsequent treatment. For each method, table 

shows reagents used, concentrations of those reagents, temperature (in ° C),and pH at which extraction is conducted. For more details, 

consult source material. RT=room temperature; heat = direct heat application; M=molar; Abnormal; cone. = concentrated]

References

Tessier and others, 1979 Popp and Laquer

Fraction removed 

in extract

Bound to 

carbonates.

Bound to 

Fe/Mn oxides.

Reagents

!Vlg\_l 2

or 

NaOAc 

NaOAc 

+ HOAc 

Na2S 2O4 

+ Na citrate 

+ H citrateor 

NH2OH.HC1 

+ HOAc 

Concentration

1 M

1 M 

1 M 

cone.

.3 M - 

.175M- 

.025 M-

.04 M 

-25%

Temper- Concenature pH Reagents tration

 RT 1 NH4OAc 1 M 

 RT 8.2 

 RT 5 

96 Na,S2O4 .115M 

 + Na citrate .094 M

96

, 1980 Kitano and others, 1980

Temper- Concen- Temperature Reagents tration ature

 __ HC1 0. 1 N RT 

Bound to 

organic matter.

Crystalline 

HNO3 .02 M 85 -

+ H2O2 30% 

+ NH4OAc- 3.2 M 

+ HNO3 20%

HF cone. heat- 

+ HC1O4 cone. 

HF cone. heat 

+ HNO, cone. 

 H2O2 30% heat 

-+HC1- 0.05N

 HF cone. heat 

+ HNO3 - cone. 

+ CH1O4

10 Metal Biogeochemistry in Surface-Water Systems
traded with the organics by the hydrogen-peroxide step. 

Finally, the silicates are released by vigorous attack with 

three strong acids (hydrofluoric, perchloric, and nitric).

In addition to these partial extraction schemes for 

fractionation of paniculate material, there has been some success with procedures for separating different forms of dissolved metals. Voltammetric procedures, particularly anodic 

stripping voltammetry (Batley and Florence, 1976; Figura 

and McDuffie, 1979), are commonly used for this type of 

separation. Ion exchange (Figura and McDuffie, 1979) and 

dialysis (Hart and Davies, 1977) are alternative procedures 

that may be used independently or in combination with each 

other to separate labile metal fractions.

Rapin and others (1986) discussed some of the pitfalls in sample treatment that may lead to erroneous results 

in partial extraction procedures to determine metal partitioning. Freezing or short-term wet storage at 1 to 2° C was 

recommended to minimize deterioration of sediments during 

storage. Drying of sediments should be avoided.

Factors That Control Metal Partitioning

Hydrogen-Ion Activity (pH)

The overwhelming importance of hydrogen ion activity is one of the most recurring themes in literature on 

aquatic trace metals. Innumerable studies have demonstrated 

dramatic changes in speciation that may be expected with 

relatively small changes in pH. Furthermore, the pH ranges 

within which such changes occur are usually near neutrality 

and are common to natural waters. Major pathways of trace 

metal partitioning, including complexation, sorption, precipitation, biological uptake, and their respective reverse 

pathways, are all highly pH dependent. The importance of 

accurate pH measurements in any assessment of metal 

behavior in aquatic systems cannot be overemphasized.

The critical role of pH in metal speciation is 

manifested in the frequency with which pH is shown as the 

primary independent variable in speciation diagrams and 

models. One leading example is shown in figure 3 (Stumm 

and Morgan, 1970, p. 270). It illustrates how Cu(II) species 

predominance is controlled by pH. The abundance of a 

species is measured as its activity relative to the activity of 

Cu +2 . At the given total copper concentration, the copper 

species are dominated by two solids: tenorite (CuO) above 

pH 7 and malachite [Cu2 (OH)2 (CO3 )] below pH 7. Among 

the aqueous species, the free metal ion is expected to 

dominate below pH 6, but various carbonate and hydroxide 

complexes become important at higher pH levels.

The diagram in figure 3 is based on chemicalequilibrium calculations and given concentrations of dominant inorganic ligands such as carbonate and sulfate. Other 

similar models, derived from similar types of computations, 

have been published by various authors (for example, Zirino 

and Yamamoto, 1972; Morel and others, 1973; Elder and 

Home, 1978; Vuceta and Morgan, 1978; McKnight and

Morel, 1979; Davis, 1984). The models all show similar 

overall patterns of pH dependence. The pH effect varies as 

ligand concentrations vary, and the precise positions of 

boundaries among different species depend on the particular 

conditions, including the total metal concentration. Nevertheless, the consistently important role of pH in controlling 

speciation is obvious from all models. The sensitivity to small 

pH changes in common natural water pH ranges is usually 

very high.

Effect of pH on Solubility

The hydroxide precipitates of most metals are quite 

insoluble under natural water pH conditions. Because the 

hydroxide ion activity has a direct relation to pH, however, 

the solubility of hydroxide precipitates increases sharply as 

pH decreases. Complexations with other ligands may tend 

to keep the metal in solution, but these complexations are 

also pH dependent. For example, the carboxyl functional 

group ( COO~) is rendered nonfunctional as a metal 

binding site by coordination with a hydrogen-ion (COOH). 

The probability of this occurrence is clearly dependent on 

hydrogen ion activity.

The variation of copper solubility as a function of pH, 

illustrated in figure 3, is one example of the role of pH in 

influencing solubility of metals. Similar sensitivity to pH in 

near-neutral ranges has been determined for cadmium (Morel 

and others, 1973) and iron (Hem, 1985). A detailed discussion of solubility variations of metals as a function of pH 

was provided by Garrels and Christ (1965, chaps. 7 and 11).

Figure 3. Activity ratio diagram showing speciation of Cu(ll) 

as a function of pH (from Stumm and Morgan, 1970, 

p. 270). Total Cu concentration is 0.01 M. Vertical axis is 

logarithmic and is based on ratio of activity of any Cu species 

to activity of the free cupric ion. s, solid; aq, aqueous.

Partitioning of Metals Among Different Phases 11
Effect of pH on Adsorption and Desorption

The effect of pH on adsorption of metals to surfaces 

of iron and aluminum oxides has been studied in considerable 

detail by Leckie and coworkers (Benjamin and Leckie, 1981; 

Davis and Leckie, 1978a, b). Sorption capacity of oxide surfaces generally ranges from near 0 percent to near 100 percent over a pH range of approximately 2 units.

The term "adsorption edge" implies a rapid change 

of adsorption capacity as a function of pH and concentrations of adsorbent and adsorbate. Plots of adsorption edges 

for cadmium, copper, lead, and zinc on oxide substrates 

(Benjamin and Leckie, 1981) illustrate this pH effect. An 

example, shown in figure 4, illustrates the sensitivity of cadmium adsorption on amorphous iron hydroxide. The pH 

range of 5 to 9 brackets neutral conditions and common 

natural water situations. Within this range, there is a span 

of approximately 2 pH units through which small changes 

in pH have a large effect on adsorption of the metal to the 

hydroxide surface. The positions of the curves along the pH 

axis are variable, depending on concentrations of adsorbing 

substance, metal ions, and competing ligands. However, the 

similarity of the shapes of the curves indicates that the intensity of pH effect is nearly independent of adsorbent 

concentration.

Similar effects of pH on adsorption of copper on iron 

oxyhydroxides were reported in earlier papers by Davis and 

Leckie (1978a, b). Nearly identical results were achieved 

by Davies-Colley and others (1984). The strong influence 

of pH can be expected to apply generally to all metal adsorption on solid surfaces, as shown by the investigations

of sorption to various clays and soil minerals by Huang and 

others (1977) and Reid and McDuffie (1981).

Mixing of waters of widely different pH may have 

significant chemical and biological consequences. Examples 

of pH-mixing effects were observed by Theobald and others 

(1963) and McKnight and Feder (1984) at the confluence 

of a high-pH stream with a low-pH stream. The acidic waters 

carry relatively high concentrations of aluminum, iron, zinc, 

and other metals, and the precipitation of these metals as 

hydrous oxides below the confluence limits the abundance 

of periphyton and benthic invertebrates.

Effect of pH on Organometallic Interactions

The affinities of metals for organic molecules are the 

result of linkages between the charged metal ion and 

oppositely charged coordination sites, such as the carboxyl 

and sulfhydryl groups, on the organic molecule. The positively charged hydrogen ion is a competitor with the metals 

for such coordination sites. Thus, hydrogen-ion activity is 

as important in determining the degree of organometallic interaction as are the concentrations of the organic and metallic 

reactants themselves.

Sorption of metals on humic acids is an example of 

an organometallic interaction for which there is some 

documentation of dependence on pH. Kerndorff and 

Schnitzer (1980) showed sharp increases in sorption efficiencies with a pH rise from 2 to 6 (fig. 5). However, the maximum slopes for chromium, lead, copper, and aluminum occur at much lower pH levels than those of cadmium, zinc, 

nickel, cobalt, and manganese. Iron and mercury show very

100

Figure 4. Adsorption of cadmium on amorphous ferric hydroxide, as a function of pH 

and concentration of total iron (from Benjamin and Leckie, 1981). Total cadmium concentration is 5 x 10~ 7 molar (M), temperature is 20° C, and sodium nitrate concentration 

is 0.1 M.

12 Metal Biogeochemistry in Surface-Water Systems
strong humic-acid affinities at any pH greater than 2. Because 

of these differences in responses among different metals, pH 

changes produce changes in the order of sorption. For example, as seen in figure 5, the sorption order at pH 2.4 is 

Hg>Fe>Pb>Cu>Al>Ni>Cr = Zn = Cd = Co = Mn, 

whereas at pH 5.8, the order is Hg = Fe=Pb = Al = Cr = 

Cu>Cd>Zn>Ni>Co>Mn.

This comparison of sorption capacities is similar but 

not equivalent to the frequently cited "Irving-Williams 

series" (Irving and Williams, 1948). Comparing the 

stabilities of metal-ligand complexes of six metals, the IrvingWilliams series reflects an increase in stability as an inverse 

function of ionic radius. According to the series

110

Mn2 + < Fe2 + < Co2 + < Ni2 + < Cu2 + > Zn2+ , copper forms 

the most stable complexes of the six metals. Like the sorption capacity order, this series is also pH dependent.

Effect of pH on Biological Uptake

Given the general concepts that (1) metal speciation 

is highly pH dependent and (2) bioavailability is highly 

dependent on speciation, it follows that hydrogen-ion activity 

is very important in affecting uptake of metals by biota. This 

importance has been documented in various studies. Schecher 

and Driscoll (1985) observed that removal of copper and lead 

by the blue-green alga Nostoc muscorum was highly pH 

dependent (fig. 6). Within acid-pH range, the sorption edges

Figure 5. Adsorption of various metals on humic acids, as a function of pH 

(from Kerndorff and Schnitzer, 1980, p. 1704). Each initial metal 

concentration =0.0005 molar.

Partitioning of Metals Among Different Phases 13
shown were similar to the pattern of adsorption on inorganic 

oxide surfaces. The percentage of metal removed increased 

sharply with increasing pH. Above pH 7, however, the effect was reversed, and removal decreased with increasing 

pH. The departure from the adsorption curve in the alkaline 

range was attributed to release of organic materials capable 

of complexing metals. Precipitation of the metal hydroxides 

was considered in calculating the percent removal by 

biological uptake. Les and Walker (1984) also observed increased binding of copper, zinc, and cadmium by the bluegreen alga Chroococcus parts with pH increases from 4 

to 7. No data were given for the pH effect above pH 7.

Environmental acidification effects on toxicity of 10 

metals to aquatic biota were considered by Campbell and 

Stokes (1985). They pointed out that pH decreases in the 

range of 7 to 4 may affect metal uptake, not only by altering 

metal speciation in solution, but also by altering the 

organisms' sensitivity to the metal at the cell surface. Metals 

may be grouped into two categories according to how their 

toxicity is affected by pH. Type I metals, including cadmium, 

copper, and zinc, produce less biological effect at lower pH. 

Type II metals, of which lead is an example, become more 

bioavailable as pH decreases.

Bioassays of nutrient uptake by the green alga 

Scenedesmus quadricauda in the presence of cadmium and 

copper (Peterson and others, 1984; Peterson and Healey,

96

72

48

24

10'6 /W Pb

10'5 /W Cu

Cell concentration

56.5mg/L(Pb)

104mg/L(Cu)

0 

3.6 6.0

pH

8.4 10.8

Figure 6. Removal of dissolved copper and lead by Nostoc 

muscorum as a function of pH (from Schecher and Driscoll, 

1985, p. 90). M, molar.

14 Metal Biogeochemistry in Surface-Water Systems

1985) demonstrated that toxicity of the metals is highly pH 

dependent. The concentration of cadmium needed to cause 

a 50-percent reduction in phosphorus uptake, for example, 

decreased two hundredfold with a change in pH from 5.5 

to 8.5. This result is consistent with Campbell and Stokes' 

(1985) description of type I metals and appears contradictory 

to the principle that metal complexation increases with pH 

to cause a decrease in bioavailability and toxicity. The explanation suggested by the authors was that the free ionic 

species (Cd +2) is the dominant cadmium species over the 

entire pH range tested. Hence, there was no direct appreciable change in cadmium bioavailability. At the same 

time, the one-thousandfold decrease in hydrogen ion activity resulted in diminished competition from hydrogen ions 

for cellular binding sites, indirectly causing greater 

bioavailability of the cadmium. This explanation was supported by the experiments with copper, which showed similar 

increases in copper toxicity from pH 5 to pH 6.5 but little 

effect above pH 6.5. Unlike cadmium and zinc, copper 

speciation is dominated by inorganic complexation in alkaline 

conditions (fig. 3), and the bioavailability is diminished 

significantly. The drop in competitive complexation by 

hydrogen ions is apparently balanced by the corresponding 

reduction in copper bioavailability.

Implications of Acidic Atmospheric Deposition

Because of the likelihood of rapid desorption of 

metals as pH drops below neutrality, one of the most dramatic 

effects of acid deposition and lake acidification tends to be 

metal mobilization. One notable example is aluminum. 

Although it is the third most common element in the Earth's 

crust (table 1), aluminum was hardly noticed as a potential 

problem in natural water systems until acid deposition began 

to receive national attention. Recent studies have shown that, 

below pH 5, significant amounts of aluminum may be leached 

from soils (Cronan and Schofield, 1979) and solubilized 

through association with organic matter (Campbell and 

others, 1983) or inorganic ligands (Hayden and Rubin, 1974; 

Driscoll and others, 1980). Nordstrom and Ball (1986) 

pointed out the importance of the pH range 4.6 to 4.9 in 

aluminum behavior. Below this range, aluminum behaves 

conservatively, and its activity is controlled by the solubilities 

of sulfate minerals. Above this pH range, aluminum becomes 

nonconservative.

Studies offish responses to acidification (Driscoll and 

others, 1980; Gunn and Keller, 1984; Robinson and Deano, 

1985) have shown that aluminum is more toxic to fish than 

previously believed. Other biotic responses, such as increased 

drift of benthic invertebrates (Hall and others, 1985), have 

been reported. Therefore, increased mobilization of this 

metal may pose a significant environmental threat.

Oxidizing or Reducing Conditions

Aquatic systems are commonly characterized by a 

rather sharp demarcation between oxidizing and reducing en-
vironments. In stratified lakes, this boundary is likely to be 

at the thermocline because of the isolation of the hypolimnetic 

waters from surface oxygenation. In more actively mixed 

waters, the boundary is likely to be in the interstitial water 

of the sediment- water interface. The boundary is characterized by a sharp change in reduction-oxidation potential (redox 

potential or Eh).

The order of metal solubility and the relative 

stabilities of different metal-containing minerals are controlled to a large extent by pH and reduction -oxidation potential (Krumbein and Garrels, 1952). Concentrations of dissolved oxygen and dissolved sulfide, influenced by the redox 

potential but more easily measurable than Eh, are important 

controls on precipitate formation (Berner, 1981). The oxides are one of the predominant precipitates in oxic waters, 

whereas the sulfides are predominant under anoxic conditions. The increasing order of solubility (based on data from 

Ringbom, 1963, p. 341-347) is as follows: 

Oxides: Fe(III) < Hg « Cu < Pb = Zn < Ni < Cd < Mn(II); 

Sulfides : Hg « Cu « Pb = Cd = Ni < Zn < Fe(II)

The position of iron in the sequence changes 

dramatically with the change in redox potential. Under oxidizing conditions, the ferric hydroxide precipitate is highly 

insoluble. Under anoxic conditions, the iron is reduced to 

ferrous iron, and its tendency to precipitate as the sulfide 

is lower than that of many other elements. However, concentrations of ferrous iron may far exceed those of other 

elements because of solubilization from the sediments under 

reducing conditions. The iron may be so abundant that it consumes most of the sulfide, leaving very little to precipitate 

other elements (Morel and others, 1973).

Combined Eh-pH Effect

The sensitivity of iron speciation to both Eh and pH 

is illustrated by the stability-field diagram of Hem (1985, 

fig. 14, p. 80) (fig. 7). Shaded areas indicate conditions in 

which iron exists as a solid; unshaded areas represent 

dissolved species. There are two general regions, representing Eh-pH combinations, where iron solubility is very 

low. One is a condition of strong reduction in the presence 

of sulfur, resulting in precipitation of the sulfide. This condition covers a broad pH range; given sufficient reducing 

potential, it may occur at any natural water pH. The second 

condition is one of moderate oxidation, above pH 4, where 

the ferric hydroxide may precipitate. Outside of these 

regions, especially at low pH, iron is relatively soluble, either 

as ferric complexes when Eh is high or as free ferrous ion 

under more reducing conditions.

The stability field diagram (fig. 7) is somewhat 

variable depending on temperature, pressure, and activities 

of iron and anions. Unless there are drastic changes in conditions, however, the changes are manifiested only as slight 

shifts in the region boundaries. The general stability-field 

pattern remains. The diagram represents inorganic interactions only; under conditions of high organic carbon concentrations, where organic complexation prevails, the field 

compartmentalization would be quite different.

The redox and pH control of iron speciation is 

especially significant because of its indirect effect on speciation of other metals. Formation of hydroxides and oxides 

of iron, as well as manganese, provides ideal surfaces for 

sorption of dissolved metal ions. Conversely, changes in EhpH conditions that redissolve the iron-manganese oxyhydroxides may release various dissolved metals into the water 

column.

Other Effects of Redox Changes

Because of the solubilization of iron and manganese 

in reduced form, a change from oxidizing to reducing environment is generally believed to favor release of metals

Figure 7. Stability field diagram for solid and dissolved forms 

of iron, as a function of Eh and pH at 25° C and 1 atmosphere 

pressure (from Hem, 1985, p. 80). Activity of sulfur 

species = 96 milligrams per liter as sulfate; carbon dioxide 

species = 61 milligrams per liter as bicarbonate; and 

dissolved iron =0.0056 milligrams per liter.

Partitioning of Metals Among Different Phases 15
from the sediments. There is some evidence, however, that 

metal release may actually be diminished under anoxic conditions, because of effects of the anoxia on other constituents 

of the medium. One of the most important is the reduction 

of sulfate to sulfide. Metals that may be released by 

solubilization of iron and manganese oxides may rapidly 

reprecipitate as sulfides. Nondetectable concentrations of cadmium, copper, and nickel in Narragensett Bay pore water 

(Elderfield and others, 1981) were attributed to sulfide 

precipitation.

The reducing environment may cause changes in pH 

that have marked impact on metal speciation. DeLaune and 

Smith (1985) observed that oxygenation of anoxic Mississippi 

River Delta sediments was followed by pH decreases because 

of oxidation of amorphous metal sulfide and pyrite. The pH 

change was greatest in saline environments, but even 

freshwater sediments exhibited a pH drop of approximately 

2 units. The resulting acidity caused solubilization of various 

metals in saline environments. In no case did the authors 

observe decreased solubility in the more oxidizing conditions. 

Apparently, the solubilization due to acidification predominated over desolubilization due to higher oxidation 

potential.

Adsorption to Hydrous Oxides of Iron and Manganese

Among the various aquatic constituents whose 

solubility and speciation are largely controlled by pH and 

redox potential, hydrous oxides of iron and manganese may 

be the most important. These iron-manganese oxide particles 

are very important substrates for sorption of dissolved trace 

metals, and as such, they have major effects on the speciation of trace metals in aquatic systems. The role of iron and 

manganese precipitates in metal speciation merits further 

attention.

Early discussions of the importance of sorption of 

metals on iron and manganese hydrous oxides were focused 

on geochemistry in marine systems (Goldberg, 1954; 

Krauskopf, 1956). The sorption process involving iron and 

manganese oxides in freshwater systems was discussed by 

Jenne (1968). Emphasizing the strong sorptive capacity of 

the hydrous oxide surfaces, Jenne pointed out that the effects of these surfaces on trace metal speciation can be far 

greater than their concentrations would suggest. The reason 

is that amorphous iron and manganese oxides tend to exist 

as coatings (usually on clay particles) rather than as discrete 

particles; thus, they have a high surface to volume ratio.

Since the publication of Jenne's paper, the roles of 

iron and manganese oxides have been tested and discussed 

frequently. One of the important considerations is the effects 

of redox potential and pH on adsorption rates. Reducing conditions and low pH are likely to prevent oxide formation, 

even with much iron and manganese available. Consequently, 

oxide adsorption of metals depends not only on the total concentrations of iron and manganese but also on the conditions 

that allow the oxides to exist in appreciable quantities.

Because of the Eh and pH control, the hydrous oxides 

can be either sources or sinks of trace metals in most systems. 

As conditions change, the direction of metal speciation can 

rapidly change from sorption of dissolved metals on particulate matter to mobilization of particulate metals back into 

the dissolved phase (Jenne, 1976).

The effect of pH on surface charge of inorganic particles (fig. 8) (Stumm and Morgan, 1981, p. 613) shows that 

manganese oxides are almost always negatively charged 

under natural water conditions, whereas amorphous iron oxyhydroxides may be either positively or negatively charged. 

This difference may explain why positively charged metal 

ions such as copper and zinc are commonly sorbed to both 

iron and manganese oxides, but anions such as silica, 

phosphate, and arsenate are sorbed only to the iron oxide 

surfaces (Jones and Bowser, 1978).

Oxidation of iron and manganese to form the hydrous 

oxides is commonly catalyzed by the same elements that are 

subsequently sorbed to them. Copper, for example, has been 

shown to catalyze oxidation of both iron and manganese 

(Jenne, 1968).

Various studies have confirmed active metal sorption 

by iron-manganese hydrous oxides. Rogers and others (1984) 

found that mercury has a strong affinity for both Fe(OH) 3 

and MnO2 , and it tends to desorb very slowly. Mercury affinity for oxides of nontransition elements such as aluminum 

and silicon is weaker. Amirhaeri and others (1984) used a 

chemiluminescence technique to show the rate of Cr(III) adsorption by iron oxide. Approximately 80 percent of 

dissolved Cr(III), at an initial concentration of 1 ^g/L, was 

removed within 3 minutes. Pierce and Moore (1982) 

measured adsorption isotherms and found up to 90 percent 

removal of arsenate on amorphous iron oxides. Arsenate was 

more rapidly adsorbed than arsenite.

A significant fraction of the metals transported by 

rivers is sorbed to iron and manganese hydrous oxide 

coatings. Gibbs (1973) compared the quantities of riverine 

metal transport in two large rivers in very different environments. In the Amazon and Yukon Rivers, he found that 

copper, chromium, cobalt, and nickel were present in the 

form of hydrous iron-manganese oxide coatings in amounts 

up to 50 percent of total metal transport.

In short, there is no apparent disagreement to the 

general statement that iron and manganese hydrous oxides 

play significant roles in trace metal speciation. From a 

physicochemical standpoint, the interaction is a virtual certainty. The hydrous oxide coatings are highly charged 

(Stumm and Morgan, 1981), and they have a surface area 

of some 200 to 300 m2/g (square meters per gram) 

(Horowitz, 1985). Quantification of the actual metal fractions involved is more difficult. The wide discrepancies 

observed in different studies probably reflect real variability attributable to different environmental conditions, 

especially reducing potential and pH. Furthermore, competing ligands may interfere considerably with sorption of

16 Metal Biogeochemistry in Surface-Water Systems
metals to hydrous iron and manganese oxides (Vuceta and 

Morgan, 1978). More study of this type of interaction is 

needed.

Concentration of Organic Matter

Metal affinity for dissolved and particulate organic 

matter is very high relative to other possible binding sites. 

Productivity and decomposition in terrestrial biotic communities ensure a substantial and continuous supply of 

organic matter in most watersheds. Anthropogenic inputs 

have added considerably to the magnitude and complexity 

of organic loads in surface waters. The combination of strong 

metal binding or sorbing capacities and abundance in surface waters renders the organics an extremely important influence on trace metal chemistry.

Effect of Organic Complexation on Solubility of Metals

In the absence of competing dissolved ligands, the 

solubility of metals is controlled by the solubility product, 

which represents the equilibrium reaction:

MAn (s) =A 

where

M = a divalent metal;

A = an anion (usually hydroxide or carbonate in oxic 

waters or sulfide in anoxic waters), which forms 

a precipitate with the metal; 

n = 2 if anionic charge = 1 ; n = 1 if anionic

charge = 2;

K = solubility product; and 

(s) = solid (precipitate).

If a competing ligand is added to this system, it adds 

a branch, or side reaction, as follows:

MAn(s) =M2+ + nA( ~ 2/"> (2)

Ksp =

, = ML(aq)

where

L = a dissolved ligand (usually charged, but the charge

is omitted here for simplicity) and 

Ks = stability constant of the M-L complexation.

As free metal ions are formed by dissolution of the 

solid phase, they are incorporated in the complexation side 

reaction. This removes free metal ions from solution and 

forces still more dissolution. In theory, an equilibrium would 

eventually be reached that conforms to both the solubility 

product, Ksp , and the stability constant, Ks.

Sholkovitz and Copland (1981) cautioned that 

solubilities of metals in natural freshwaters may be very different from those observed in artificial solutions or chemical 

models. The primary reason is interaction with organic 

substances. For example, a number of metals may be held 

in dissolved state by complexation with humic acids, even 

in alkaline conditions (pH as high as 9). At very low pH levels 

(less than 3), humics may be precipitated, taking the metals 

with them. Iron is one of the elements most affected by this 

association with humic acids. More than 70 percent of iron 

is removed by humics at pH 2. This removal rate is very 

different than inorganic solubility considerations would 

predict. On the other hand, manganese and cobalt association with humic acid precipitation is negligible.

Other Factors Affecting Attainment of Equilibrium

One important factor that is ignored in this simplified 

view of metal-ligand-precipitate interaction is the effect of 

kinetics. Reaction rates in natural water systems may vary 

from nanoseconds to decades (Morgan and Stone, 1985), and 

the rates are subject to considerable variation with 

temperature. In the double equilibrium represented by equation 2, one reaction may be several orders of magnitude faster 

than the other. Hence, the theoretical equilibrium may never 

be reached in natural water environments because a ratelimiting step prevents attainment of a steady-state condition

Figure 8. Charge variation on surfaces of some common suspended solids, as a function of pH (adapted 

from Stumm and Morgan, 1981, p. 613). The curves are meant to exemplify trends and are not necessarily 

quantitative.

Partitioning of Metals Among Different Phases 17
(Hem, 1985, p. 25). Types of reactions that are typically 

slow include certain metal-ion oxidations and polymerizations, aging of oxide and hydroxide precipitates, dissociation of metal-organic chelates, and precipitation of metal-ion 

silicates and carbonates (Stumm and Morgan, 1970, 

p. 66).

A second factor that complicates the interaction is that 

each of the reactants may enter into additional side reactions 

(Elder, 1975). The charge on the organic ligand, for example, 

depends on the ligand's degree of association with hydrogen 

ions (protonation), which, in turn, depends on pH. If conditions that control such side reactions change, there is a 

domino effect on the association reactions and an appreciable 

shift in speciation.

Effect of Organic Complexation on Adsorption of Metals

The effect of organometallic complexation on metal 

adsorption is similar to the effect on solubility; it creates a 

multicomponent interaction among the adsorbing surface, the 

organic ligand, the metal, and the other constituents of the 

system. At least three simultaneous types of association compete with each other: sorbing surface-to-metal, sorbing 

surface-to-ligand, and metal-to-ligand. The "winner" of this 

interaction depends on the relative concentrations of the three 

reactants and on the relative association constants of the three 

complexes.

The variability of effects of complexing ligands on 

adsorption was directly addressed by Davis and Leckie 

(1978b). By using model experimental systems to determine 

uptake of Cu(II) and Ag(I) on amorphous iron oxide, the 

authors observed that some ligands enhance uptake, whereas 

others retard it. The mechanism of enhancement, as described 

by the authors, is adsorption of the ligand in a configuration 

that leaves a strongly complexing functional group directed 

outward toward solution. This functional group is then 

available for binding the metal. Examples are thiosulfate, 

glutamic acid, and ethylenediamine. Retardation occurs in 

two ways: (1) when the ligand is not adsorbed but instead 

competes with the adsorbing surface for coodination of the 

metal ion or (2) when the ligand is adsorbed with its functional group involved in the surface bonding, leaving no 

binding sites available for metal ions. The first type of retardation is caused primarily by inorganic ions such as 

chloride or cyanide; the second type may be caused by 

organic ligands such as picolinic acid or salicylic acid.

Later work by Davis (1984) further confirmed the 

importance of adsorbed organic matter in determining trace 

metal partitioning and demonstrated how different metals are 

affected in different ways. An alumina surface covered with 

adsorbed organic matter (extracted from lake sediments) had 

greater adsorptive capacity for Cu(II) than did uncovered 

alumina. The enhancement could again be explained by affinity to exposed functional groups. However, Cd(II) adsorption to alumina was not significantly affected by the organic

coating because cadmium affinity for the organic functional 

groups is weak.

Other investigations of complexation effects on adsorption provide additional documentation of both enhancement and retardation. Elliott and Huang (1979) reported that 

Cu(II) adsorption to alumina was greater in the presence of 

organic chelating agents than in their absence. They also 

pointed out that maximum adsorption should occur at a ligand 

to metal ratio of 1:1. At higher ratios, the excess ligand would 

compete for the surface-binding sites, and, at lower ratios, 

a smaller proportion of the metal would be in the form of 

preferentially sorbed organometallic chelate. Baccini and 

others (1982) obtained nearly opposite results with the same 

reactants Cu(II) and alumina. Natural organic matter from 

lake water was found to retard adsorption because the organic 

matter tended to saturate the binding sites of the mineral surface. The explanation for the different results of these two 

studies may be parallel to the interpretation offered by Davis 

and Leckie (1978b). The organic chelators in the experiments 

of Elliott and Huang (1979) may adsorb with functional 

groups available for binding metal ions, whereas the natural 

organic matter used by Baccini and others (1982) may adsorb without leaving this kind of functional group exposure.

A study of the influence of surfactants on metal-ion 

adsorption by clays (Beveridge and Pickering, 1983) illustrated some of the complexities of the interaction between 

organic binding and surface adsorption of metals. The effect of the surfactants was found to be dependent on the 

charge of the surfactant, as well as on the type of clay. Cationic surfactants reduced metal sorption on montmorillonite 

by competing for surface sites, but the surfactants had little 

or no effect on sorption on two other clays (kaolinite and 

illite). Nonionic surfactant effects were variable with both 

the clay type and the metal involved. Anionic surfactants did 

not interfere directly with adsorption but reduced the concentrations of dissolved metals by precipitating them.

Laxen (1985) demonstrated again the variability of 

complexation effects on adsorption. Adsorption of cadmium, 

nickel, and lead to iron hydrous oxides was enhanced by 

humic acids. The effect on copper adsorption was more 

uncertain. Laxen recognized that humic acids could work 

both ways to facilitate adsorption of the metal or to keep 

the metal in solution by forming stable soluble complexes. 

He suggested that complexation with the sorbed humic acid 

is generally stronger than complexation with the dissolved 

humic fraction.

Fulvic acids typically constitute about 50 percent of 

the total dissolved organic carbon in surface waters 

(McKnight and others, 1983). Complexation by fulvic acids 

may also be important in affecting adsorption (Saar and 

Weber, 1982). Again, adsorption may be either favored or 

retarded, depending on whether the fulvic acids themselves 

tend to be dissolved or adsorbed to sediments. Because fulvic 

acids are more hydrophilic than humic acids, the effect of

18 Metal Biogeochemistry in Surface-Water Systems
fulvic acids on metals may be more likely to favor dissolution and mobility.

Effect of Organics on Other Reactions Incorporating Metals

Just as organic complexation may have either a 

positive or a negative effect on adsorption and solubility, it 

may also either increase or decrease biological uptake of 

metals. Vymazal (1984) observed that increases of humic 

substance concentration from 1.6 to 2.7 mg/L significantly 

decreased uptake of metals by periphytic algae. Baccini and 

Suter (1979) reported similar results. Uptake of copper by 

phytoplankton was positively correlated to the concentration 

of Cu2+ in solution, hence negatively correlated to the proportion of complexed copper. On the other hand, complexing 

agents may increase retention of metals in phytoplankton cells 

by increasing the solubility of the metals and facilitating their 

translocation across the cell wall and plasma membrane 

(Schulz-Baldes and Lewin, 1976). The translocation 

mechanism is important for blue-green algae that produce 

siderophores to improve the organism's capacity to incorporate iron (Murphy and others, 1976; McKnight and Morel, 

1980).

Important oxidation-reduction reactions also are controlled to some extent by organics. The oxidation of ferrous 

iron may be inhibited by organic matter, even in highly oxidizing environments (Theis and Singer, 1973). Two explanations for this interaction were given by the authors: (1) 

iron(II)-organic complexes that are resistant to oxidation may 

be formed and (2) the ferric iron that is produced is promptly 

reduced again by the organics.

Overall Effect on Metal Speciation

In view of the variable effects of organics on metal 

behavior in aquatic systems, it is not surprising that studies 

of metal-organic interactions have produced notably inconsistent results. The presence of organic ligands in natural 

waters causes metal speciation to deviate considerably from 

predictions based solely on equilibrium constants of inorganic 

association and precipitation reactions. Each of the different 

types of association reactions tends to pull the system in a 

different direction. Various physical, chemical, and 

biological controls determine what type of association 

predominates at any given moment. The end result is that 

the presence of organic ligands is likely to have some effect 

on metal speciation, but the magnitude, and even the direction, of that effect is variable.

Biological Activity

Recent studies indicate that aquatic biota are not only 

affected by changes in bioavailability of metals due to partitioning but also play important roles in controlling partitioning mechanisms and pathways. Transformations of metals 

from one species to another are commonly mediated directly

by rnicrobiota. Extracellular sequestering agents may be 

secreted by organisms in response to environmental stimuli 

and may act as important chelators of metal ions. Metals that 

are ingested or otherwise assimilated by organisms may 

undergo species transformations internally before reentering 

the aquatic system.

Methylation and Other Bacterially Mediated Transformations

Bacteria play very active roles in species transformations of metals. One of the most important microbially 

mediated transformations from a human health standpoint 

is methylation (Summers and Silver, 1978; Thayer and 

Brinckman, 1982). Transfer of a methyl group from methyl 

B12 (methylcobalamin) to the metal ion may proceed either 

aerobically or anaerobically and gives the organism a means 

of detoxification in metal-contaminated systems (Wood and 

Wang, 1985). Methyl mercury is an especially hazardous 

product of this process; it is responsible for the catastropic 

cases of "Minamata disease" in Japan. Microbial methylation of mercury, which is most active at the surface of bottom 

sediments, is favored by high temperatures and organic 

enrichment. It is inhibited by oxygenation of sediments 

(Callister and Winfrey, 1986). Other elements that are known 

to undergo biomethylation are chromium, arsenic, selenium, 

and tin (Ridley and others, 1977). Methyl forms of other toxic 

elements, including cadmium and lead, have been observed 

but their formation is less clearly understood.

Numerous other types of microbial transformations 

of metals are important controls on metal cycles and 

biological availability in aquatic systems. Oxidation-reduction 

reactions at the sediment-water interface or the anoxic-oxic 

boundary in stratified lakes are largely mediated by bacteria, 

which gain energy from the conversions (Goldman and 

Home, 1983, p. 105). Sulfate reduction by bacteria (Lovley 

and Klug, 1983) produces the sulfide ion, which binds and 

precipitates many metals. Thus, the process has major 

significance in controlling metal partitioning in anoxic waters.

Algal Activity Extracellular Sequestering Agents

Another type of biological activity that has recently 

been shown to have important effects on speciation and 

bioavailability of metals is extracellular production of sequestering agents capable of significantly altering the metal 

complexing capacity of the medium. Most commonly observed as a function of blue-green algae, this extracellular 

chelator release provides the organism with selective advantages over other species. As nitrogen fixers, some bluegreen algae are able to capitalize on limitation of other species 

in nitrogen-poor waters and undergo dramatic population 

blooms (Home and Goldman, 1972). However, their 

nitrogen-fixing capability is dependent on an adequate supply 

of iron (Murphy and others, 1976; Wurtsbaugh and Home, 

1983). To contend with the likely shortage of available iron 

in surface waters of lakes, the algae are capable of producing

Partitioning of Metals Among Different Phases 19
siderophores (trihydroxamates) that selectively chelate ferric iron (Murphy and others, 1976). The siderophores serve 

as carrier molecules, specifically for blue-green algae, 

transporting iron across the cell membranes. At the same 

time, they render the iron less available to other algal species 

and thus suppress competitive growth.

The work of McKnight and Morel (1980) provided 

evidence that the hydroxamate siderophores released by bluegreen algae are not as selective to iron as previously believed. 

Copper is also complexed by the siderophores. The resulting 

removal of free cupric ions from solution probably diminishes 

toxicity of copper by making it less available. Unlike the ironsiderophore complex, the copper-siderophore counterpart is 

apparently not assimilated by blue-green algal cells. This may 

be attributable to known differences in coordination numbers 

and molecular geometries of the two complexes. The bluegreen algae thus derive secondary benefit from siderophore 

production. Other metals also may be complexed in similar 

fashion.

Figure 9 (from McKnight and Morel, 1980) illustrates 

siderphore complexation of copper. The graph represents 

titration of copper solutions with media from two separate 

cultures of the blue-green alga Anabaena cylindrica. One

I 

b 6

QC 

(SI

o

11

12

6543 

TOTAL COPPER CONCENTRATION, 

AS NEGATIVE LOGARITHM

Figure 9. Free cupric-ion activity as a function of total copper concentrations and concentration of total iron, in 

cultures of Anabaena cylindrica (from McKnight and Morel, 

1980, p. 64). M, molar.

20 Metal Biogeochemistry in Surface-Water Systems

culture was incubated at an iron concentration of 10~ 8 M 

(molar) and the other at an iron concentration of 10~ 5 M. 

Free cupric-ion activity is plotted as a function of total copper; both variables are represented as negative logarithms. 

The solid line represents no complexation; hence, cupricion activity equals total copper concentration.

Extensive complexation of copper occurred in media 

from the iron-limited Anabaena cultures, as shown by the 

curves (fig. 9) which drop well below the no-complexation 

line. Furthermore, more severe iron limitation produced the 

greatest copper complexation and left cupric-ion activity 

several orders of magnitude lower than total copper concentration. The lower iron concentrations stimulated production 

of siderophores, including some that were strong copperbinding agents (stability constant >108).

Earlier studies by McKnight and Morel (1979) indicated that copper complexation by extracellular algal products is common to many algal species. Green algae and other 

eucaryotic species produced copper-complexing agents, 

although of lower complexing strength than those produced 

by blue-green algae. For all species, the benefit was the 

same detoxification of copper. This study was an extension 

of work by Swallow and others (1978), who had found that 

very few algal species produced exudates that could alter 

cupric-ion concentrations. The later study extended the 

analyses to lower exudate concentrations, at which considerably more copper complexation was found to occur. 

However, green algal adaptation to copper (Kuwabara and 

Leland, 1986) could not be explained by media conditioning 

by algal exudates.

Further evidence of detoxification of copper by algal 

extracellular products was provided by Van den Berg and 

others (1979). A diatom, a blue-green alga, and a green alga 

were all shown to produce an exudate that diminished the 

toxicity of copper to photosynthetic activity of Chlorella 

vulgaris.

Enhancement of Iron-Manganese Oxide Formation

Iron and manganese oxidation in natural water 

systems may be catalyzed by microbial activity (Emerson 

and others, 1982; Nordstrom, 1985). The tendency for metals 

to be associated with hydrous oxides or concretions of iron 

and manganese is, therefore, partially dependent on 

biological characteristics of the system. This relation was 

demonstrated experimentally in the field by Kepkay (1985), 

by using azide-poisoned and nonpoisoned dialysis enclosures. 

Formation of manganese concretions, as well as uptake of 

dissolved copper, iron, manganese, and nickel, proceeded 

at a higher rate in nonpoisoned cells. Nordstrom (1985) 

demonstrated similar effects on iron oxidation by Thiobacillus 

ferrooxidans in a stream degraded by acid-mine drainage in 

California.

Other Biological Effects on Partitioning

Assimilation in biological tissue, resulting in bioaccumulation, is an important process affecting speciation
and transport of metals. The biota constitute a major reservoir for both toxic and nontoxic metals. Uptake by biota of 

a metal creates a multitude of transport possibilities, either 

by transfer through various trophic levels (Mathis and 

Kevern, 1975) or by transformation in the organism and 

subsequent release into the abiotic environment (Starkel, 

1985). Further discussion of bioaccumulation and effects on 

biota follows in the Bioavailability and Bioaccumulation section of this paper.

Bioturbation of bottom sediments by benthic 

organisms may cause resuspension of previously deposited 

metals and result in increases in the metal concentrations in 

overlying water (Krantzberg and Stokes, 1985). Some benthic organisms may also alter sediment redox potential, 

which, in turn, affects metal cycling.

Hydraulic Energy Flow and Turbulence

Water movement plays an important role in partitioning of metals in both riverine and lacustrine systems. In 

rivers and streams, the capacity for transporting suspended 

sediments depends on hydraulic flow, turbulence, and morphological characteristics (Guy, 1970). Therefore, the proportion of metals in the suspended phase increases with fluid 

energy. In lakes and reservoirs, the mixing characteristics 

of the system greatly influence the relation between metal 

partitioning and hydrologic factors.

Riverine Hydrologic Effects

In contrast to the correlation between riverine 

discharge and suspended sediment loads, dissolved metal 

concentrations are not necessarily an increasing function of 

discharge (Yeats and Bewers, 1982). If the major inflows 

to a river carry metal concentrations that are relatively constant, and if there are no major seasonal shifts in 

predominance between high-concentration and lowconcentration sources, then increasing discharge is unlikely 

to produce either a dilution effect or a flushing effect. This 

is commonly the case in large rivers (Yeats and Bewers, 

1982; Elder and Mattraw, 1984). Highly concentrated pointsource inputs, which might cause discharge-related variability 

of metal loads in small rivers, are likely to be masked beyond 

detection in the volume of a large river.

Other studies have provided examples of both positive 

and negative effects of discharge on dissolved-metal concentrations in rivers. Positive effects may occur where new 

sources are activated by high discharges, resulting in maximum metal concentrations on the falling limb of the flood 

peak (Bradley, 1984). Negative effects are likely when heavy 

pollution produces high inputs of dissolved, labile forms of 

metals. The dissolved concentrations may decrease during 

high flow because of removal by sedimentation as the ions 

are adsorbed by suspended-sediment particles (Benes and 

others, 1985).

The relation, if any, between metal concentration and 

discharge in riverine systems is a function of several factors, including dilution, desorption of metals from sediments, 

and resuspension of sediments. The relative importance of 

these processes changes as discharge changes, as illustrated 

in figure 10 (from Turk and Troutman, 1981). These curves 

were derived from a study of polychlorinated biphenyls in 

the Hudson River, but the trends also apply to other solutes, 

including some metals, that adsorb to sediments. The authors 

observed that, as river discharge increased from minimum 

levels to about 500 m3 /s (cubic meters per second), the contaminant concentration (dissolved plus paniculate) decreased, 

but, as discharge rose further, contaminant concentration 

(dissolved plus particulate) increased again, reaching maximum levels at maximum discharge.

The explanation of the contaminant-discharge relation illustrated in figure 10 is that different processes 

predominate in different flow conditions. At low flows, 

desorption is the primary mechanism of contaminant movement from sediments to water. As discharge decreases, the 

dilution factor decreases, while desorption remains nearly 

constant, resulting in an inverse discharge-concentration relation. During high-flow periods, resuspension becomes the 

predominant mechanism of contaminant movement from 

sediments to water. As discharge increases, the transport of 

greater sediment loads more than compensates for the dilution effect, and contaminant concentrations increase.

An important result of the overall pattern illustrated 

in figure 10 is that the species composition of the contaminant 

load is highly dependent on discharge. Dissolved forms 

predominate in low-flow conditions, and particulate forms 

predominate in high-flow conditions. In view of the general 

rule that the soluble, labile species are more bioavailable than 

particulate species, the contaminants transported during 

low flows may be of disproportionately high biological 

importance.

The expectation that the dissolved to particulate 

species ratio should be relatively high at low flow is consistent with the observation by Jackson and others (1982) 

that dissolved mercury in the Wabigoon River, Canada, increased in summer, while particulate mercury declined. The 

same pattern was observed by Benes and others (1985) for 

lead in a smelter-polluted river. In the Fraser River system 

in British Columbia, Geesey and others (1984) found that 

the bulk of total copper, nickel, and zinc in the bottom 

sediments consisted of reactive species (extractable with 

dilute hydrochloric acid and peroxide). During periods of 

low flow, there was more rapid deposition of particulate 

metals, causing an increase in the concentrations of reactive 

metals in surface sediments.

Lacustrine Hydrologic Effects

In stratified lakes, the iron cycle sets up in a 

characteristic pattern, illustrated in figure 11 (from Davison 

and others, 1980, fig. 14). Sometimes described as the "iron 

wheel," this cycle exists because of the marked differences 

in solubility and mobility of the two oxidation states. Below

Partitioning of Metals Among Different Phases 21
the oxic-anoxic boundary (near the top of the hypolimnion 

in a stratified lake), ferrous iron predominates. Although the 

ferrous sulfide is insoluble and represents a pathway for 

removal of iron from the hypolimnion, additional iron moves 

in the opposite direction because of dissolution of sedimentary 

oxides and hydroxides. Eddy diffusion in the upper layers 

of the hypolimnion, and turbulence in the epilimnion, mixes 

dissolved iron vertically through the water column. 

Therefore, the entire process is largely driven by wind mixing 

and the hydrodynamics of the system.

Similar effects characterize manganese cycles (Verdouw and Dekkers, 1980; Davison and Woof, 1984). The 

principal difference is that manganese sulfides and oxides 

are somewhat more soluble than their very insoluble iron 

counterparts, and the metal therefore tends to be more 

mobile.

The iron and manganese cycles are highly significant 

in the partitioning and transport of other metals in lakes 

because of the common association between the metals and 

the iron-manganese hydrous oxides. The cycling of iron and 

manganese provides the other metals with a mechanism for 

movement from the sediments upward through the water column to the photic and oxygenated zones of the lake.

As illustrated in the diagram (fig. 11), vertical eddy 

diffusion and turbulent mixing are important processes in the 

iron-manganese cycle. Evidence of vertical eddy diffusion 

in the upper hypolimnion was provided by Hesslein and Quay 

(1973). Turbulent mixing within the epilimnion and diel vertical shifts of the thermocline (Elder and others, 1979) make

further metal transport possible in the water column. 

Biological uptake of the elements in the epilimnion may be 

active and rapid, as this type of cycle is most likely to occur 

during summer months when biological productivity is at 

maximum levels.

In unstratified systems that are wind mixed from top 

to bottom, the hydrologic effect is similar to that in rivers. 

Higher mixing energy produces greater capacity for movement of bottom sediments and all materials associated with 

them, but calm conditions may favor high dissolved to particulate ratios.

Nature of Substrate Material

Metal partitioning depends largely on the types of 

suspended matter and dissolved ligands that are available for 

binding. In systems that contain elevated concentrations of 

organic matter (for example, those that receive large inputs 

from heavily forested watersheds), metals are predominantly 

chelated and adsorbed to organic matter. It is important in 

such systems to examine the ratio of paniculate to dissolved 

organic matter (POM:DOM) and to identify, if possible, the 

major organic species present. Much of the organic speciation will depend on the degradation processes that may break 

down POM before its transport into the aquatic system. It 

will also depend on the timing and severity of hydrologic 

activity that may initiate transport before degradation has progressed appreciably. In watersheds that produce runoff containing predominantly inorganic loads, the clay to silt ratio

1.4

O

<

DC

II o -1

O cc

1-2

1.0

a- 0.8

CO C3

§-0.4

I 

o

Low discharge High discharge

100 200 300 400 500 600 800 1,000 

DISCHARGE, IN CUBIC METERS PER SECOND

1,200 1,400 1,600 1,800

Figure 10. Polychlorinated biphenyl concentration in water (either dissolved or participate), as a function of discharge 

in the Hudson River, N.Y. (from Turk and Troutman, 1981, p. 4).

22 Metal Biogeochemistry in Surface-Water Systems
and the composition of the clay load (fractions of different 

types of clays) will be important in affecting the adsorptiondesorption processes associated with those particles.

Specific surface areas and exchange capacities of 

several sorption active surfaces were tabulated by Forstner 

and Wittman (1981, p. 209). This tabulation revealed 

substantial differences among clays, ferric hydroxide, and 

humic acids. Significant differences in exchange capacities 

existed for different types of clay minerals and for other important sorptive surfaces. Especially noteworthy was the very 

high exchange capacity of humic acids 170 to 590 milliequivalents per 100 grams. This was an order of magnitude 

greater than exchange capacities of most other surfaces listed. 

Similar data for fulvic acids, measured as the conditional formation constants for complexation of copper, were reported 

by McKnight and others (1983).

Sediment Grain Size

General Effects of Grain Size

Because the primary mechanisms of association between metals and sediments are complexation and adsorption processes at the particle surfaces, it follows that the extent of such association will be directly related to the total 

particle surface area.

For a perfectly spherical particle, the surface to 

volume ratio (5: V) may be expressed as a function of radius 

(R) as follows:

Inflow

s

V

(3)

R

Hence, the S: V ratio is inversely proportional to R. The result 

is that, for a constant mass, the surface area increases as particle size decreases. This is illustrated in figure 12, which 

shows the change in surface area as a hypothetical mass of 

uniform density is subdivided into smaller but equally sized 

spheroidal particles. The total surface area approximately 

doubles for every tenfold increase in number of particles. 

If these particles act as sorption surfaces, there should 

theoretically be a corresponding increase in adsorption sites. 

This relation is an oversimplification of natural phenomena, 

because natural particles are not perfect spheres and always 

have a certain degree of porosity, which adds considerably 

to their surface area. Nevertheless, the spherical-particle concept shows that, from a purely physical standpoint, smaller 

particle size correlates with greater surface area.

Grain-Size Effects in Natural Waters

The influence of surface area on metal adsorption was 

demonstrated by Oliver (1973), who determined metal concentrations on various sediment samples from the Ottawa and 

Rideau Rivers in Canada. True surface area of the sediments, 

determined by measuring nitrogen adsorption by a sorptometer, was an important factor in controlling sediment-metal 

concentrations. Concentrations increased with surface area,

Outflow

-- - 

Figure 11. Iron cycle in a lake having an anoxic hypolimnion, showing various transport processes (from Davison 

others, 1980, fig. 14). Relative importance of processes indicated by arrow thickness.

and

Partitioning of Metals Among Different Phases 23
as shown in figure 13. The shape of the curve was found 

to be roughly the same for many other metals, including lead, 

zinc, copper, cobalt, iron, manganese, and chromium. The 

curve is also notably similar to the surface-particle relation 

shown in figure 12.

Other factors, however, usually cause the relation between particle size and metal concentration to deviate appreciably from the simple correspondence to physical changes 

in surface area. Gosz and others (1980) found that the origin 

of the sediments was an important factor. Sediments from 

limestone and sandstone generally showed the characteristic 

inverse relation between metal concentration and grain size, 

but this was not the case for sediments originating in basalt 

or granite. Some of the differences may be attributed to differences in the type of vegetation and organic-matter content. Rickert and others (1977) found that some metals are 

more likely than others to concentrate in the fine fractions. 

The observed concentrations of most metals were much 

higher in the <20-/*m fraction than in the larger fraction, 

but chromium and mercury were notable exceptions. 

Association with relatively coarse organic particles of anthropogenic origin may account for the departure of chromium 

and mercury from the expected concentration-grain size 

relation.

As pointed out by Rickert and others (1977), the 

grain-size effect has major implications for analysis of data 

on total metal concentrations in sediments. Concentration differences among samples may reflect differences in proportion of fine material rather than reflecting variable metal 

loading.

In the Mattole River, Calif., a relatively pristine 

system, Jenne and others (1980) found concentration trends

as a function of grain size for a large number of trace 

elements. Most metals conformed to the normal pattern of 

inverse relation between concentration and grain size. 

However, other elements, such as barium, sodium, and strontium, showed an opposite trend their concentrations were 

higher in larger grain-size fractions. These differences were 

attributed to differences among metals in their tendency to 

sorb to clays, hydrous oxides, and organic matter, all of 

which are concentrated in fine fractions. The discussion included considerable description of methodology for sampling 

and processing both suspended and bed sediment to achieve 

grain-size fraction separation.

Bimodal Distribution

A common deviation from the inverse relation between metal concentration and grain size is bimodal distribution. This phenomenon is characterized by concentration 

peaks in the smallest grain-size fraction and in another fraction of larger particles. Concentrations in intermediate fractions are much lower.

In the Saddle River, N.J. (fig. 14), Wilber and Hunter 

(1979) found that zinc distribution in sediments was a 

decreasing function of grain size at an upstream site but was 

bimodal near the mouth of the river. In both cases, the highest 

concentrations were clearly in the smallest grain-size fraction (< I/mi), but at the river mouth site, another concentration peak occurred in the largest size fraction (1,000-2,000 

/im). The probable reason was that the downstream site was 

affected by municipal effluents. These effluents contained 

relatively high concentrations of coarse organic materials, 

which, because of their high metal sorptive capacity, carried 

disproportionately large amounts of zinc.

3,000

2,500

CO

£ 2,000

LU

| 1,500-

1,000

500

200

10 20 30 40 50 60 

NUMBER OF PARTICLES

70 80 90 100

Figure 12. Total surface area of spherical particles whose total mass equals 1,000 grams and density equals 1 gram per cubic centimeter, as a function of the number of particles.

24 Metal Biogeochemistry in Surface-Water Systems
Bimodal metal distribution on different grain-size 

fractions has been observed in diverse environments, ranging 

from marine sediments (Horowitz, 1985) to small streams 

(Kennedy and others, 1985) and rivers (Rogers, 1983). In 

marine sediments, the secondary concentration peak in the 

larger grain-size fractions can be attributed to coatings and 

atmospherically deposited particles. Also, ferromanganese 

micronodules, rich in associated metals, generally fall in the 

size range of 10 to 20 ^m (Horowitz, 1985). In the stream 

situation, described by Kennedy and others (1985), the 

bimodal distribution was attributable to variability of cation 

exchange capacity (CEC) in going from small to large particle sizes. In the midsize ranges, the grains tended to be composed of a higher proportion of minerals of low CEC, 

whereas both the smaller and the larger grains contained 

higher proportions of highly reactive clay-silt aggregates.

Effects of Grain Size on Metal Transport

Although metal concentrations and grain sizes are 

usually inversely correlated, it does not follow that metal 

transport in rivers is also inversely correlated to grain size. 

Larger grain size generally signifies higher discharge, 

although various other factors, such as particle shape, turbulence, and flow depth, are also important (Guy, 1970). 

Hence, greater total sediment loads and greater total metal 

loads may be expected if grain size increases. However, 

because of the effects of grain size on concentration, this increase in metal transport is not proportional to the increase 

in sediment load.

Although the dependence of metal concentrations on 

sediment grain size is well documented, the assessment of 

metal loading in sediments by examining only the small grainsize fractions may be misleading (Forstner, 1982; Salomons 

and Forstner, 1984). The larger grain sizes are less subject 

to scour; hence, they are less mobile and more likely to provide an accurate long-term record at a given site. The 

midrange size fraction (20-200 ^m) is the most common in 

most bulk sediments. The largest proportion of the total metal 

load may reside in this fraction because of its abundance, 

even if the concentrations are lower than in the smaller fractions. Therefore, particle-size distribution and chemical 

analysis of different size fractions are both necessary to fully 

describe paniculate metal loading under different flow 

regimes.

Temperature, Light, Salinity, and Water Hardness

Numerous other characteristics of the aquatic environment exert various degrees of influence over metal partitioning. Among these are temperature, light, salinity, and 

water hardness.

The importance of temperature derives largely from 

high sensitivity of most chemical reaction rates to temperature 

changes (Morgan and Stone, 1985). An increase of 10° C 

can double the reaction rate (Hem, 1985). Because reaction 

rates determine the tendency of a system to attain equilibrium, 

temperature is a controlling factor on the chemical speciation 

of the system. Various chemical reactions also are sensitive 

to light. Among the important effects of light is photoreduction of iron, releasing soluble Fe(II) from ferric hydroxides (McMahon, 1969; Miles andBrezonik, 1981; Collienne, 

1983). Photo-oxidation of organic compounds may 

simultaneously release metals bound in organic complexes 

(Collienne, 1983). Temperature and light are also important 

controls on biological activity (Goldman and Home, 1983),

10 20 30 40 50 

SURFACE AREA, IN SQUARE METERS PER GRAM

60

Figure 13. Nickel concentrations as a function of surface area in bottom sediments 

of the Ottawa and Rideau Rivers near Ottawa, Canada (from Oliver, 1973, p. 136).

Partitioning of Metals Among Different Phases 25
which may affect speciation of metals associated with 

biological products.

High salinity or water hardness exerts important control on metal speciation. Major ions, particularly magnesium 

and calcium, interfere with complexation and adsorption, 

presumably by competing with trace metals for surfacebinding sites (O'Connor and Kester, 1975; Gupta and Harrison, 1981; Millward and Moore, 1982). In addition, 

alkaline-earth ions, especially calcium, seem to control 

permeability of cell membranes to solutes (Oduleye, 1976). 

Reductions in metal toxicity attributable to increases in water 

hardness have been documented repeatedly (Howarth and 

Sprague, 1978; Chakoumakos and others, 1979; Chapman 

and Dunlop, 1981; Rai and others, 1981; Stendahl and 

Sprague, 1982).

EFFECTS OF METALS ON AQUATIC BIOTA 

Bioavailability and Bioaccumulation

Bioaccumulation of metals in tissues of aquatic biota 

and the physiological and ecological effects of metal uptake 

are important results of metal contamination. Measurement 

of these factors is essential to complete understanding of the 

environmental effects of metals.

There is a tendency to assume that metal uptake rates 

of organisms are directly related to metal concentrations in 

their habitat medium, either sediments or water. In fact, this 

relation has often been documented. Dixit and Witcomb 

(1983) found high levels of copper, lead, and zinc in in5,000

'148 16 32 63 125 250 5001,0002,000 

GRAIN SIZE, UPPER LIMIT, IN MICROMETERS

Figure 14. Concentration of zinc in sediments of the Saddle River, N.J., as a function of grain size (from Wilber 

and Hunter, 1979, p. 798).

vertebrate fauna of the River Irwell, England, and a linear 

relation between concentrations in the animals and those in 

the substrate. In a much less polluted system (Kootenay Lake, 

British Columbia), Evans and Lasenby (1983) found a correlation between concentrations of lead in crustaceans (Mysis 

relictd) and in sediments, even though M. relicta is a benthic detritivore only during daylight hours (it migrates upward to feed on plankton at night).

It is equally common, however, to find cases where 

there is no apparent relation between bioaccumulation and 

concentrations in the habitat medium. Heit and Klusek (1985) 

found no relation between trace element concentrations in 

fish dorsal muscle and those in surficial sediments, despite 

the bottom-feeding habits of the fish species studied. Elder 

and Mattraw (1984) found little evidence of correlation between metal bioaccumulation in Corbicula manilensis and 

concentrations of the same metals in the riverine sediments 

where the animals were found.

In view of the wide range of bioavailability of different forms of metals, it is indeed apparent that substrate 

concentration alone cannot be the sole factor that determines 

uptake by organisms. Marquenie (1985) discussed this 

phenomenon, with description of some case studies. For example, a contaminant contained in a fine sediment might be 

less toxic than the same contaminant at equal concentrations 

in coarse sediment, because the tendency to adsorb to fine 

particles diminishes the bioavailability of the metal.

Multielement Effects

During the past decade, a number of investigations 

have focused on bioaccumulation as influenced by effects of 

multiple elements. In particular, the investigations have 

shown that the concentrations of iron and manganese, in addition to those of the metal in question, influence the ultimate 

effect of the subject metal on the organism. While examining 

the inhibitory effects of cadmium on green algae (Chlorella 

pyrenoidosa), Hart and Scaife (1977) found that cadmium 

accumulation was effectively blocked by moderate concentrations of manganese or iron.

The effect of iron or manganese concentrations on 

metal bioavailability complicates the relation between concentration of a metal and its uptake by biota. Rather than 

a direct relation, bioaccumulation of a metal probably is 

related to an extractable metal to iron ratio or to an extractable 

metal to manganese ratio. More specifically, the ratio of the 

extractable metal concentration in the sediments to the concentration of amorphous oxides of iron and manganese might 

be the important control. As this ratio increases, the number 

of oxide-binding sites that are occupied by the adsorbate 

metal also increases, resulting in a progression toward saturation of the binding sites. Increased occupancy of the binding 

sites diminishes the likelihood that additional metal will be 

sorbed, leaving higher concentrations available for biological 

uptake.

26 Metal Biogeochemistry in Surface-Water Systems
The importance of element ratios in determining 

bioaccumulation was confirmed in a study by Luoma and 

Bryan (1978). They found that the lead to iron ratio in 1 N 

HC1 extract of sediment could explain nearly all variability 

of lead concentrations in the bivalve Scrobicularia plana from 

English estuaries. There was no doubt that this ratio was a 

better factor than total sediment lead concentrations to use 

as a predictor of lead concentrations in 5. plana. The model 

was consistent enough to be a quite reliable predictor of concentrations of metals in bivalves in all estuaries where it was 

tested. Its accuracy was corroborated by Langston (1980) 

in studies of arsenic availability to 5. plana in English 

estuaries.

Similar study of another pelecypod mollusk (Elliptic complanata) by Tessier and others (1984) confirmed the 

findings of the Scrobicularia plana studies. Copper, zinc, 

and lead concentrations in E. complanata were found to be 

much more closely related to sediment extracts (using 

magnesium chloride, acetate, acetic acid, and hydroxylamine 

as extracting agents) than to total sediment metal concentrations. The authors suggested that the protective or competitive 

role of iron was the primary reason for the close relation 

between metal concentration in the organism and the total 

metal extract to iron extract ratio in the sediments.

All of these studies illustrated the importance of the 

association of metals with the amorphous iron oxides, as had 

been proposed by Jenne (1968) and documented by various 

authors since then. Amorphous iron oxide is likely to be a 

relatively small fraction of the total iron in the system, but 

it exerts a disproportionate control over speciation of metals. 

The selective extraction methods used in the studies by 

Luoma and Bryan (1978) and Tessier and others (1984) attack 

the amorphous iron oxide but not the more crystalline iron 

oxide phases. When more vigorous extraction procedures are 

used, the crystalline iron is removed, and the correlation with 

bioaccumulated metal becomes much weaker.

Mechanisms by which iron might exert such control 

over bioaccumulation of other metals were suggested by both 

Luoma and Bryan (1978) and Tessier and others (1984). 

Three scenarios were envisioned, all of which probably occur, depending on the species. The first is that iron and metal 

are ingested together and iron is solubilized by the acidic, 

reducing conditions in the digestive tract of the organism. 

Iron then competes with the other metal for uptake sites in 

the digestive system. The second is that both iron and metal 

are ingested together, but iron remains in particulate form. 

The iron oxide continues to compete with other binding sites 

for metals, just as it does outside the organism. Thus, 

availability of the ingested metal is reduced. The third 

scenario is that uptake of the metals is not through ingestion 

but through the gills and mantle and the association of iron 

oxide and metal in the external medium reduces the availability of the metal.

A study of bioaccumulation of lead in freshwater 

snails (Newman and Mclntosh, 1983b) indicated that the protective role of iron and manganese is not universal. Although 

lead was slowly accumulated by the snails from aufwuchs 

(attached seston and biota) and sediments, bioaccumulation 

was not affected by iron or manganese concentrations in aufwuchs. The authors acknowledged that one reason for this 

may have been insufficient concentration differences to induce different responses. The lead to iron ratio ranged from 

0.007 to 0.016. The other possible explanation offered was 

that the original hypothesis that iron or manganese oxides 

may interfere with lead accumulation may have been incorrect. The need for more research on this topic was 

emphasized.

Multielement effects involving other metals were 

described by Popham and D'Auria (1982) and by Calabrese 

and others (1984). Uptake of copper by the mussel Mytilus 

edulis was enhanced by high concentrations of lead, zinc, 

or silver. Copper was apparently accumulated preferentially 

to the other elements. The presence of the other metals may, 

in some way, facilitate transport of copper across the cell 

membrane, but specific explanations for this were not given.

Kuwabara (1985) found that interaction between zinc 

and phosphorus can increase the growth-limiting effect of 

phosphorus for algae (Selanastrum capricornutum). Elevated 

zinc concentrations may interfere with phosphorus 

metabolism. Other metals that are involved in nutrient 

assimilation and metabolism include manganese, iron, 

copper, molybdenum, and cobalt (Goldman and Home, 

1983).

Nonuniform Bioaccumulation Among and Within Individuals

Bioaccumulation and toxic effects of metals correlate 

in some cases with size and age of the organisms (Bryan and 

Hummerstone, 1978; Bryan and Uysal, 1978; Tsai and 

Chang, 1984). However, in other cases, there is no apparent 

relation between bioaccumulation and animal size (Harvey 

and Knight, 1978). Of the two gastropod species studied by 

Newman and Mclntosh (1983a), one showed no relation between lead concentration and size, whereas the other showed 

a curvilinear relation. An explanation for the inconsistent importance of size in controlling bioaccumulation was given 

by Smock (1983), with reference to concentrations in aquatic 

insects. Two general patterns of relation to size were proposed, as illustrated in figure 15. The mode of uptake is 

shown to be a factor that affects bioaccumulation and the 

effects on the organism. Where adsorption on the exterior 

surfaces of the organism is the primary mechanism of uptake, the effect of size is precisely the same as it is in 

sediments larger size means smaller surface-to-volume ratio 

and decreased uptake. This relation is illustrated by curve 

A in figure 15. If absorption is the primary mechanism of 

uptake that is, if the metal is ingested with food and absorbed through the gut, size of the organism has little effect. If anything, there is a slight increase of metal concentration with increasing size, as illustrated in curve B in 

figure 15.

Effects of Metals on Aquatic Biota 27
The results of Zaroogian and Johnson (1984) show 

that uptake of nickel by the bivalves Crassostrea virginica 

and Mytilus edulis is indeed dependent on size and follows 

the general pattern of curve A (fig. 15). These animals are 

filter feeders and might be expected to assimilate the metal 

internally, as had been demonstrated for another bivalve 

mollusk by Bryan and Uysal (1978). According to the model 

of Smock (1983), curve B should apply, and there should 

be no effect of size on metal concentration. The apparent 

contradiction indicates that the processes involved are not 

fully understood and that the effect of size may be more complicated than is indicated by Smock's model.

Bioaccumulation rates may vary considerably among 

different organs of the same individual. The freshwater 

mussel Anodonta cygnea was exposed to 10 ^ig/L mercury 

and 16 ^ig/L cadmium, and separate analysis of bioaccumulation in different organs were performed by Balogh and 

Salanki (1984). Both metals in nearly all organs accumulated 

linearly during the 840-hour experiment, but at different 

rates. A ten- to twenty fold difference was observed between 

kidney and adductor muscle accumulation.

Effects of Speciation on Bioaccumulation

Different metal species in sediments and water have 

different degrees of bioavailability. Aquatic biota respond 

quickly to changes in concentrations of free metal ions, but 

the availability of metals bound to organic and inorganic

CO

QC 

O

< w

QC 2

o -

O

_J

<

LLJ

INCREASING * 

ORGANISM DRY WEIGHT

Figure 15. Metal bioaccumulation by an organism, as a 

function of weight of the organism (from Smock, 1983, 

p. 319). Curve A, bioaccumulation primarily by adsorption; 

curve B, bioaccumulation primarily by absorption.

ligands, or adsorbed to suspended matter, is generally less 

than that of the free ions (Marquenie, 1985). Although studies 

of both speciation and biological effects of metals are common, there are relatively few studies that link measured concentrations of specific metal forms with bioaccumulation or 

biological responses.

Studies such as those by Diks and Alien (1983) and 

Kuwabara and others (1986) do provide some insight into 

the nature of speciation effects on bioaccumulation. Diks and 

Alien (1983) analyzed concentrations of copper in various 

sediment and water fractions from Dlinois and Indiana rivers. 

These data were compared to metal concentrations in tissues 

of tubified worms that were exposed for up to 12 days to 

the same sediments and water. A high correlation (ft > 0.97) 

was found between the copper extracted from the amorphous 

manganese oxide fraction and the concentration of copper 

in the worms. Other copper fractions were less correlated 

to bioaccumulation. Kuwabara and others (1986) tested 

growth response of the green alga Selenastrum capricornutum 

to dissolved and particle-bound orthophosphate and zinc. The 

particle-bound fractions of both elements (adsorped to particles of titanium oxide) were less available to the algae than 

zinc and phosphorus in solution.

Toxicity

Use of Biota as Pollution Indicators

Increasing concern over hazardous waste contamination of natural water systems in recent years has led to increasing development of methodology for assessing contaminant effects. One area of development has been in the 

use of aquatic biota as indicators of contamination. This use 

generally entails one of two approaches: (1) examination of 

occurrence and distribution of natural populations in the affected habitat or (2) examination of response of nonnative 

organisms when placed in the affected habitat or a simulation of the habitat. The principal advantage of using bioindicators is that they integrate effects of whatever 

contaminants may be in the system. Conversely, one 

disadvantage is that observed effects are difficult to attribute 

to a particular contaminant in any system having multiple 

contaminants and sources.

Use of bioindicators to assess trace metal contamination has had some success in recent years. Examples are 

studies by Sheppard and Funk, 1975; Winner and others, 

1980; Whitton and others, 1981; Bryan, 1983; Leland and 

Carter, 1985; and Swift, 1985. The results of all of these 

studies indicated that the existence of water-quality problems 

was signaled by changes in the biotic community either in 

the community structure or community metabolic processes, 

or both.

In rivers, the distribution of different types of 

organisms can be a useful bioindicator. The characteristic 

continuum is that shredders and collectors dominate in headwaters, grazers and collectors in midwaters, and gatherers 

in large rivers (Vannote and others, 1980). Responses of ben28 Metal Biogeochemistry in Surface-Water Systems
thic invertebrates to pollution abatement in the Penobscot 

River, Maine, were interpreted by Rabeni and others (1985) 

as illustrating a shifting of the riverine community continuum 

because of changes in contaminant levels. Pollution tended 

to shift this continuum in a downstream direction.

Whitton and others (1981) listed several advantages 

to using bioindicator species, particularly plants, to monitor 

heavy metals in rivers. The advantages are significant but 

should be interpreted with caution. For example, the metal 

composition of a particular plant species is likely to be less 

heterogeneous than that of the sediments. The restriction to 

"a particular species" is critical here, because differences 

among species in their resistance and adaptation to contaminants yield a great deal of cross-species heterogeneity 

(Luoma, 1977; Norris and Lake, 1984). Comparisons oif 

bioaccumulation in different species should not be attempted 

as a means of comparing pollution effects in different 

systems. Even within one species, individual variations and 

factors that may not be taken into account, such as differences 

in size, age, or reproductive state, may produce 

heterogeneous results (Luoma, 1977; Ney and Van Hassel, 

1983). If these variables are adequately taken into account, 

bioindicator organisms can be valuable tools for assessing 

problems of contamination, especially when used in combination with chemical analyses.

Norris and Lake (1984) noted that the reliability of 

using bioindicators depends on the metal being tested. Copper and zinc bioaccumulation by fish correlated quite well 

with water concentrations of the same elements, but the same 

degree of correlation could not be shown for cadmium. This 

difference probably is caused by distinct speciation patterns 

of the different elements and distinct pathways for their incorporation into biological tissues. Compartmentalized bioaccumulation among different tissues of rooted plants was 

demonstrated by Buhl and McConville (1984).

Among the advantages of using plants as bioindicators, Whitton and others (1981) included the integrated 

record that is produced. Knowing what facet of the system 

is being integrated, however, is also important. Some species 

of aquatic macrophytes extract metals primarily from the 

sediments, while others extract from the water (Heisey and 

Damman, 1982). Even an individual species may extract different metals from different phases. Campbell and others 

(1985) observed that most of the copper accumulated by the 

water lily Nuphar variegatum was derived from the 

sediments, whereas most of the zinc in the plant came from 

the water.

Effects of Metals on Phytoplankton

A considerable amount of work has been done over 

the past several decades that demonstrates impacts of trace 

metals on phytoplankton activity in aquatic systems. 

Phytoplankton are quite sensitive to most metals, and most 

studies that examine algal response to changes in metal concentrations, whether in the laboratory or in the field, show 

positive results.

The effects of metals are not limited to toxicity; 

metals may also stimulate activity in some situations. The 

most common metal to show stimulatory effects is iron, but 

similar effects have also been observed for other metals, including manganese, cobalt, vanadium, and molybdenum 

(Wiessner, 1962; Goldman, 1972). Copper and zinc, which 

are frequently toxic to algae and other biota at relatively high 

concentrations, are essential micronutrients. Deficiency 

symptoms may develop in algae when copper or zinc concentrations fall below 1Q- 7 or 10~ 6 M (Wiessner, 1962).

Copper and zinc are good examples of elements that 

have ambivalent biological impacts, especially with respect 

to phytoplankton. They may either be nutrients or toxicants 

at natural water total concentrations. This is explained by 

limited bioavailability of bound and precipitated species. If 

natural water ligands are present in concentrations that exceed that of the metal, complexation may leave a bioavailable 

metal fraction that is several orders of magnitude less than 

the total concentration and less than the micronutrient requirements for the metal (Manahan and Smith, 1973). 

Relatively small changes in the total ligand to metal ratio, 

or changes in ambient conditions, particularly pH, may shift 

the equilibrium such that the concentrations of the 

bioavailable metal become high enough to be toxic.

The stimulatory capacity of copper has some implications for possible "backfiring" when using this element as 

an algicide (Telitchenko and others, 1972). The amount of 

the copper addition is critical. Although too much may be 

costly and dangerous to other organisms, too little may serve 

only to raise the Cu2 + concentration slightly toward an optimal level and actually encourage algal growth (Elder and 

Home, 1978).

Toxicity as a Function of Bioaccumulation

Results of studies show that the question of whether 

or not toxicity is related to the concentrations of metal accumulated by the organism does not have a simple answer. 

Cadmium and copper toxicity to daphnids was affected by 

humic-acid concentrations, but bioaccumulation was not affected (Winner, 1984). Chlorella pyrenoidosa growth was 

sensitive to cadmium in experiments performed by Hart and 

Scaife (1977), but a great deal of cadmium could be accumulated in the organism without shutting off its productivity, possibly because of production of protective cadmium sequestering agents. The relation between toxicity and 

bioaccumulation is clearly variable among different species 

and is probably dependent on the mechanism of uptake and 

metabolism (Leland and Kuwabara, 1985).

Baccini (1985) noted that, when a phytoplankton 

community is exposed to a trace metal stress, the community structure tends to shift toward those species that accumulate the lowest concentrations of the metal in their

Effects of Metals on Aquatic Biota 29
tissues. Species that accumulate high concentrations suffer 

the greatest effects on metabolic rates and may eventually 

disappear from the community.

Effects of Metal Interactions

The biological effect of one metal may be influenced 

by the presence of others. In some cases, this effect may be 

to reduce toxicity (antagonism), and in other cases, the effect may be to increase it (synergism). Tolerance of white 

suckers to cadmium increased when they had been previously 

exposed to cadmium or mercury (Duncan and Klaverkamp, 

1983). Several mechanisms for this enhanced tolerance were 

suggested, including decreased uptake, increased excretion, 

redistribution of metals to less sensitive biochemical storage 

sites, and induced synthesis of metallothionein (a protein that 

facilitates metal storage and detoxification). A similar conditioning effect of prior metal contamination was observed 

by Titus and Pfister (1984). Bacteria in sediments of the Ottawa River showed higher tolerance to cadmium when concentrations of various other metals were high.

Growth of copepods that were exposed to various pair 

combinations of arsenic, cadmium, copper, lead, mercury, 

and zinc was inhibited in all cases more than when exposed 

to single metals (Borgmann, 1980). However, very little 

synergism and no antagonism were observed; the toxicities 

of the metal pairs could be accounted for by summation of 

the individual toxicities. Hart and Scaife (1977) could not 

find any evidence of effects of metal interactions with most 

combinations of metals in their experiments with Chlorella 

pyrenoidosa. Cadmium accumulation, although unaffected 

by extracellular presence of cobalt, copper, molybdenum, 

or zinc, was almost totally blocked by manganese.

The capacity of an organism to develop tolerance to 

metals also was demonstrated by Seim and others (1984). 

Steelhead trout that were continuously exposed to copper 

were better able to grow and survive than if the exposure 

was intermittent (4.5 hours per day). Continuous exposure 

may allow the organism's defense mechanisms to develop 

more effectively than would be allowed by temporal variation in exposure levels. Unfortunately, the intermittent exposure situation is more common in natural waters.

TRANSPORT OF TRACE METALS 

IN RIVERS AND LAKES 

Forms of Transport in Rivers

Riverine transport of trace metals is generally 

characterized by high proportions of particulate species, 

especially during high flows (Jackson and others, 1982; Yeats 

and Bewers, 1982; Benes and others, 1985). Even when soluble species dominate the composition of inflowing water, 

they have strong affinity to three types of suspended 

materials organic detritus, clay mineral particles, and ironmanganese oxide coatings all of which are abundant in most 

river waters. Many studies of metals in rivers indicate the

relative amounts of metals in dissolved and particulate phases, 

but few go beyond that. However, some information is 

available to describe partitioning of metals transported in 

large rivers.

Some of the most detailed work on forms of metal 

transport in rivers has been done by Gibbs (1972, 1973, 

1977). Five phase categories were recognized: (1) dissolved, 

(2) adsorbed on solids, (3) associated with particulate organic 

material, (4) as metallic coatings (adsorbed to hydrous oxides 

of iron and manganese coated on suspended particles), and 

(5) in crystalline material structures where the metals are 

almost totally unavailable for further cycling. Most work was 

done on two very large rivers the Amazon and the Yukon. 

Striking similarities were found between the two rivers with 

respect to the relative proportions of different metals 

transported in the respective phases. The proportions varied 

among different metals, however. Manganese speciation was 

dominated by coatings, as might be expected, but relatively 

high proportions were also found in crystalline and dissolved 

forms. Chromium was distributed primarily between 

dissolved and crystalline species, with very little of the element associated with coatings. More than three-quarters of 

the copper was in crystalline form, with the remainder almost 

evenly distributed among the more available forms.

Rivers that carry heavy loads of suspended sediments 

are naturally those that exhibit the highest fractions of particulate metal transport. This may explain the similarity between the Amazon and the Yukon Rivers, both of which are 

characterized by heavy sediment loads (Gibbs, 1977). In contrast, the St. Lawrence River, which carries a relatively 

lighter suspended sediment load, displays lower particulate 

to dissolved metal ratios (Yeats and Bewers, 1982). Even 

then, the ratios are still much greater than unity.

A study of 20 of the world's major rivers by Martin 

and Meybeck (1979) also demonstrated how different metals 

vary considerably in form of transport. This variation was 

attributed primarily to differences in weathering rates and 

coordination chemistry. Dissolved fractions constitute a 

substantial fraction of the total transport of arsenic, copper, 

molybdenum, and zinc, whereas 90 percent of the elements 

chromium, cobalt, lead, manganese, and nickel are carried 

in particulate form. Some elements, such as zinc, may occur in higher concentrations in river water than can be accounted for by composition of watershed substrate materials. 

These high concentrations indicate remobilization of such 

elements from the bottom sediments, which provide an 

autochthonous source that adds to constituents in inflow from 

watershed drainage.

Differences in speciation behavior of cadmium, copper, lead, iron, and zinc in the Yarra River, Australia, were 

observed by Hart and Davies (1981). The analytical approach 

permitted separation of particulate, filterable, ionexchangeable, dialysable, and bound-metal fractions. The 

dominant forms of the metals in the river were, respectively:

30 Metal Biogeochemistry in Surface-Water Systems
Cadmium: free Cd2+

Copper: ion exchangeable

Iron: bound (nonion exchangeable) 

and particulate

Lead: PbCO3°

Zinc: free Zn2+ 

Of the ion-exchangeable fraction, organic complexes 

were important for copper but not so important for lead and 

zinc. Increases in concentrations of iron, lead, or zinc were 

manifested primarily as increases in bound and particulate, 

not ion-exchangeable, species. This trend presumably reflects 

equilibrium shifts toward adsorption when metals are introduced as the free ion or other labile forms.

Particulate transport is favored in relatively unpolluted systems, where metal inputs are principally from 

erosion of watershed substrates. In polluted systems, where 

sources are likely to deliver metals in a soluble phase or 

loosely associated with organic material, the proportion of 

dissolved metals tends to be higher (Salomons and Forstner, 

1984, p. 162).

Studies in a variety of localities have demonstrated 

that the species of metals transported in rivers are transformed when the metals enter coastal estuaries (Kharkar and 

others, 1968; Hart and Davies, 1981; Ellaway and others, 

1982; Danielsson and others, 1983). In most cases, these 

transformations were observed to proceed toward greater particulate fractions, as flocculation or coagulation was triggered 

by increased dissolved solids in the estuarine water. 

However, Kharkar and others (1968) reported that significant amounts of some metals are released from particles into 

dissolved forms when reaching the oceans. This release was 

calculated to contribute most of the dissolved cobalt supplied 

to the oceans from rivers. In their study of the Amazon 

plume, Boyle and others (1982) found that copper and nickel 

were relatively unreactive upon discharge to the ocean, but 

there was some indication of desorption of cadmium.

Transport in Lakes

Most investigations of metal transport and cycles in 

lakes and estuaries have focused on iron and manganese 

(Evans and others, 1977; Jones and Bowser, 1978; Davison, 

1982; Morris and others, 1982; Davison and Seed, 1983; 

Sundby and Silverberg, 1985). The cycles of these two 

elements have significant influence over the cycles of other 

elements because of sorption and desorption from iron and 

manganese hydrous oxides.

Particle flux, or movement of suspended particles 

through the system and into the bottom sediments, is another 

important control on metal cycles in lakes (Baccini, 1985; 

Joshi, 1985; Sigg, 1985). Data from various lakes of different morphological features were used by Santschi (1984) 

to compare mean residence time of zinc with particle flux 

to the bottom, in grams per centimeter per day. A close inverse correlation was found. Elements that have the greatest

affinity for particles, such as iron and cobalt, are most likely 

to conform to this type of correlation with particle flux 

(Hesslein and others, 1980). Characteristics of the particles, 

especially ionic radius, electronegativity, mineralogy, and 

porosity, are also important in determining their capacity for 

transporting dissolved metals (Jackson and others, 1980).

OVERVIEW

The metals that have been discussed in this review 

are extremely important constituents of natural water 

systems. They are ubiquitous, and they are readily associated 

with other constituents both abiotic and biotic. Characterized by a small ionic radius and high surface charge, their 

biogeochemical interactions are diverse and complex. Metals 

may accumulate in particulate matter and biota. In water, 

they commonly occur at concentrations of less than 1 /*g/L, 

but even at such concentrations, they are likely to have important biological effects. Their effects on sensitive populations can substantially affect the structure and function of 

the aquatic community. Effects on human populations also 

are important. Transferred through food chains, or 

transported to potable water sources, metals can create special 

problems for human health and aquatic-resource 

management.

Research on occurrence, transport, biogeochemical 

cycles, and biological effects of metals in aquatic systems 

has been active, especially in the past two decades. In view 

of the complexity of metal behavior and extremely low levels 

of effective concentrations, researchers are faced with some 

formidable problems. With the aid of improved partial extraction techniques and analytical methodology, many studies 

have provided detailed insight into specific situations having 

well-defined conditions. When biota are involved, the success of the investigation usually depends on limitation of the 

study to one species or to a few very similar species. 

Generalizations of results from one situation to another are 

extremely difficult and usually unsuccessful. Models are 

useful to facilitate understanding of fundamental chemistry 

of the metals, but applied to natural systems, they are prone 

to inaccuracy because of the complexity and natural variability of environmental factors.

In spite of such difficulties, considerable progress has 

been made in understanding trace metal biogeochemistry. 

Although research results commonly conflict as a result of 

natural variability, a number of common themes emerge from 

the vast body of literature on the subject. Specific studies 

explore various facets of these basic themes and serve to provide additional documentation and detail to them. The 

following itemized statements synthesize important themes 

and represent the current state of our understanding of metals 

in surface waters, derived from results of research to date. 

 Anthropogenic and natural processes are important as 

sources of trace constituents. Some metals, such as 

aluminum and iron, are very common in the Earth's

Overview 31
crust and are present in natural waters at substantial concentrations. These concentrations vary geographically 

(being controlled by the geologic characteristics of the 

watershed) but are likely to be quite stable over time. 

Other elements (for example, cadmium, mercury, 

selenium, and silver) are rare in the Earth's crust. 

Background concentrations of such elements are low but 

usually detectable by state-of-the-art analytical tools. The 

presence of any of the metals in natural waters at concentrations higher than background is usually attributable 

to one of three common anthropogenic sources: (a) 

deposition of atmospheric pollutants, (b) point-source 

discharges of municipal or industrial waste products, and 

(c) nonpoint-source runoff from agricultural or 

silvicultural areas or urban impervious surfaces. 

In areas of mining activity, mine drainage can 

significantly increase concentrations of various metals, 

such that the contaminated waters may be uninhabitable 

by most aquatic biota and the water may be unusable 

for agricultural or domestic purposes. The metal contaminants are by no means limited to those that are 

mined. Contaminants are released in the processing of 

mineral ores, and their transport is facilitated by the high 

acidity usually characteristic of mine drainage. 

Metals accumulate in sediments and biota at much higher 

concentrations than those that remain in the water. 

However, the dissolved fractions can be very important 

biologically because they represent a principal source 

of bioavailable metals.

Anthropogenic input of metals is likely to increase the 

proportion of dissolved metals in the system, at least near 

the input source, because metal pollution is usually accompanied by other types of pollution, such as acidification or elevated concentrations of dissolved organic 

matter, which favor solubilization of the metals. 

Because of their charge and small ionic radius, freemetal ions in solution have a very strong tendency to 

become associated with various other constituents of the 

aquatic medium. Some of these associations cause 

precipitation, some lead to incorporation in colloidal or 

suspended particles, and some leave the metal in adissolved state but bound with dissolved ligands as complexes or chelates (complexes with multiple binding 

sites). Nearly all of the association reactions that result 

in the formation of different forms, or species, of metal 

are reversible. Slight changes in environmental factors, 

such as temperature or pH, can produce appreciable 

shifts in species composition. The result is that, at any 

point in time, the metal exists as numerous species and 

the proportions of these species are highly subject to 

change. Activities of the dissolved free-ion species are 

generally several orders of magnitude lower than those 

of the total metal concentration. 

Two important partitioning processes precipitation and 

adsorption cause the predominance of paniculate metal

fractions over dissolved fractions. Metals are only 

slightly soluble in water (solubility products of hydroxides, carbonates, and sulfides are low), assuring low concentrations of the dissolved forms under common natural 

water conditions. If charged surfaces are present, which 

is virtually always the case in natural waters, adsorption is a predominant process, sometimes accounting for 

nearly 100 percent of the metal partitioning. Both 

precipitation and adsorption are reversible a fact that 

is critical in trace metal cycling. 

Three types of paniculate substances, abundant in natural 

waters, provide surfaces for metal adsorption. All have 

very strong affinities for metal ions. They are (a) claymineral particles, (b) paniculate organic matter, and (c) 

hydrous oxides of iron and manganese. 

The most important single variable that controls metal 

partitioning is pH. Because the equivalence points of 

many metal-ligand associations fall within the pH range 

of 5 to 9, major shifts in proportions of different metal 

fractions can occur with relatively mild changes in pH, 

as are common in the natural environment. Studies of 

aquatic metal biogeochemistry must include reliable pH 

measurements.

Complexation by dissolved organic ligands is almost certain to affect adsorption and bioavailability of the metal, 

but the direction of the effect may be either positive or 

negative, depending on the conditions and the metals and 

ligands involved. This effect can be predicted if association constants are known for the metal-to-ligand, metalto-surface, and ligand-to-surface combinations. 

Amorphous hydrous oxides of iron and manganese are 

among the most important types of adsorption surfaces 

for trace metals. Because of their surface charge, these 

oxides have a strong affinity for other metals, and they 

are usually abundant in natural waters. Because they 

commonly occur as coatings on other particles, they tend 

to have high surface exposure relative to their mass, and 

their effect on metal adsorption may be disproportionately high.

Changes in the redox potential between oxidizing and 

reducing conditions profoundly affect metal partitioning. 

Because of the solubilization of iron and manganese oxides, reducing conditions are generally expected to favor 

release of metals from sediments to water. However, 

this is not always the case, because other factors that 

accompany a redox drop may counteract the release from 

iron-manganese binding. Most important among such 

accompanying factors are increases in pH and reduction of sulfate to sulfide, both of which favor reprecipitation of released metals.

Trace metal concentration in sediments generally increases with decrease in grain size, primarily because 

of the higher surface to volume ratio of finer particles. 

But this general rule may be altered by the nature of the 

particles. The existence of metal-rich nodules raises con32 Metal Biogeochemistry in Surface-Water Systems
centrations in large-size particles. This commonly results 

in a bimodal distribution, whereby elevated concentrations are observed in both very fine and coarse fractions, 

with smaller concentrations in intermediate-size fractions. The same type of distribution may occur in 

systems subject to large inputs of coarse organic particles, such as from a municipal-waste discharge because 

of strong affinity between metals and organic surfaces. 

In rivers and streams, hydrologic changes are important to both transport and speciation of metals. In 

medium- to high-discharge conditions, an increase in 

flow promotes higher metal concentrations because of 

the greater sediment load. The effect of this change is 

an even greater relative increase in metal transport (concentration x flow). In low-discharge conditions, concentrations tend to decrease as flow increases, because 

of dilution of metals desorbed from bottom sediments. 

During high flows, paniculate to soluble fraction ratios 

are likely to be much higher than during low flows. 

In lakes and reservoirs, the hydrologic action is also 

critical to metal transport. In stratified systems, the 

highest concentrations of dissolved metals are in the 

hypolimnion, whereas the primary producing populations are in the epilimnion. The thermocline barrier to 

upward transport of nutrients and metals from hypolimnion to epilimnion may be partially overcome by current and eddy effects caused by wind action. Wind action 

can provide mobility of the elements from their sources 

to where their biological effects may be most evident. 

Metals tend to accumulate in bottom materials and biota. 

As a result, metal "sinks" exist in the bottom materials 

and biota. However, such accumulation is not necessarily equivalent to permanent removal from circulation 

through deposition in an immobile reservoir or to complete export from the system. Instead, the immobilization is often only a temporary stop in a continuous cycle 

through the system. The accumulated metals may be 

returned to general circulation upon suitable changes in 

environmental conditions.

Bioaccumulation of a metal in native organisms may or 

may not correlate to concentrations of the same metal 

in the habitat medium either water or sediments. If 

bioaccumulation does not correlate, the concentrations 

in the medium probably do not correspond to 

bioavailability. The latter situation is very common 

because of the abundance of iron and manganese oxides, 

particulate organic matter, and fine clay-mineral 

sediments in water systems. All of these substances 

strongly adsorb metals, diminishing their bioavailability, 

and thus "protect" the organisms from the 

contaminants.

Toxicity of a metal to aquatic populations is manifested 

as impairment of metabolic functions or limitation of 

distribution or abundance of the populations. For any 

given population, toxicity may or may not be correlated

to bioaccumulation of the metal by organisms in the 

population. Bioaccumulation occurs nonuniformly 

among different tissues and organs, and the distribution 

factors vary greatly among different species. Thus, some 

species may accumulate considerable toxic metal loads 

without suffering significant impairment, whereas others 

are much more sensitive. The duration and continuity 

of exposure also influence an organism's capacity to 

resist toxic effects.

Bioaccumulation and toxicity can be affected by various 

characteristics of the organism, such as size, age, 

reproductive state, and feeding habits. However, there 

is no general pattern of such relation, and in many cases, 

the variables may be independent of each other. The 

magnitudes and directions of the relation, if any, depend 

to a large extent on the mechanism of uptake of metals 

by the organism.

Biological effects of metals are not limited to toxicity. 

At concentrations higher than background, but within 

common natural water ranges, many metals may be 

nutritive, toxic, or variable between the two, depending 

on the biota present and the environmental conditions. 

The biological effect of one metal may be modified by 

the presence of others. If various metals are present at 

contaminant levels and a new contaminant is introduced, 

its effect may be greater or less than it would be in an 

uncontaminated situation. Its effect would depend on 

which of the following interactions predominates: (1) 

improved effectiveness of the organism's toxicity 

defense mechanisms, such as synthesis of metallothioneins, due to its prior exposure to toxic elements, 

(2) weakening of the organism's defense mechanisms 

caused by the stresses imposed by the existing contamination, and (3) cation exchange with the introduced 

metal that releases previously sorbed metals into a 

bioavailable pool. The amelioration of metal 

bioavailability by adsorption to hydrous oxides of iron 

and manganese is another example of effects of element 

interactions.

Various forms of aquatic biota, particularly blue-green 

algae, can partially control bioavailability by production 

of extracellular sequestering agents. These substances 

are advantageous to the organism in at least three ways: 

(1) they facilitate uptake of limited quantities of nutritive 

elements such as iron; (2) they complex toxic elements, 

thereby diminishing their bioavailability; and (3) they 

reduce the availability of nutritive elements to competing 

species.

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