EPA/600/3-86/023
March 1986
THE EFFECTS OF VARIABLE HARDNESS, pH, ALKALINITY,
SUSPENDED CLAY, AND HUMICS ON THE CHEMICAL SPECIATION
AND AQUATIC TOXICITY OF COPPER
Henry Nelson
. 2
Duane Benoic
2
Russ Erickson
2
Vince Mattson
Jim Lindbergh
^Science Applications International Corporation, McLean, Virginia.
2
U.S. Environmental Protection Agency, Environmental Research Laboratory,
Duluth, Minnesota.
3 . . •
University of Wisconsin at Superior, Center for Lake Superior Environmental
Studies, Superior, Wisconsin.
ENVIRONMENTAL RESEARCH LABORATORY
OFFICE OF RESEARCH AND DEVELOPMENT
U.S. ENVIRONMENTAL PROTECTION AGENCY
nnii iTu mm rron/i
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NOTICE
This document has been reviewed in accordance with
U.S. Environmental Protection Agency policy and
approved for publication. Mention of trade names
or commercial products does not constitute endorse-
ment or recommendation for use.
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ABSTRACT
The effects of variable hardness, pH, alkalinity, hutnics, and suspended
clay on the chemical speciation of copper and its toxicity to fathead minnow
larvae in Lake Superior water were investigated. Two proposed methods
(toxicity factors and chemical speciation) for predicting LC50 values in
specific natural waters from laboratory toxicity data and the average site-
specific values of general water quality parameters were evaluated. The
~ 2
accuracy of the cupric ion selective electrode in determining Cu activities
in ambient and chemically altered Lake Superior water was also determined.
Increases in calcium and magnesium hardness at constant ambient Lake
-3
Superior alkalinity (approximately 1 x 10 eg/L) increased LC50 values in
+2
terms of total, dissolved, and free copper (Cu activity) as did increases in
sodium. Increases in pH from 6.6 to 8.7 at ambient Lake Superior alkalinity
increased the total and dissolved copper LC50s. However, the free copper LC50
increased from pH 6.6 to 7.3, remained relatively constant from pH 7.3 to 8.0,
and then decreased from pH 8.0 to 8.7. At approximately three times the
-3
ambient Lake Superior alkalinity (3 x 10 eg/L), the total and dissolved
copper LC50s increased monotonically, and.the free copper LC50 decreased raono-
tonically with increasing pH from 7.1 to 8.5. Increases in alkalinity from
approximately one-third to six times the ambient alkalinity of Lake Superior
did not significantly affect the total and dissolved copper LC50s but
decreased the free copper LC50s.
The differences between LC50 values for some waters with higher than
ambient Lake Superior alkalinity and/or pH appear to be due primarily to
changes in the proportions of inorganic copper species with different toxici-
ties/unit concentration. However, the differences between LC50 values for
waters with lower than ambient alkalinity and/or pH and for waters at ambient
alkalinity and pH but different hardnesses cannot be explained by changes in
the proportions of inorganic species. In such cases, it appears likely that
changes in general water quality parameters such as pH and hardness, change
the toxicity/unit concentrations of one or more toxic copper species.
i i i
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Additions of humics arid/or suspended clay Co Lake Superior water
increased total and dissolved copper LC50s, but generally decreased free
copper LC50s even though they did not have any appreciable affect on pH or
alkalinity. The decrease in the free copper LC50 with additions of humics
and/or clay may have at least been partially due to increases in the toxici-
ties/unit concentrations of inorganic copper species due to a possible
increase in the stress on the organisms. However, it is also possible that
some copper humic and/or copper clay complexes are directly toxic to fathead
minnow larvae.
The apparent changes in the toxic ities/unit concentration of toxic copper
species with changes in the values of general water quality parameters such as
pH or alkalinity indicates that the feasibility of developing a chemical spe-
ciation method for deriving site-specific criteria is low. The development of
a toxicity factors method, which involves empirically deriving from laboratory
data multi-variable equations relating LC50 values to general water quality
parameters such as pH and hardness, appears to be more feasible but is still
being evaluated.
+ 2
The ratio of Cu activities determined by the ion selective electrode to
+ 2
Cu activities predicted from inputing dissolved copper into the REDEQL
chemical equilibrium computer program (assuming no organic or clay complexa-
tion) varied from 0.85 to 1.15 for 12 of the 18 test waters to which no humics
or clay were added. The four test waters for which the ratios were less than
0.85 had ambient Lake Superior or lower pH and/or alkalinity. Most of the
test waters for which the ratios were close to one had greater than ambient
Lake Superior pH and/or alkalinity. These observations along with the
observed dependence of the slope of the cupric ion electrode response on
various parameters indicated that a substantial proportion of copper may be
bound by organics in Lake Superior water at ambient or lower pH and/or
alkalinity even though the TOC of Lake Superior water averages only 1 ppra.
The reason may be due to the relatively high stability constants for the
formation of many copper-organic complexes, and to a reduction in the
- -2 +2
competition between OH , CO^ , and organic ligands for Cu at ambient or
lower pH and/or alkalinity.
i V
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ACKNOWLEDGEMENTS
We would like to thank the following individuals for their assistance:
1. Charles Stephan, who was responsible for the design of the flow-
through electrode chamber, and with whom we held many helpful discus-
sions on the use, calibration, and behavior of the cupric ion
electrode.
2. Ron Carlson, who conducted the bioassay work for the part of the
Connecticut field study concerning electrode determinations of LC50s,
and with whom we hald helpful discussions on the interpretation of
the field study data. Mr. Carlson is currently writing a much
larger, more detailed report on the Connecticut field study.
3. Robert Andrews, with whom we held several helpful dicsussions on the
use, calibration, and behavior of the cupric ion electrode.
5. Ed Leonard and Don Ruppe, who performed some of the atomic absorption
analyses of total and dissolved metals.
This research was partially funded under EPA contract No. 68-01-6388,
which Science Applications International Corporation (formerly JRB Associates)
had with the Office of Water, Criteria and Standards Division, U.S. Environ-
mental Protection Agency, Washington, D.C. This report constitutes the final
product of work assignment 16 under this contract.
V
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TABLE OF CONTENTS
Page
CHAPTER 1: INTRODUCTION 1- 1
1.1 The Chemical Speciation of Copper in Freshwater Systems...1- 2
1.2 The Effects of General Water Quality Parameters
and Chemical Speciation on Copper Toxicity to Fish 1- 5
1.3 Toxicity Factors Method 1- 8
1.4 Chemical Speciation Method 1- 9
1.5 Useful Equations for the Interpretation
of Toxicity Test Data 1-13
CHAPTER 2: METHODS AND MATERIALS 2- 1
2.1 General Apparatus and Materials for Cu
Activity and pH Determinations 2- 1
2.2 Cupric Ion and pH Electrode Calibrations 2- 1
2.3 Cu+^ Activity and pH Measurements
in Flow-Through Systems 2- 6
2.4 Atomic Absorption Spectroscopy 2- 7
2.5 Flow-Through Exposure Systems 2- 8
2.6 Flow-Through Toxicity Tests 2- 9
2.7 Cd+^ Electrode Calibration and Cd+^ Activity
Determinations in Flow-Through Exposure Systems 2-10
2.8 Connecticut River Field study on the Naugatuck River 2-13
2.9 Steady State/Equi 1 ibrium Comparisons 2-16
2.10 Copper Titrations of Lake Superior
Water and Reconstituted Water ' 2-17
2.11 Chemical Speciation Calculations 2-18
vi
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TABLE OF CONTENTS (continued)
Page
CHAPTER 3: RESULTS AND DISCUSSIONS 3- 1
3.1 Set 1 — Separate Effect of MgCl^, CaCl^i
and NaCl Additions on Copper Toxicity
in Lake Superior Water 3-14
3.2 Set 2 — The Effect of pH on Copper Toxicity
in Lake Superior Water at Ambient Alkalinity 3-20
3.3 Set 3 — The Effect of pH on Copper Toxicity in
Lake Superior Water with the Alkalinity Increased
to Approximately Three Times Ambient Alkalinity..... 3-26
3.4 Set 4 — The Separate and Joint Effects
of Humics and Suspended Clay on Copper Toxicity 3-33
3.5 Set 5 — The Effects of Variable Humic Concentrations
on Copper Toxicity in Lake Superior Water 3-37
3.6 Set 6 — The Effect of Variable Alkalinity
on Copper Toxicity in Lake Superior Water 3-41
3.7 Field Study on the Naugatuck River, Connecticut 3-44
3.8 Possible Variations in Larvae Sensitivity
to Copper Toxicity Over Time 3-48
3.9 Differences Between Steady State
and Equilibrium Concentrations 3-51
3.10 Theoretical Copper-Organic Complexation
Based on the Slopes of the Electrode Response 3-53
3.11 Multiple Linear Regressions 3-66
CHAPTER 4: CONCLUSIONS AND RECOMMENDATIONS 4- 1
4.1 The Apparent Dependence of Toxicities/
Unit Concentration on the Values of
General Water Quality Parameters 4- 1
4.2 Potentially Toxic Copper-Organic Complexes 4- 3
4.3 Conclusions with Respect to the Feasibility
of Developing a Chemical Speciation Method
for Predicting Copper Toxicity in Site Waters 4- 5
4.4 Conclusions with Respect to the Feasibility
of Developing a Toxicity Factors Method for
Predicting Copper Toxicity in Site Waters 4- 7
4.5 Recommendations
vii
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LIST OF FIGURES
Page
2-1 A Typical Calibration Curve for the
Cupric (Cu ) Ion Electrode 2- 5
+ 2
2-2 The Calibration Curve for the Cadmium (Cd ) Ion Electrode.... 2-12
2-3 Sampling Sites Along the Naugatuck River ....2-15
3-1 The Separate Effects of > Mg*^ and Na+ on the
Negative Logarithms of Cu and Dissolved Copper LC50s 3-16
3-2 Th^Effect of pH on the Negative Logarithms of
Cu and Dissolved Copper LC50s in Lake Superior
Water at Ambient Alkalinity 3—? 1
3-3A Th|2EffecC pH on the Negative Logarithms of
Cu and Dissolved Copper LC50s in Lake Superior
Water at Three Times Ambient Alkalinity 3-28
3-3B A Comparison of the Effect of pH on the Negative Logarithms of
Cu and Dissolved Copper LC50s in Lake Superior Water
at Ambient and Three Times Ambient Alkalinity 3-29
3-4 The Separate and Joint Effects of Humics and
Suspended Clay on the Negative Logarithm of Cu
and Dissolved Copper LC50s 3-34
3-5A The Effect of Variable Humic ^ncentrations on
the Negative Logarithms of Cu and Dissolved
Copper LC50s in Lake Superior Water ..3-38
3-5B The Effect of Variable Humic Concentrat ions on
the Negative Logarithms of Cd and Dissolved
Cadmium LC50s in Lake Superior Water 3-40
3-6 The Effect of Vaj^able Alkalinity on the Negative
Logarithms of Cu and Dissolved Copper LC50s in
Lake Superior Water at Ambient pH 3-43
+2
3-7 The Variation in the Negative Logarithms of Cu
and Dissolved Copper LC50s in Ambient Lake Superior
Water Between Different Sets Over Time 3-50
3-8 The Effects of UV Irradiation and/or HPLC Clean Up
on Copper Titrations of Lake Superior Water and
Burdick and Jackson Water 3-60
3-9 The Effects of Ca ^ Zn Na+^, pH and Alkalinity
on Copper Titrations of Lake Superior Water 3-63
vi i i
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LIST OF TABLES
3-1 General Water Quality Characteristics and 96-Hr LC50
Va^es in Terms of T^al Copper, Dissolved Copper,
Cu Activity and Cu Concentration for the Test
Waters of Each Set of Copper Toxicity Experiments 3- 3
3-2 The Average pH, 96-Hour LC50 in Terms of Diss^ved
Copper (Moles/L), 96-Hour LC50 in Terms of Cu
Concentration (Moles/L), and Corresponding Calculated
Concentrations (Moles/L) of CuOH+, CutOH)^0, CuCO^",
Cu(CO ) , and Cu-Org for the Test Waters of Each
Set or Copper Toxicity Experiments 3- 6
3-3 The Average pH, 96-Hr LC50 in Terras of Dissolved Copper
(Moles/L), Percentages of Dissolved Copger Made Up By
Cu , CuOH+, CuCOH^", CuCO^0, CutCO^)^ > Respectively,
the Sura of the Inorganic Percentages and the Theoretical
Copper-Organic Percentages 3-10
3-4 Static 96-Hour Acute Value9 for One-Day-Old Fathead
Minnow Larvae in Water Taken from Various Sites
Along the Naugatuck River in Connecticut 3-46
+ 2
3-5 Comparison of Equilibrium Cu Activities
to Steady State Cu Activities 3-52
3-6 Comparison of Computed Free Copper Percentage of
Dissolved Copper at the LC50 Point Based on Electrode
Determinations to RSDEQL Computed Free Copper Percentages
Based on Inorganic Speciation Alone for the Six Sets
of Copper Toxicity Tests..... 3-56
3-7 Comparison of Computed Free Copper Percentages of
Total Copper Based on Electrode Determinations to
REDEQL Computed Free Copper Percentages Based on
Assumes Variable Glycine Concentrations 3-65
ix
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1. INTRODUCTION
The aquatic toxicities of metals have been shown to be at least partially
dependent upon the values of general water quality parameters such as tempera-
ture, dissolved oxygen, hardness, pH, alkalinity, total organic carbon (TOC),
and suspended solids (1-20). Since the average values of such parameters can
vary substantially over the range of natural waters normally encountered (21),
there is interest in formulating protocols by which site specific criteria for
metals can be derived (22). EPA is currently evaluating the feasibility of
developing protocols for formulating site specific criteria based on non-site
laboratory toxicity data bases and on average site specific values of general
water quality parameters. Such protocols are potentially more cost effective
than those based on performing toxicity tests at each site.
The development of protocols for site specific criteria depends upon the
determination of which water quality parameters significantly affect metal
toxicity. It also depends upon the development of one or more methods by
which the toxicity of metals in a site water can be estimated (e.g., the LC50
values for an appropriate sensitive species) from non-site laboratory toxicity
data bases and average site specific values of those general water parameters
that significantly affect toxicity.
For the past year, scientists at the EPA lab in Duluth have been conduct-
ing toxicity tests to determine the effects of five general water quality
parameters (hardness, pH, alkalinity, humics, and suspended solids) on the
toxicity of copper to fathead minnow (Pimephales promelas) larvae. The tests
were designed to determine which of those general water quality parameters
exert a significant effect on copper toxicity. In addition, the tests are
being used to evaluate the feasibility of predicting copper toxicity for fat-
head minnow larvae in site waters from the results of laboratory toxicity
tests and the average site specific values of the general water quality
parameters found to significantly effect toxicity.
Currently, scientists at the lab are evaluating two proposed methods for
estimating copper LC50 values in site water:
1-1
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a) Toxicity Factors Method - involves empirically deriving one or more
equations relating the LC50 or a LC50 related transformation to one
or more general water quality parameters (or transformations) such as
pH, and hardness. This method is described in Section 1.3.
b) Chemical Speciation Method - involves determining the relative con-
centration and toxicity/unit concentration of the various copper
species present. This method is described in Section 1.4.
Brief summaries of the literature on the chemical speciation of copper in
freshwater systems and the effects of chemical speciation on copper toxicity
are presented in Sections 1.1 and 1.2, respectively.
1.1 THE CHEMICAL SPECIATION OF COPPER IN FRESHWATER SYSTEMS
The chemical speciation of copper in actual and model freshwater systems
has been predicted by numerous authors (6-13,23-36) based on the use of chemi-
cal equilibrium computer programs such as REDEQL. In addition, several groups
have also based their calculations of the concentrations of various copper
+ 2
species on the actual experimental determinations of Cu activities with a
cupric ion selective electode (6,8,32-36). We have also based our calcula-
tions of the concentrations*of copper species in toxicity test waters on
+ 2
cupric electrode determinations of Cu activities (Section 1.4). However,
for comparative purposes, calculations of copper species concentrations were
also performed by inputing dissolved copper determinations into the REDEQL
chemical equilibrium computer program.
The relative proportions of the free Cu+^ and various copper hydroxy and
copper carbonate complex concentrations depend upon the pH and alkalinity of
the system and the magnitudes of the stability constants for the formation of
the complexes. Although the pH' and alkalinity of freshwater systems vary
substantially (21), most authors have predicted that the chemical species
+ 2 + -2
Cu , CuOH , Cu(0H)2°, CuCO^", and CutCO^)^ together represent greater than
98% of the dissolved inorganic copper in the typical freshwater systems they
have considered (6-13, 23-26, 32-35). Those conclusions concur with our
chemical speciation calculations based on both cupric electrode determinations
of Cu+^ activities and on REDEQL calculations for the toxicity test waters
described later in this report. Although we considered such chemical species
1-2
-------
as Cu2(OH)2+^, CuHC03+, CuSO^ and various copper chloride, anionic copper
hydroxy, copper ammonia and copper phosphate species, che calculated contri-
butions of all such species to the total inorganic copper in the toxicity test
waters were negligible (individually <1%, together <2%).
The total dissolved copper in many freshwater systems may also consist of
substantial amounts of copper-organic complexes including complexes with amino
acids, carboxylic acids, hunic acids, and various effective copper chelators
(24, 25, 27, 31, 33, 35-54). Although chelators such a NTA and EDTA do not
appear to occur naturally (37), they can be introduced to freshwater systems
by industrial and municipal sewage discharges (37, 38, 55). Furthermore,
there is evidence that aquatic organisms such as algae and some types of
invertebrates and fish secrete substantial quantities of chelating agents in
response to copper stress (37, 56) .
The stability constants for the formation of many copper-organic com-
plexes and the extent of copper binding by many types of humics (37) appear to
be of sufficient magnitude to possibly result in substantial copper-organic
complexation, even in waters with relatively low organic content (24, 35, 39).
However, there is still substantial debate over the concentrations and types
of organics required to complex substantial amounts of copper (37). The TOC
of Lake Superior water (which served as the base and diluent water for all of
our in lab toxicity tests) is on the order of only 1 ppm (57). however, the
magnitude of the greater than Nernstian slopes of the cupric ion electrode
response along with theoretical REDEQL calculations using glycine as a model
of the average organic nitrogen present in Lake Superior water indicate that
substantial copper-organic complexation may occur in Lake Superior water and
Lake Superior water with lowered pH and/or alkalinity, as will be further
discussed in Section 3.10.
The total copper in some freshwater systems substantially exceeds the
dissolved copper due to copper precipitation and adsorption to or binding by
suspended solids and colloids (33). We have attempted to determine which
solids control the copper solubility in our toxicity test waters by consider-
ing the pH and alkalinity of the test waters, and the solubility products as
1-3
-------
listed in Martel and Smith 1976 (58) for the following solids: CuO, . (teno-
I s)
rite), Cu„(0H)oC0., . (malachite), and Cu,(0H)„(CO.)., Our calculations
I I .Hs; J L J ivs;
indicated that CuO^^ controlled the solubility of copper for all of our
toxicity test waters and that the copper solubility decreased with increasing
pH over the pH range of our toxicity tests. Our calculations also indicate
that the highest concentration of total copper used in each type of toxicity
test water was below the calculated total solubility of copper which was
defined as the estimated sum of the Cu CuOH , Cu(OH)„°, CuCO ° and
-2
Cu(CO^^ concentrations at the saturation point.
Differences between total and dissolved copper in freshwater systems may
also be due to adsorption onto various types of suspended solids and colloids
including hydrous iron and manganese oxides (59), clays (60, 61), aluninum and
silica oxides (24, 62) and hunic material (35-54). The adsorption of copper
onto suspended clays and oxides generally increases with increasing pH
(60-62). Increases in the organic content of sediments and suspended solids
also appear to increase copper adsorption (63). Although increases in the
overall organic content of the water are reported to enhance the adsorption of
copper onto suspended clays and oxides by some groups (60-62), increases in
the dissolved organic content of water is reported to greatly reduce such
adsorption (35, 64). The apparent discrepancy has been attributed to differ-
ences between the effects of suspended or colloidal organic matter and dis-
solved organics (65). Lake Superior water has a very low suspended solid
content so that adsorption of copper onto suspended material was relatively
low in all of our toxicity test waters which did not have humic or clay
additions. That was demonstrated by the small differences (eg., <15%) between
dissolved and total copper measurements. The addition of Lake Superior shore
clay and of Aldrich soil derived hunic acid did lead to substantial adsorption
of the copper onto suspended solids and colloids. The increase in the dif-
ference between total and dissolved copper with hunic additions indicated that
some of the humic material remained in suspension either as particulate matter
or as colloidal material.
1 -4
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1.2 THE EFFECTS OF GENERAL WATER QUALITY PARAMETERS AND CHEMICAL SPECIATION
ON COPPER TOXICITY TO FISH
Our data (Chapter 3) along with most of the data in the literature indi-
cate that increases in hardness, alkalinity, pH, humic content, TOC in
general, and suspended solids generally decrease copper toxicity to fish
(increase LC50 values and chronic level values in terms of total and dissolved
copper) (1-20). The observed changes in total and dissolved copper toxicities
with changes in the values of general water quality parameters can probably be
primarily attributed to:
a) Changes in the relative concentrations of different copper species
which may exert different toxicities/unit concentration on the test
o rg an i sm s
and/or
b) Changes in the toxicity/unit concentration exerted by toxic copper
species on the test organisms which could be reflective of physiolog-
ical changes within the organisms and/or changes in copper transport
rates into and out of the organisms.
Although not completely conclusive, our toxicity data tends to support the
postulate that hardness affects total and dissolved copper toxicity primarily
through mechanism b, that alkalinity effects are primarily through mechanism
a, and that pH, hunic and suspended clay effects are through both mechanism a
and b.
A number of groups have previously attempted to determine the relative
contributions of different copper species to copper toxicity to fish by com-
paring the results of copper toxicity tests in terms of total or dissolved
copper with the corresponding calculated concentrations of the various pos-
sible copper species (6-16). The results of such analyses have not been
completely conclusive and are not always in agreement. Difficulties with
these types of analyses in the past have included a statistically small number
of data points compared to the number of chemical species considered and
substantial correlations (e.g. lack of independence) between the concentra-
tions of the different chemical species as discussed by Magnuson et al., 1979
(12). At least part, but generally not all of the correlation between the
1-5
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concentrations of some of the different chemical species can be eliminated by
varying the values of general water quality parameters such as hardness,
alkalinity and pH independently of one another as we have done for the sets of
experiments described in this report. However, in addition to affecting the
relative concentrations of different chemical species, changes in the values
of general water quality parameters such as pH may also change the toxicity/
unit concentration exerted by one or more of the toxic copper species on the
organism (15). Our data tends to support that postulate as will be discussed
in Chapter 3 and 4. Therefore, it may not be possible to quantitatively
determine the relative toxic ities/unit concentration of different chemical
species whose relative concentrations depend upon the values of general water
quality parameters that also affect the toxic ities/unit concentration.
Despite difficulties with quantitative analyses, the groups who have
tried to determine the relative toxic contributions of different copper
species have, at least qualitatively, attempted to identify those copper
species which do, and do not, contribute significantly to copper toxicity
+ 2
(6-16). Either computer estimated or electrode determined Cu activities
have been correlated with copper toxicity to fish by all of the groups (6-16).
In addition, most have suggested that CuOH+ probably also contributes signi-
ficantly to toxicity (7, 9-14). There is no general agreement on the contri-
bution of Cu(OH) ° to toxicity perhaps at least partially due to the greater
4
than 10 range in the reported magnitude of the stability constant for its
formation (35,66). The species CuC0^° and Cu(C0^)^° have generally been
described as contributing very little, if any, to toxicity based on a lack of
correlation between the concentrations of those species and copper toxicity
(7-16). However, one group has suggested that CuCO^0 can contribute signifi-
cantly to copper toxicity when it constitutes a high proportion of the copper
present (6). Copper bicarbonate CuHCO. and the anionic copper hydroxy
-2
species such as Cu(OH)^ and Cu(OH)^ are not believed to contribute signifi-
cantly to copper toxicity based on either a lack of correlation between
predicted concentration and copper toxicity, or to the low level of the pre-
dicted concentrations (6-16). Although several groups have suggested that
Cu2(OH)2+^ may be toxic (9, 10, 11), the predicted concentration of that
species in most freshwater systems appears to be too low to contribute signi-
ficantly to toxicity (11).
1-6
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Huraic acid (17, 18, 20), NTA (6), EDTA (16), glycine (16, 20), citrate
(16), and municipal sewage effluent (67) have all been reported to reduce the
toxicity of copper to fish in terms of total and/or dissolved copper. How-
ever, it is still possible, as pointed out by Borgmann 1983 (15), that copper-
organic complexes do exert some toxicity to fish, although lower than that
+ 2 ...
exerted by Cu and perhaps other toxic inorganic copper species. Although
+2
constant LC50 values in terms of free copper (Cu ) in the presence and
absence of such organics would indicate that the copper-organic complexes were
+ 2
non-toxic, a reduction in the Cu LC50g would indicate that the copper-
organic complexes might be contributing to toxicity (15).
+ 2
There is some evidence which indicates that Cu LC50 remain relatively
s 3
constant for algae and bacteria in the presence and absence of various kinds
of organics (15) although there is little, if any, comparative data of that
nature available on fish. However, a recent study on the effects of various
amino acids on copper toxicity to the aquatic invertebrate Daphnia indicated
that the addition of any of the amino acids tested substantially reduced the
Cu ^ LC50 (68). Although that indicates that the copper-amino acid complexes
may contribute to toxicity, the increase in total copper LC50fi with the amino
acid additions indicates that the toxicity/unit concentration exerted by the
copper-amino acid complexes is lower than the concentration weighted average
toxicity/unit concentration exerted by the inorganic copper species. Quali-
tatively similar results were obtained with guppies, but the apparent
toxicity/unit concentration of the copper-amino acid complex was far less than
for the Daphnia (68). Our toxicity data indicate that humic additions
+ 2
increase total and dissolved copper LC50s, but decrease Cu LC50 . There-
s
fore, the copper-humic complexes may exert some toxicity, but with a lower
toxicity/unit concentration than the concentration weighted average toxicity/
unit concentration of the inorganic copper species.
The toxicity of copper precipitates and copper adsorbed onto suspended
solids has not been well characterized, but the small amount of data available
indicates that those forms of copper are probably relatively non-toxic com-
pared to comparable quantities of some dissolved forms of copper (9, 16, 69).
However, our data on the effects of suspended clay on copper toxicity indicate
1-7
-------
+ 2
that Cu LC50g are reduced by suspended clay additions which indicates that
copper-clay complexes might exert some toxicity to fathead minnows.
1.3 TOXICITY FACTORS METHOD
The toxicity factors method involves estimating LC50 values in site
waters by first empirically developing an equation relating the LC50 value in
terms of an appropriate measurable form of copper to general water quality
parameter variables through raultivariable regression on laboratory toxicity
data. Average site specific values of the general water quality parameters
are then substituted into the equation. An example of such an equation can be
found in Howarth and Spraque 1979 (10).
The selection of an appropriate measurable form of copper should be based
upon the ease at which the measurement can be made in natural waters and upon
how closely the measured form approximates the bioavailable fraction of cop-
per. The measurable forms of copper whose LC50 values are dependent upon the
least number of general water quality parameters will be the ones which most
closely approximate the bioavailable fraction of copper. Measurable forms of
copper which have or will be considered in the future include total copper,
0.45 um filtered (dissolved) copper, copper fractions with molecular weights
< 1,000 as determined by dialyses or ultrafiltration and uncomplexed "free"
— *2 . . ...
copper as determined by Cu activity measurements with an ion selective
electrode.
If an equation relating LC50 values in terms of a chosen measurable form
of copper to general water quality parameters is developed from laboratory
toxicity data, the accuracy of the equation for estimating LC50 values in site
waters can be tested by comparing equation estimated LC50 values to LC50
values determined from toxicity tests conducted at several different sites
that are representative of the wide range of natural waters normally encount-
ered.
1-8
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1.4 CHEMICAL SPECIATION METHOD
The chemical speciation method of estimating LC50 values in site waters
involves determining the relative contributions of different soluble species
of copper to the overall toxicity of the copper in the site waters. Assuming
additive toxicity, the contribution of a particular soluble species of copper
to the overall toxicity of copper in a site water will depend upon the product
of the concentration of the chemical species times the toxicity that would be
exerted by the chemical species alone per unit concentration. Before describ-
ing the chemical speciation method in more detail, it will be necessary to
briefly discuss chemical speciation calculations and the concepts of toxicity
units and toxicity per unit concentration.
The major soluble inorganic species of copper in most natural waters are
Cu+2, CuOH+, and CuCO-° usually followed to a lesser extent by Cu(OH).* and
-2 +2
CuCCO^)^ (Section 1.1). If an LC50 value in terms of the Cu activity is
determined in a given type of test water from ion selective electrode measure-
+ 2 . .
ment of Cu activities, the concentrations of the major inorganic species at
the LC50 point can be calculated from the following equations:
^Cu 'lC50 = (Cu )LC50/yCu+2 (1-1)
[Cu(0H flLC50 = KCuOH VCu+2)LC50 (1~2)
[Cu(OH)2]LC50 =BCu(OH) Kw2(Cu+2)LC501°2PH/''cu(OH)i (1~3)
[CuC°3 1LC50 = KCuC03 (Cu ^LC50 (C°3 )/yCuC03 (1-4)
[Cu(C03)2 ]LC50 =BCu(C0 (Cu ^lc50 (c03 > /^^COg)^"2 (1-5)
where (
+ 2 +2
(Cu ^LC50 = point in tfcrms of the Cu activity at the LC50 point
Alkalinity - K 10P / yOH~ + 10~P /'V
(CO, ') = — d-6)
2/ y,
3 *-2 + 10 , y,
C03 ' a2 'HC03
( „ +, B„ » K. *, Bo /oo \ -2 = stability cons
CuOH ' Cu(OH)- CuC03 ' "Cu(C03)2
tants
1-9
-------
= ion product for water
K „ = acid dissociation constant for HCO,
a2 3
= activity coefficients
Likewise, the concentration of any other soluble raonomeric copper
species at the LC50 point, organic or inorganic, which can be represented as
+2
being formed from Cu can be calculated from the general equation
ICuL 2-"J]lc5o * kCuL. (C"*2)iLC50
-------
The toxicity of a copper solution can be expressed in terras of toxicity
units (70) where 1 toxicity unit is defined as the amount of toxicity require
to kill 50% of the test organisms during a 96 hr exposure. Therefore, by
definition, the toxicity of a copper solution at the LC50 point (e.g., a
copper solution in which 50% of test organisms will be killed during a 96 hr
exposure) is equal to one toxicity unit. The fractional contribution of any
given toxic copper species to the 1 toxic unit at the LC50 point will, assum-
ing additive toxicity, be equal to the product of the concentration of the
species at the LC50 point times the number or fraction of toxicity units that
the toxic species would exert if it was alone in solution at unit concentra-
tion (e.g., 1 molar). Therefore, assuming additive toxicity and that all
toxicity is due to dissolved copper species, the following equation should be
valid at the LC50 point for any type of test water (70).
1 = T1[Cu ]LC50 + t2[Cu0H 1LC50 + T3lCu(0H)2 1LC50 + T4 tCuC03 ^LC50 +
T.[Cu(C03)2 ]lc50 Tn(Cu.Lj J1Lc50 (1"9)
where
T^ = number or fraction of toxicity units which would be exerted by Cu+^ if it
was alone in solution at a unit concentration (e.g., toxicity/unit
concentration of Cu ^ or alternatively the inverse of the LC50 in terms
+ 2 +2
of Cu if all of the copper was in the form of Cu )
T2 = toxicity/unit concentration of CuOH
T^ = toxic it y/un it concentration of CutOH^'
T = toxicity/unit concentration of Cu.L.
. n ' i j
In theory the right side of equation (1-9) should include all souble
chemical species which contribute to the overall toxicity of the solution.
In practice, the left side of equation (1-9) can be approximated by including
on the right side only those species which might contribute significantly
(e.g. >1%) to toxicity.
+ 2
The chemical speciation method of estimating Cu ^53 values in site
waters depends upon first using multiple linear regression analyses to
1-11
-------
mine the values of the toxicity/unit concentration (e.g. T., T , T ... T )
i i- j n
exerted by all of the chemical species which contribute significantly to
toxicity. Substituting equations (1-1) to (1-7) into equation (1-9) and then
... +2
dividing both sides of the equation by the Cu activity gives an equation
which is more suitable for multiple linear regression analyses than equation
(1 -9 ).
1/Cu+2LC50 = T1 X1 * T2X2 + T3XT3 + T4X4 + Vs + " ' ' TnXn
(1-10)
where
X1 = " yCu
+ 2
,P»/
x, = K - K 10* /y_ n +
2 CuOH w CuOH
x. = B
Cu(OH)-
v a K •
4 Cu CO :
K2 102pH
Aik-KwiopH/yofi +iopH/yH+
x = B "2
5 Cu(C03)I
2/yco;2 * 1°"p'V12Vo3J
Aik-K iopH/y- + io~pH/y,+
w OH H
2/ycc -2 • io"pH/Ka2yHCO-
^Cu(C03)2*
x = K (L)J/7r T
n CuL. CuL.
J J
The toxicity/unit concentration (Tj, T21 ... T ) exerted by each of n
chemical species can be theoretically determined by
a) Determining the inverse of LC50 values in terms of the Cu 2 activity
for N>n types of test water for which the relative concentrations of
the n species change, but the toxicities/unit concentration of the n
species remain constant
b) Calculating the values of x. , x^, x^, ••• x of equation (1-10) from
the values of the pH and alkalinity at the Ec50 point for each of the
Ntest waters
c) Performing multiple linear regression of the inverse of the Cu LC50
on the variables x^, x^, x- ... x^ using the data from the N>n test
waters to determine the values of the toxicities/unit concentration
(T T T . . . T )
UT 2* 3* n
1-12
-------
If the toxicities/unit concentration (T., T„, T_, ... T ) can be deter-
+2 1 2 3 n
rained, the Cu LC50 in a given site water can be estimated by:
a) Determining the average site specific values of pH, alkalinity, and
hardness
b) Estimating the values of activity coefficients in the site water from
the substitution of the estimated site water ionic strength (based
primarily on site water alkalinity and hardness) into the Davies
equation (1-8)
c) Substituting the toxicities/unit concentration, the estimated activ-
ity coefficient, and the average site specific vali^s of pH and alka-
linity into equation (1-10) and solving for the Cu activity at the
LC50 point-
If the Cu ^ LC50 can be estimated for a site water, the total copper and
dissolved copper LC50s can be estimated by titrating a sample of the site
+2
water with copper until the activity of Cu as determined with a cupric ion
+ 2
electrode equals the estimated Cu LC50. The titrated sample can then be
analyzed by atomic absorbtion spectroscopy to determine the total and dis-
solved copper in the sample which will correspond to the estimated total and
dissolved copper LC50s for the site water.
1.5 USEFUL EQUATIONS FOR INTERPRETATION OF TOXICITY TEST DATA
Before discussing the results of the toxicity tests, it might be helpful
to define a few terms and to transform a more generalized form of equation 1-9
to a couple of other forms. Equation 1-9 equates the one toxicity unit at the
LC50 point to the sura of the product of the toxicity/unit concentration T^ of
each toxic species i times its concentration at the LC50 point ^^lc50 an(^
valid for additive toxicity. A more generalized form of equation (1-9) is
1 "Xv'i'ixso <1-"'
The product of the toxicity/unit concentration of each species i times its
concentration at the LC50 point is, assuming additive toxicity, equal to the
overall fractional contribution f^ of the particular species to the one
toxicity unit at the LC50 point
1-13
-------
fi - TilCilLC50 °-12)
Substituting equation (1-12) into (1-11) gives
1 -£f. (1-13)
Assuming additive toxicity, equation (1-13) is valid for any test water
at the LC50 point. Subtracting equation (1-13) for a test water A from
equation (1-13) for a test water B gives the following equation which may be
useful for comparing the changes in the fractional contribution of various
species to toxicity in going from any test water A to any test water B.
0 -Z^UB-A) U"14)
i
i
where by considering equation (1-12), it can be seen that
^fi(B-A) = TiB ^ Ci ^LC 50( B) ~ TiA^°i'LC50(A) (1-15)
If the toxicity/unit concentration of each species i remains constant in going
from test water A to B then
^fi(B-A) = Ti( [Ci'LC50(B) ~ [Ci1LC50(A)) (1-16)
and substituting equation (1-16) into (1-14) gives
0 -L'i^i'l-CSOU-A) ll"17)
An equation relating the change in the LC50 in going from test water A
to B, to changes in the sum over all toxic species of the product of the
toxicity/unit concentration times the proportion of dissolved copper repre-
sented by the species can be derived as follows. The concentration of any
copper species i at the LC50 point in any test solution is equal to the pro-
duct of the proportion of dissolved copper at the LC50 point that is repre-
sented by the species times the dissolved copper LC50:
1-14
-------
^CJlX50 ~ Pi(LC50) Dissolved Cu (1-16)
Substituting equation (1-18) into equation (1-11) gives:
1 = P. (LC50) = (LC50) ^T.P. (1-19)
i i
Rearrangement of equation (1-19) give:
LC50
(1-20)
Using equation (1-20) it can be seen that the difference in the dissolved
copper LC50 in going from test water A to B is given by:
(Tt. P. ~ P. )
. iA iA iB iB;
<4lc50)
-------
solution. The first grouped term of equation (1-22) represents the overall
contribution of differential proportional changes in toxic chemical species
to the differential change in the LC50 value. The second grouped term of
equation (1-22) represents the overall contribution of differential changes
in the toxic ities/unit concentration of toxic chemical species to the dif-
ferential change in the LC50 value.
It can be seen from the negative signs within the first and second
grouped terms of equation (1-22), that a differential decrease in the propor-
tion of a toxic species j will make a positive contribution to the differen-
tial change in the LC50 value, as will a differential decrease in the
toxicity/unit concentration of the species. Likewise, a differential increase
in the proportion of a toxic species j will make a negative contribution to
the differential change in the LC50 value, as will a differential increase in
the toxicity/unit concentration of the species. The sign and magnitude of the
differential change in the LC50 value will be equal to the sum of all positive
and negative differential contributions of all n toxic species j.
Although the above discussion concerned differential changes, the quali-
tative relationship between the signs of proportional and toxicity/unit con-
centration changes and the sign of a LC50 change should be valid for interval
changes as well.
1-16
-------
2. METHODS AND MATERIALS
2.1 GENERAL APPARATUS AND MATERIALS FOR Cu+2 ACTIVITY AND pH DETERMINATIONS
+ 2
All Cu activity determinations were made with an Orion cupric ion elec-
trode (model 94-29) referenced against an Orion Ag/AgCl double junction
reference electrode (model 90-02) (36, 71). The inner chamber filling solu-
tion used for the reference electrode was an Orion supplied AgCl solution
(90-00-02). The outer chamber filling solution used was an Orion supplied 10%
KNO^ solution (90-00-03). The pH measurements were obtained with either an
Orion pH glass electrode (910100) or a Fisher glass pH electrode (13-639-3)
referenced against the same Orion double junction reference electrode used to
+ 2
reference the cupric ion electrode. Both pH and Cu mV readings were
obtained from an Orion 801 meter attached to either a model 605 or 855 elec-
trode switch.- The use of an electrode switch allowed the cupric ion and pH
electrodes to be referenced against the same reference electrode simultane-
ously which has diagnostic advantages, and allowed alternate cupric ion
electrode mV and pH readings to be made by simple switch adjusment.
2.2 CUPRIC ION AND pH ELECTRODE CALIBRATIONS
Two point pH calibrations were performed daily in pH 7 and pH 10 buffers
supplied by Preiser for pH measurements in natural waters, and in pH 7 and ph
4 buffer supplied by MCB for pH measurements in acetate buffer.
The potential of an Orion cupric ion electrode/double junction reference
electrode can be theoretically related to the Cu 2 ion activity in solution by
the Nernst equation (72)
E = E' + 2.3 RT log (Cu+2) (2-1)
2F
where
E = potential reading in mV of the electrode pair
2-1
-------
E' = constant dependent upon the Ag /Ag° standard reduction potential,
copper sulfide and silver sulfide solubility products, the reference
electrode potential and any associated junction potentials
(Cu+^) = activity of Cu+^
2.3 RT/2F = "Nernstian slope" = 29.3 raV/log unit at 22°C
Although the slope is sometimes theoretically assumed to be equal to the
Nernstian slope, the value of E' cannot be theoretically estimated. In prac-
tice, both the slope and E1 (intercept) are determined experimentally through
calibration of the electrode. The calibration of the electrode consists of
+2 , +2
plotting mV readings versus Cu activity in solutions where the Cu activity
can be independently determined, and then performing linear regression on the
data to determine the experimental slope and intercept (E'). If the resultant
plot appears linear and the experimental slope is approximately equal to the
theoretical slope, the electrode is described as having a "Nernstian" slope or
as exhibiting "Nernstian" behavior.
The cupric ion electrode was calibrated in 0.01 M acetate buffer (36, 71,
79) consisting of equal molar amounts of Fisher reagent grade anhydrous sodium
acetate and high purity ("Ultrex") acetic acid purchased from Baker. The
cupric ion electrode was calibrated in acetate buffer (pH < 5.0) to maintain a
relatively constant pH with copper addition and to make the contribution of
CutOH^0 to the total copper negligible because the reported stability con-
stant values in the literature for that complex vary substantially. However,
there is very little disagreement between stability constant values reported
for copper acetate complexes.
The calibrations of the cupric ion electrode were carried out in an
initial volume of 200 ml acetate buffer in a 250 ml teflon beaker wrapped with
electrical tape to shield the light sensitive cupric ion electrode. The solu-
tions were stirred with a teflon stirring bar and the beaker was covered with
a commercial opaque teflon electrode holder. The calibrations were performed
-3 -1
by making 7-10 additions of a 10 M standard and 3 additions of a 10 M
+2
standard which gave a Cu activity calibration range from approximately 3 x
-8 -4
10 M to 10 M. After each addition of the CuNO^ standard, a stable mV
2-2
-------
reading was made followed inmediately by the withdrawal of 2 ml of the buffer
for total copper analysis by atomic absorption spectroscopy.
+ 2
The Cu activity in each 2 ml buffer sample was calculated from the pH
and total copper in the buffer using the following two equations:
+2 ^UT
(L/^Cu + 2) +(KCuAc (Ac")/*^CuAc)',"{BCuAc2 (Ac")2)
where
Act
1KHAc (formation) 10~pH) + (^Ac")
and
(Ac ) = acetate ion activity
Ac,j, = total acetate
KCuAc + ' ^Cu(Ac)2 = stability constants for copper acetate complexes
= stability constant for formation of acetic acid
Calibration curves were constructed from linear regressions on mV read-
+ 2
ings versus Cu activities. Calibration curves were almost always linear
-7 -t-2
with approximately Nernstian slopes down to 10 M Cu activity. There was
generally some scatter of points below 10 ^ M, but there were no discernible
+ 2 ...
non-linear trends down to the blank. The blank generally had Cu activities
_g
> 1.5 x 10 M corresponding in .01 M acetate buffer to the total copper
detection limit with the atomic absorption spectrometer. The scatter of
-7 +2
points below 10 M Cu activity was probably due to a combination of sample
contamination and increasing inaccuracy of atomic absorption spectroscopy
determinations below 10 ppb. The often reported non-linear behavior ;of the
+2 -6 +2
electrode at Cu activities below 10 M was observed only when mV vs Cu
+ 2 . .
activity curves were based on Cu activity determinations from nominal total
copper instead of total copper determined from atomic absorption spectroscopy.
+ 2
The pseudo non-linear behavior of the electrode when Cu activity deter-
minations are based on nominal total copper instead of actual total copper,
-8 +2
and the presence of >1.5 x 10 M Cu activities in blank acetate buffer may
2-3
-------
be due to oxidative dissolution of copper off the membrane as postulated by
several authors (73, 74, 75).
Calibration curves were very reproducible over many weeks. Figure 2-1
shows the typical reproducibility of calibration curves for two calibrations
performed over one month apart. The figure also demonstrates the difference
*2
between basing calibration curves on Cu activities determined from nominal
total copper instead of actual total copper (e.g., determined by atomic
absorption spectroscopy). The curves based on actual total copper are linear
with approximately Nernstian slopes down to the blank, whereas curves based on
nominal total copper are non-linear with positive curvature below a Cu+^
activity of approximately 5x10 ^M.
The slope and intercept of cupric ion electrode calibration curves were
+ 2
generally obtained from linear regression performed only on data with Cu
activities above 10 ^ M due to the usual scatter of data with Cu ^ activities
-7 +2
below 10 M. However, many of the Cu activities in copper toxicity test
-7 -9
solutions were below 10 M down to as low as 10 M. Therefore, it was
necessary to calculate those activities from linear extrapolations of the
calibration curves. Several authors have shown that the Orion cupric ion
electrode exhibits linear Nernstian behavior to well below 10 ^ M Cu+^
activity by using copper buffer solutions (76, 77, 78). The errors associated
with extrapolation due to uncertainties in the slope should have been reduced
by the performance of the linear regression on numerous data points (e.g., 8
or 9) with Cu+^ activities above 10 ' M.
+2
In retrospect, the accuracy of Cu determinations could have possibly
-7 -9
been improved by extending calibrations into the range of 10 to 10 M by
using a copper buffer instead of acetate pH buffer. However, electrode
response times in such regions are generally slow and errors introduced by the
use of a copper buffer (e.g., buffer complexing interferences with spike
recoveries with atomic absorption analyses) may offset any advantage gained by
not using extrapolation.
2-4
-------
170
160
150
40
130
20
10
100
90
• 9/50 Calibration bOMd or total copper
til/3 Calibration based on total copper
^ 11/3 Calibration bated on nominal copper
o 9/30 Calibration based on nominal copper
80
70
Blank
-LogCu*2
+ 2
Figure 2-1. A typical calibration curve for the cupric (Cu ) ion electrode
-------
2.3 Cu+2 ACTIVITY AND PH MEASUREMENTS IN FLOW-THROUGH SYSTEMS
All of Che laboratory toxicity tests were performed in flow through tanks
+2 . .
which will be described in Section 2.5. The Cu activity measurements in tne
flow-through tanks were performed with the use of an electrode chamber based
on a design by C. Stephan of the Duluth lab. The electrode chamber was
designed to hold the cupric ion, reference, and pH electrodes, shield the
cupric ion electrode from light, and to direct all flow in the tank through
the electrode chamber. The passage through the electrode chamber and the
placement of the cupric ion electrode were designed so that the membrane sur-
face of the electrode was continually impacted by the flow through the
chamber.
Since cupric ion electrode measurements require a certain minimal flow,
use of the flow through electrode chamber eliminated the necessity of stirring
or bubbling fcfaich can, in some cases, affect the pH of the test solution.
Furthermore, the use of a flow-through electrode chamber should have minimized
any errors associated with possible oxidative dissolution of Cu*2 off the
membrane surface (75). Comparisons of mV readings obtained with the electrode
chamber to mV readings obtained with a stirring system in Lake Superior water
+ 2
at various Cu activities were within 1 mV corresponding to a difference
between measured activities of <8%.
The cupric ion electrode was polished with Orion supplied polishing
strips every day before calibration or flow-through tank measurements were
taken to remove any surface irregularities or oxidized material (72, 73, 80)
from the electrode membrane. The electrode was then thoroughly rinsed with
deionized water or with .001 M EDTA/.01 M ascorbic acid (46, 71) and then
deionized water. However, rinsing with .001 M EDTA/.01 M ascorbic acid, which
is a mild copper reducing and complexing solution, did not appear to have any
beneficial affect when following polishing. That concurs with previous obser-
vations by others (77) although it appeared to improve the performance of
electrodes if they were not polished. The internal and external chambers of
the double junction reference electrode were also refilled daily before
measurements were taken.
2-6
-------
As previously discussed, two point pH calibrations were performed each
day prior to measurements. However, due to the reproducibility of cupric ion
electrode calibrations over several weeks (see previous section and Figure
2-1), full calibrations of the cupric ion electrode were generally performed
only just before, and just after, an entire set of measurements during a toxi-
+ 2
city test lasting 2 weeks. However, replicate measurements of Cu activities
were conducted over the test period in each test chamber in vrtnich measurements
were taken. If replicate measurements uniformly varied by over 2 mV in a
given diluter system, the cupric ion electrode was repolished, the inner and
outer chambers of the reference electrode were refilled, and measurements were
redone again. If large discrepancies between replicate measurements still
-5 +2
existed, the electrode performance was checked in a 10 M Cu standard.
The response time of the cupric ion electrode was dependent upon the
+2
activity of Cu being measured but was generally less than 1.0 hours. The mV
readings were taken when the chart recording indicated that mV readings were
level for at least 15 minutes, or when the average mV reading from the meter
did not change by greater than 0.1 mV over a 15 minute interval.
2.4 ATOMIC ABSORPTION SPECTROSCOPY
All total and dissolved copper measurements were determined by furnace
atomic absorption spectroscopy with a Perkin Elmer Model 5000 atomic absorp-
tion spectrometer equipped with either a model HGA 500 graphite furnace and
AS-40 auto sampler, or a model HGA 2200 graphite furnace and AS-1 auto
sampler. Total copper determinations were approximated by total acid
exchangeable determinations which involve only acidification prior to analyses
instead of more extensive digestion procedures, which have been found to be of
little or no value in analyses of Lake Superior water. Dissolved copper
determinations were performed by filtering the samples through a 0.45 mu
Millipore'filter prior to acidification and analyses. Quality assurance was
maintained with a minimum of one standard curve, two duplicate determinations,
and two spike recovery determinations for every 20 samples analyzed.
2-7
-------
2.5 FLOW-THROUGH EXPOSURE SYSTEMS
All laboratory toxicity tests other than the 96 hour static exploratory
tests were flow-through tests conducted on fathead minnow larvae. Six sets of
laboratory copper toxicity tests were completed. Each set of toxicity tests
consisted of determining copper toxicity in four types of test water simul-
taneously using four diluter systems, one for each type of test water. The
four diluter systems were designed and constructed to be as identical as
possible (81). Each diluter system consisted of two interconnecting stainless
steel headboxes, a glass diluter, randomized glass delivery tubes to exposure
tanks, and 24 glass flow-through exposure chambers.
The diluent test waters used in these experiments were Lake Superior
water and various types of chemically altered Lake Superior water. Filtra-
tion through 20 mesh screens, aeration, heating to 22C° and general chemical
alteration (other than copper addition) in the diluent test waters were
carried out in the headboxes. Chemical alterations of Lake Superior water
included changes in hardness, alkalinity, pH, suspended clay and humics.
Hardness was increased independently of alkalinity by pumping stock solutions
of MgCl^ and CaCl^ into the headboxes. Alkalinities were increased indepen-
dently of hardness by pumping a combination of NaHCO^ and KHCO^ into the
headboxes. The pH was decreased independently of alkalinity by bubbling CO^
into the headboxes. Increases in humic and suspended clay content were
obtained by pumping in suspensions of Lake Superior shore red clay and solu-
tions of hunics purchased from Aldrich. Neither increases in hunics nor
suspended clay affected hardness or alkalinity significantly, but did increase
pH 0.1-0.2 units above ambient Lake Superior water.
The diluent water flowed from the headboxes to the mixing chamber of the
diluter by gravity where it was mixed with CuSO^ stock that was dripped into
the mixing chamber after being pumped from a stock bottle to the diluter
mixing chamber with a FMI metering pump. The resultant copper solution then
underwent 4 successive approximately 1:1 dilutions with diluent water such
that each successive concentration was approximately 1/2 of the next highest
concentration. Therefore, the lowest concentration of the five was designed
2-8
-------
to be approximately 1/16 of the highest concentration. Each diluter continu-
ously delivered control test water and the five different concentrations of
copper in the test water to 24 glass flow-through exposure chambers. There-
fore, the control and each of five different copper concentrations were
replicated four time9 within a given diluter system. Each diluter delivered
solutions to the exposure chambers at approximately 15 ml/minute and main-
tained a volume of approximately 750 ml in each chamber corresponding to
retention times of approximately 50 minutes.
Each exposure chamber in each diluter system was designated with a number
indicating the relative concentration of copper in the chamber (e.g., 1 for
the highest concentration to 6 for the control) and a letter (A, B, C or D)
indicating which replicate delivery tube supplied the chamber from the given
concentration cell in the diluter. The A and B chambers were randomly located
in a row on one side of each diluter system, and the C and D chambers were
randomly located on the other side of each diluter system.
2.6 FLOW-THROUGH TOXICITY TESTS
Prior to the beginning of each set of copper toxicity tests on fathead
minnow larvae, eggs were obtained from the laboratory culture unit and accli-
mated to each type of test water for 5-6 days prior to hatching. During
acclimation, the eggs were placed in screened bottom (40 mesh) 120 ml glass
jars which were slowly oscillated up and down in control test waters within
the control exposure chambers. The screened bottcm glass jars were suspended
from rocker arras which rotated at approximately 2 revolutions/minute parallel
to, but above, the rows of exposure chambers. At the beginning of the copper
toxicity tests, the hatched larvae were also placed in screened bottom (40
mesh) 120 ml glass jars, 10 per jar. One jar with 10 larvae in it was placed
in each exposure chambers and again, slowly oscillated up and down in the
control and copper test waters within the chambers.
In the flow-through copper toxicity tests described in this report, 9b-hr
LC50s were determined from mortality data on the 10 larvae introduced to each
exposure chamber. The larvae in the A and B chambers of each diluter system
2-9
-------
were not fed, and the surviving larvae were terminated at the conclusion of
the 96 hr test . The larvae in the C and D chambers were fed with brine shrimp
and maintained for an additional three days after the conclusion of the 96 hr
acute te9ts to determine the affects of copper on 7 day growth. The Trimmed
Spearman-Karber computer program was used to calculate LC50 values for the
+2
unfed and fed tanks in terms of total copper, dissolved copper, and Cu
activity. At the conclusion of the 7 day growth tests, the surviving larvae
were terminated, dried in an oven, and weighed. An analysis of variance was
then performed on the dry weight data to determine any statistically signifi-
cant (e.g., >95% probability) differences in the dry weights between larvae in
control test water and larvae in test water with various levels of copper.
The hardness, alkalinity, pH, dissolved oxygen, conductivity and tempera-
ture were determined in each control test water every day during the 96 hr
acute tests. In addition, the same parameters were determined at least once
in the low, middle, and high copper concentration of each test water. Samples
for atomic absorption spectroscopy were withdrawn from chambers representing
every copper concentration in each test water at least once prior to the run-
ning of the toxicity tests, and at least twice during the 7 day test period.
+ 2
Duplicate sets of Cu activity measurements were made on different days in
tanks representing every copper concentration in each test water during the 7
day period, and additional sets were usually taken if large discrepancies
between duplicate sets were observed. In addition, duplicate measurements,
particularly in the middle concentration chambers close to the LC50, were made
within the same set of measurements on a given day. In addition to the ph
measurements determined with a combination electrode when general quality
parameters were determined, additional pH measurements were made with a single
glass electrode referenced against an Ag/AgCl double junction reference elec-
+ 2
trode in conjunction with each Cu determination.
2.7 Cd+2 ELECTRODE CALIBRATION AND Cd+2 ACTIVITY DETERMINATIONS IN FLOW
THROUGH EXPOSURE CHAMBERS
One set of laboratory flow-through cadmium tests was performed in addi-
tion to the six sets of flow-through copper toxicity tests. The set of
cadmium toxicity tests was performed primarily for support of previous cadmium
2-10
-------
toxicity tests conducted at the Duluth lab. Since the subject of this report
primarily concerns copper toxicity, the set of cadmium toxicity tests will not
be described in detail in this report. However, the objective of the set of
cadmium toxicity tests was to determine the effects of various humic levels on
metal toxicity, which is identical to the objective of set 5 of the copper
toxicity experiments. Therefore, the results of the cadmium toxicity tests
will be briefly compared to the results of set 5 of the copper toxicity tests
in Section 3.5.
The set of cadmium toxicity tests consisted of determining cadmium tox-
icity in four types of test water simultaneously using four diluter systems,
one for each type of test water. The diluter systems used were the same ones
used in copper toxicity tests, and are described in Section 2.5. The general
biological and chemical procedures used for the cadmium tests were also almost
identical to those described in Section 2.6 for copper. However, there are
three major differences between the cadmium and copper procedures.
The first major difference involves electrode calibration. Although both
+ 2 +2
the Cu and Cd electrodes were calibrated in .01M acetate buffer, they
exhibit different behavior at low activities. The apparent non-Nernstian
+ 2 + 2
behavior of the Cu > electrode at low Cu , activities appears to be due to
copper dissolution off the membrane and not to an actual non-Nernstian re-
+ 2
sponse of the electrode, as was discussed in Section 2.2. However, the Cd
+ 2
electrode appears to actually exhibit non-Nernstian behavior at lower Cd
+2
activities, as can be seen by the Cd calibration curve presented in Figure
(2-2). Even though the calibration curve is based on total rather than
nominal cadmium, the calibration curve becomes non-Nernstian at cadmium
activities below 3 x 10 ' M.
The non-Nernstian response of the Cd ^ electrode has at least two detri-
+2 .
mental effects on the determination of Cd activities. The determination of
+ 2
Cd activities in the non-Nernstian response range of the electrode is sub-
ject to greater error than in the Nernstian range due to the decrease in the
slope of the electrode response versus the log of the activity in the non-
Nernstian range. Also, the limit of detection of the cadmium electrode caused
2-11
-------
-200
-210
o 1/16
-220
A 2/6
V 3/7
-230
-240
-250
-260
-270
-230
-290
7
6
,+2
-Log Cd
+ 2
Figure 2-2. The calibration curve for the cadmium (Cd ) ion electrode
2-12
-------
by the flattening of the calibration curve at low-activities makes the deter-
mination of some of the lower Cd+^ activities in some of the toxicity tests
+2
impossible. Since the non-Nernstian response of the Cd electrode is actual
instead of apparent due to any Cd+^ dissolution off the membrance, the use of
metal buffers for calibration would probably not extend the Nernstian response
range of the electrode, nor lower the limit of detection.
The second major difference between copper and cadmium procedures
involves the way in which electrode readings were taken in the flow-through
+2
chambers. Although a flow-through electrode chamber was used to perform Cu
electrode readings as discussed in Section 2.3, the flow supplied by the
+2
electrode chamber was insufficient for use with the Cd electrode. There-
~ 2
fore, flow was supplied to the Cd electrode by bubbling air through the
solution with the use of an air stone positioned below, and directed at, the
electrode membrane surface.
The third major difference between copper and cadmium electrode proce-
dures involves the way in which the electrode membrance surface was cleaned.
Whereas the use of Orion supplied electrode polishing strips improved the
+ 2
performance of the Cu electrode, it was detrimental to the performance of
+ 2 +2
the Cd electrode leading to high and often unstable mV readings. The Cd
electrode membrane surface was therefore cleaned with tissue paper.
2.8 CONNECTICUT FIELD STUDY ON THE NAUGATUCK RIVER
+2
Static 96 hour LC50s in terras of total, dissolved, and free copper (Cu )
were determined for one day old fathead minnow larvae in water samples taken
from several different sites along the Naugatuck River in Connecticut. The
tests were conducted as part of a much larger study to determine the applic-
ability of site specific criteria derived in relatively unpolluted upstream
waters to site water downstream containing industrial and/or municipal sewage
+ 2
effluents. The Cu LC50 determinations were made so that they could be
+ 2
compared to any future theoretical estimates of Cu LC50 values if a suitable
chemical speciation methodology can be developed. Since R. Carlson of the EPA
lab in Duluth is preparing a report on the Connecticut field study, the study
2-13
-------
will not be described in detail in this report. However, the results of the
tests in which Cu+^ LC50s were determined will be briefly discussed in Section
3. 7.
Static tests were run in Lake Superior water for comparative purposes,
water taken from a relatively unpolluted upstream site (site Nl), and water
taken from several downstream sites containing industrial and/or municipal
sewage effluents (N4-A, N5, N6, N7). Figure 2-3 shows the location of the
various sampling sites along the Naugatuck River.
The static tests were run in 1000 ml plastic test chambers in 700 ml of
test water at 25°C. Duplicate test chambers, each containing 10 organisms,
were used for each copper concentration and type of test water. Also, an
additional test chamber containing the test water but no organisms, was sec up
for each copper concentration and type of test water so that electrode deter-
minations of Cu+^ activities could be performed without disturbing the test
organi sms.
Samples for atomic absorption analyses of total and dissolved copper, and
for pH, hardness, and alkalinity determinations were taken both prior to the
initiation of, and after the termination of, the 96 hour acute tests. The
Cu+^ activity for each nominal copper concentration and type of test water was
determined in water from the exposure tanks without organisms within the first
+ 2
2.4 hours of each toxicity test. Additional Cu activities were also deter-
mined in water from the exposure chambers with organisms after the termination
~ 2
of the 96 hr acute toxicity tests. However, due to time limitations, Cu
activity determinations after the termination of the toxicity tests were per-
formed only in water from exposure chambers bracketing the dissolved copper
LC50 point.
The Cu+^ activity determinations were performed as follows. Water from a
given exposure chamber was poured into a 500 ml polyethylene bottle to the
top of the bottle, and then closed to the atmosphere with a screw polyethylene
top into which a cupric ion selective electrode, double junction reference
electrode and pH electrode had been previously tightly fitted. The sample was
2-14
-------
NAUGATUCK RIVER
TOI
TUftNCN ft
SCYMOUR
HJ
¦3.4-To -•Yh* STP
^ WA A TM0MAiT0M
% \p° . ]
\ H^STP J
Mt
SB I
MONlTOWIMg LOCATIONS
O SlOLOOICAL
~ CHEMICAL/ PHYSICAL
X
-» DISCHARGE SITES NCCOMMKNOD FOR STUOY
AMftOI. SCALE : I INCH* 10 KILOMETERS
Figure ,2-3 Sampling sites along the Naugatuck River
2-15
-------
closed to the atmosphere with minimal head space to prevent CO^ exchange and
the associated pH changes from occurring during the stirring required for the
+ 2
Cu activity determinations. The sample was then stirred with a Teflon bar,
and mV readings were recorded until they leveled off to changes of no more
than O.lmV in a 5 minute period.
2.9 STEADY STATE/EQUILIBRIUM COMPARISONS
A number of tests were performed to determine how close the steady states
in the flow-through toxicity tests were to equilibrium. Tests were performed
on Lake Superior water and Lake Superior water with elevated pH and alkalin-
ity, lowered pH and alkalinity, and added hunic acid. The tests were per-
formed on test waters with nominal copper concentrations comparable to pre-
viously determined LC50 values in similar types of water.
The tests were conducted as follows. A sample was withdrawn from the
flowthrough chamber for later analyses for total copper by atomic absorption
analysis. The pH and Cu+^ activity were determined in a duplicate flow-
through chamber by the same method described in section 2.3. Water was then
poured from the exposure chamber into a 500 ml wide mouth Teflon bottle to the
top of the bottle and then closed to the atmosphere with a screw polyethylene
cap into which a Cu+^ ion selective electrode, and a double junction reference
electrode had been previously tightly fitted. The sample was then stirred
with a Teflon bar, and the mV readings from the cupric ion selective electrode
were recorded every 5 minutes for the first 15 minutes, every 15 minutes for
the next 45 minutes and every 30 minutes thereafter for at least 6 hours or
until the readings leveled out.
At the termination of the Cu ^ electrode mV readings, a sample was taken
from the bottle for analysis by atomic absorption for total copper. That was
done so that any decrease in mV readings due to copper absorption could be
differentiated from decreases due to the difference between the flow-through
steady state and the closed system equilibrium. After the sample was taken,
the pH of the remaining solution in the Teflon bottle was determined to make
sure that the stirring during the experiment had not led to any significant
2-16
-------
~2
CC>2 exchange and associated pH change which would have affected the Cu
activity in addition to any differences between the flow-through steady state
and the closed system equilibrium.
2.10 COPPER TITRATIONS OF LAKE SUPERIOR WATER AND RECONSTITUTED WATER
Two sets of titration experiments were performed to determine if copper-
organic complexation was primarily responsible for the magnitude of the
observed greater than Nernstian slopes of the cupric ion selective electrode
response in Lake Superior water. The first set of experiments was designed to
observe the effects of reduced organic content on the slope of the electrode
response. The second set of experiments was designed to observe the effects
+ 2
of ionic strength, cation competition with Cu for organic ligands, alkalin-
ity, and pH on the slope of the electrode response.
The first set of experiments consisted of 10 \ CuCNO^^ titrations of
Lake Superior water, UV irradiated Lake Superior water, Burdick & Jackson HPLC
grade water, UV irradiated Burdick & Jackson water, and Burdick & Jackson
water passed through a C-18 reverse phase HPLC column. The alkalinities of
the Burdick & Jackson waters were raised to that of ambient lake water by
adding either NajCO^ or NaHCO^. The second set of experiments consisted of
10 3M CuNO titrations of Lake Superior water and seperate Lake Superior
-3
waters with the following nominal concentrations of added chemicals: 2x10 M
Ca(N0^)2> 10"4M Zn(N03)2, 6xl0~3M Na NO^ and 4xl0~3NaHCO3. Titrations of the
4x10 M NaHCO^ water were performed at pHs of 8.00 and 8.50.
The equipment and procedures used for the titrations were similar to
those discussed in Sections 2.1 and 2.2 for the electrode calibrations. The
major difference was that the titrations were performed primarily in solutions
with a low pH buffering capacity. Changes in pH during a titration can con-
tribute to the non-Nernstian slope, as will be discussed in Section 3.9.
Since we were interested in differentiating the contribution of copper-organic
complexes from that of pH changes to the non-Nerstian slope, and since the pH
in equilibrium with the atmosphere can vary during the course of the titra-
tion, the titrations were performed at constant pH as follows.
2-17
-------
The Burdick & Jackson waters to which Na^CO^ was added to approximate
ambient lake water alkalinity and lake waters with added NaHCO^ were bubbled
with CC>2 until the pH of the solution was below 8.00. The ambient lake water
and other chemically altered lake water already had pHs below 8.00 due to
supersaturation with C^. If the pH of the solution was below 7.95, it was
then bubbled with air until the pH was 7.95 and then stirred until the pH was
+2
8.00. When the pH was 8.00, the mV reading related to the Cu activity in
the blank was taken, a sample was withdrawn for atomic absorption of total
-3
copper, and an aliquot of 10 M CuNO^ was added. The addition of the CuNO^
would drop the pH below 8.00 again. After stirring slowly brought the pH back
to 8.00, another mV reading was taken, another sample was withdrawn for total
copper analyses, and another 10 CuNO^ addition was made. The same proce-
dure was repeated until the titration was terminated. All of the constant pH
titrations were performed at a pH of 8.00 except one at a pH of 8.50.
2.11 CHEMICAL SPECIATION CALCULATIONS
The chemical speciation calculations in this report are based primarily
+ 2
on the substitution of Cu LC50s, derived from cupric ion selective deter-
+ 2
ruinations of Cu activities, into equations (1-1 to 1-7). In addition, for
comparative purposes, the calculations were also performed by the input of
dissolved copper LC50s into the REDEQL chemical equilibrium program. The
REDEQL program was also used to model the effect of the amino acid glycine on
chemical speciation.
+ — 2
All equilibrium calculations of CuOH , CuCOH)^0, CuCO^" and CuCCO^^ i-n
this report, including those using REDEQL, were performed using the stability
constant values experimentally determined by Sunda and Hansen (35) with a
cupric ion selective electrode. The stability constants are for 25°C and zero
(by correction) ionic strength. No standard enthalpy of formation values for
the copper hydroxy or carbonate complexes could be found in the literature, so
temperature corrections of the stability constants to the experimental temper-
ature in the flow-through systems of 22®C were not possible. However, approx-
imate values of the ion product of water K and the acid dissociation constant
-2 o W
for HCO^ going to CO^ at 22°C were used in the calculations based on
graphical interpolation between values at several other temperatures.
2-18
-------
The stability constants of formation reported by Sunda and Hansen (35)
+ -2
for CuOH , CuCO^' and (^(COj)^ are very close to those listed in Martel and
Smith (58). However, there are major disagreements in the literature over the
magnitude of the stability constant for the formation of Cu(OH) ° with
4
reported values differing by as much s 10 (35). Although there have been
several methods used to determine the CutOH)^ stability constant, the method
using the cupric ion selective electrode appears to be the most direct.
There have been three values of the CuOH)^0 stability constant deter-
mined with a cupric ion selective electrode (35,66,81), all of which have been
reported subsequent to the most recent listings of stability constants by
Martel and Smith (58). The value of the CutOH)^0 stability constant reported
by Sunda and Hanson (35) is very close to that reported by Paulson (81), but
is more than three orders of magnitude less than that reported by Vuceta and
Morgan (66). We have decided to use the value of the stability constant
reported by Sunda and Hanson (35) for the following reasons:
1. The reported value is close to that reported, by Paulson (81).
2. Both Paulson (81) and Sunda and Hanson (35) corrected for copper
absorption vAiereas Vuceta and Morgan (66) did not.
+2
3. The agreement between oyjj electrode determination of Cu activities
and REDEQL predicted Cu activities in systems with higher than
ambient lake water alkalinity and/or pH is greater using the constant
reported by Sunda and Hanson than that using the constant originally
listed in the REDEQL data base or the constant reported by Vuceta and
Morgan (66).
The activity coefficient values used in equations (1-1 to 1-6) were
determined from the substitution of the estimated ionic strength for each test
water into the Davies equation (1-8). Estimates of the ionic strength of each
test water were made from the following equation based on the assunption that
most of the alkalinity of natural freshwaters below a pH of 8.5 is due to
HC0~:
I = (2x10 ^) (Hardness in mg/L as CaCO^ + l/^ (Alkalinity in
eq/L)+l/2(7xlO_5)^l/2(2xlO~A) ~ W2^C.Z^
2-19
-------
where
I = estimated ionic strength
(7x10 = sum of the average concentrations of Na+ and K+ in Lake
Superior water (57).
-4 _
(2x10 ) = sum of the average concentrations of Cl^ and NO^ and four
times the average concentration of SO, in Lake Superior
water (57).
^C. Z? «= sum of the product of the concentration of any added cations
i and anions above the average concentration in ambient Lake
Superior water times the ion charge squared.
2-20
-------
3. RESULTS AND DISCUSSION
Six sets of flow-through copper toxicity tests, one set of flow-through
cadmium toxicity tests, and a field study on the Naugatuck River in
Connecticut were completed. Each set of flow-through toxicity tests consisted
of 96 hr LC50 determinations and 7 day growth studies in four types of test
water. The results of the LC50 determinations and related speciation calcu-
lations for the unfed tanks are summarized for every test water in each set in
+2
Tables 3-1, 3-2, and 3-3. The results of the one set of Cd toxicity tests
are compared to those of set 5 of the Cu* toxicity tests in Section 3.5.
The LC50 determinations for the fed tanks in terms of total, dissolved
+2
and free copper (Cu ) were very close to that of the unfed tanks, which indi-
cates that the absorption of copper onto the food and the effects of feeding
. . +2
were probably negligible. However, because most of the Cu determinations
were performed in the unfed tanks and because feeding is not a standard pro-
cedure for 96 hr acute tests, the chemical speciation calculations are based
on the unfed tanks. The LC50 values in terms of total, dissolved, and free
copper (Cu for the static field 96 hour acute tests performed on samples
from the Naugatuck River are presented in Table 3-4. The sites along the
Naugatuck River from which samples were taken were shown previously in figure
2-3.
Table 3-1 lists for each test water in each set, by column, any chemical
constituent added to ambient Lake Superior water, the average ph, alkalinity,
and hardness of the test water, and the LC50 determinations for the test water
+ 2 +2
in terms of total copper, dissolved copper, Cu activity and Cu concentra-
tion. The table lists tv*> mean pH values for each test water. The first
value in each case is the mean pH determined with a combination pH electrode
that was calibrated in one buffer at pH 7. The second value in each case is
the mean pH value determined with a single glass pH electrode (Fisher or
Orion) referenced against a double junction Ag/AgCl electrode and calibrated
using two buffers (pH 7 and pH 10). The discrepancies between mean pH values
are usually < 0.2 pH units, but are large enough to significantly affect some
of the speciation calculations at higher pHs as shown in Tables 3-2 and 3-3.
3-1
-------
Table 3-2 lists for each test water in each set, by column, any chemical
constituent added to the ambient Lake Superior water, the mean pH of the test
water, the LC50 in terms of dissolved Cu in moles/L, the LC50 value in terms
+2
of the Cu concentration in moles/L and the corresponding calculated concen-
+ ° ° -2
trations of CuOH , CuCOH)^ , CuCO^ > Cu( > an<* Cu-Org in moles/L where
Cu-Org stands for a sun of the concentrations of all theoretical copperorganic
or other unknown complexes. The calculated concentrations of the inorganic
copper species were determined by substituting in the experimentally deter-
+2
mined pH, alkalinity, and Cu activity at the LC50 point into equations (1-1)
through (1-7). The sun of theoretical copper-organic or other unknown com-
plexes represented by Cu-Org was calculated from the difference between the
dissolved copper and the sun of all major inorganic copper species at the LC50
point. A value of zero means that the sum of all major inorganic species was
equal to or greater than the dissolved copper. Table 3-2 lists three sets of
calculated concentrations for each test water, one from substituting the com-
bination electrode mean pH into equations (1-2) to (1-7), one from substi-
tuting in the single glass electrode mean pH values, and one from substituting
in the average of the two mean pH values.
Table 3-3 lists for each test water in each set, by column, any chemical
constituent added to the ambient Lake Superior water to form the test water,
the mean pH of the test water, the LC50 in terms of dissolved copper in mol/L,
the calculated percentages of dissolved copper at the LC50 point contributed
+2 + 0 -2
by Cu , CuOH , CutOH^ , CuCO^°, and CutCO^)^ > respectively, the sum of
those inorganic percentages, and the possible Cu-Org percentage calculated
from the difference between 1002 and the sum of inorganic percentages. Again,
as in Table 3-2, there are three sets of data for each test water correspond-
ing to calculations based on pH values determined with the combination pH
electrode, pH values determined with the single glass electrode, and averages
of the mean combination and single glass electrode pH values.
A discussion of the results of the six sets of copper toxicity tests
concluded is given below. Unless otherwise stated, all discussion concerning
chemical speciation will be based on calculations which used the average of
the combination pH electrode and single glass electrode mean pH values (e.g.,
the bottom pH for each test water in Tables 3-2 and 3-3).
3-2
-------
TABLE 3-1. Gener^ water quality characteristics and 96 hr LC50 values in terms of total copper,
dissolved copper, Cu activity and Cu concentration for the test waters of each set of copper toxicity
experiments. The parentheses show 95% confidence intervals.
PH
Electr. LC50 LC50 LC^U LCJO
Comb. vs Alk. Hard. Tot. Cu Diss. Cu (Cu ) |Cu |
Elect. Kef. mg/L mg/L ug/L ug/L ug/L ug/L
Set 1 — 4/83 (Ca*^, Mg*^, Na* Affects)
Ambient Hardness 8.10 8.16 42.6 47.0 32.0 28.0 0.4tt 0.57
_3 (25-41) (22-36) (0.3-0.7)
2 x 10 M CaCI 8.01 8.09 43.1 243 119 109 3.64 5.42
2
'2
2 x 10 3 MgCl_ 8.01 8.10 43.3 255 48.0 40.0 1.13 1.64
-3
2 x 10 NaCl 8.10 8.14 43.3 47.2 75-0 70.0 2.85 3.70
Set 2 — 6/83 (pH Effects at Ambient Alkalinity)
(89-166) (82-152) (2.5-5.2)
48.0 40.0 1.13
(37-63) (31-53) (0.8-1.6)
75-0 70.0 2.85
(50-111) (47-104) (2.0-4.0)
CO Bubbling 6.63 6.53 48.9 49.2 8.0 7.0 0.79 0.95
(6.9-8.5) (6.0-7.4) (0.6-1.1)
CO Bubbling 7.30 7.40 44.5 46.2 23.0 19.0 l.bO 1.90
(18-29) (15-24) (1.1-2.2)
Ambient pH 8.02 8.10 42.8 45.1 60.0 50.0 1.67 1.99
(42-86) (35-72) (1.0-2.1)
NaOH Addition 8.65 8.81 47.4 45.2 82.0 71.0 0.45 0.5b
(70-120) (61-104) (0.3-0.6)
Set 3 — 7/83 (pH effects at 3X Ambient Alkalinity)
(K , Na ) HCO , CO 7.14 7.16 155 46.2 22.0 19.0 1.64 2.14
(17-28) (15-24) (1.3-2.1)
-------
Table 3-1 (continued)
PH
Elec t r .
Comb. vs Alk. Hard.
Elect. Ref. mg/L mg/L
(K+, Na) HC03> C02 7.90
(K+, Na) HC03, C02 8.50
HC1 Added to Lower pH 7.16
and Alkalin i ty
Set 4 — 10/83 (Effects of Clay,
OJ ¦ - — ¦¦ '¦ — — ¦ ¦ i ¦ f.
I
Ambient Lake Water 7.93
Clay Added (70 NTU) 8.12
Humics (5 mg/L TOC) 7.91
Clay and Humics Added 8.17
Set 5 -- 11/83 (Effects of Variable Humic Concentrations
Ambient Lake Water
7.94
8.12
43.0
45.0
Humics (1.25 mg/L TOC)
7.91
8.11
42.2
45.0
Humics (2.5 mg/L TOC)
7.94
8.06
43.8
45.5
Humics (5.0 mg/L TOC)
7.95
8.11
42 .8
45.0
7.99 148 45.0
8.64 150 45.0
7.17 26.4 45.0
> unties)
8.15 40.8 42.5
8.36 46.5 48.0
8.10 44.2 45.5
8.28 48.5 46.8
LC50 LC50 LC^y L(J$y
iTot. Cu Diss. Cu (Cu ) ICu |
ug/L ug/L ug/L ug/L
79.0
(53-116)
157
(99-249)
22.0
(16-31)
56.0
(45-98)
136
(86-216)
21 .0
(15-30)
1 .00
(0.7-1.5)
0.61
(0.4-1.0)
3.64
(2.4-5.5)
1 .30
0.79
4.20
105
(69-163)
149
(117-191)
289
(229-367)
430
(313-593)
87.0
(57-135)
66.0
(52-84)
249
(197-317)
254
(185-350)
2.75
(2.0-4.3)
1.11
(0.9-1.4)
1 .04
(0.8-1.5)
0.96
(0.7-1.5)
3.25
1 .33
1 .24
1 .14
83.0
(64-107)
132
(113-153)
244
(208-326)
298
(238-373)
74.0
(57-96)
102
(87-117)
198
(169-264)
245
(196-306)
2.6 8
1-3.5)
1 .t>8
b-2.3)
1 .86
5-2.9)
0.96
(0.8-1.5)
(2.
(1.
(I .
i . IB
2.23
2.24
I .14
-------
Table 3-1 (continued)
E»
Electr.
Comb. vs AIk.
Elect . Ref. mg/L
Set 6 -- 5/84 (Effects of Variable Alkalinities)
HCL Added, (»2 reduced 7.84 7.84 17.0
Ambient Lake Water 7.87 7.99 42.0
(K*. Na+) HC03, C02 7.92 8.04 161
(K+, Na+) HC03, C02 7.96 8.04 318
LC50
Hard. Tot. Cu
mg/L ug/L
44.0
68.0
46.0
4.51
5.2/
(58-80)
(30-54)
(3.b8-5.52)
44.0
95.0
79.0
3.33
3.95
(75-119)
(62-99)
(2.54-4.36)
44.0
51 .0
39.0
0.51
0.67
(32-83)
(25-64)
(0.33-0.80)
45.0
66.0
46.0
0.29
0.42
(57-77)
(40-54)
(0.23-0.34)
LC50 LC^O LC^U
Di ss. Cu (Cu ) ICu J
ug/L ug/L ug/L
-------
I
+2
Table 3-2. The average pH, 96 hr LC50 in terms of dissolved copper (mol|s/L), 96 for LC50 in terms of2Cu concentration
(moles/L), and corresponding calculated concentrations (moles/L) of CuOH , Cu(OH)2 , CuCO^, CuCCO^^ and Cu-Org for the test
waters of each set of copper toxicity experiments.
pH
LC50
Diss. Cu
9 -1
x 10 M
LC*9
(Cu ]
9 -1
x 10 M
LC50+
[CuOH ]
9 -1
x 10 M
LC50
[Cu(OH)^ °
9 -1
x 10 M
LC50
(CuC03°l
9 -1
x 10 M
LC50
|Cu(C03)2 J
9 -1
x 10 M
LC50
ICu-Org1
9 -1
x 10 M
Set 1 — 4/83 (Ca*^, Mg*^, Na* Effects)
Ambient Hardness
2 x 10 3 M CaCl,
CO
I
2 X 10 J M MgCl2
2 x 10 3 M NaCl
8.10
8.16
8.13
8.01
8.09
8.05
8.01
8.10
8.05
8.10
8.14
8.12
441
441
441
17 20
1720
1720
630
630
630
1110
1110
1110
7.55
7.55
7.55
85.3
85.3
85.3
25.8
25.8
25.8
58.3
58.3
58.3
23.9
27.4
25.5
155
186
168
48.1
59.1
53.0
145
160
152
4.61
6.07
5.26
23.1
33.4
27.5
7.18
10.9
8.73
27.4
33.0
30.0
180
206
193
1070
1290
1170
335
411
368
1060
1160
1110
2.92
3.84
3.36
17.1
24.7
20.4
5.25
7.92
6.36
18.8
22.5
20.5
221
189
205
368
98.2
248
208
115
168
ob
0b
0b
Set 2 — 6/83 (pH Effects at Ambient Alkalinity)
C02 Bubbling
C02 Bubbling
6.63
6.53
6.58
7.30
7.40
7.35
110
110
110
299
299
299
12.4
12.4
12.4
30.0
30.0
30.0
1.34
1.06
1.18
12.6
16.0
14.1
0.0087
0.0055
0.0068
0.386
0.612
0.479
10.2
8.05
8.87
101
127
111
0.0057
0.0036
0.0044
0.279
0.436
0.334
83.6
86.0
85.0
155
125
143
Ambient pH
8.02
8.10
8.06
787
787
787
31.3
31.3
31.3
69.3
83.3
75.6
11.1
16.1
13.2
525
631
567
7.18
10.3
8.38
14.2
14.6
91.3
-------
Table 3-2 (continued)
LC50
Diss. Cu
a 9-1
pH x 10 M
NaOH Added 8.65 1120 7.08
8.81 1120 7.08
8.72 1120 7.08
Set 3 — 7/83 (pH Effects at 3X Ambient Alkalinity)
KHCO- and NaHC03
Added, CO„
^ Bubbling
KHCO» and NaHCO
Added, C02
Bubbling
KHC03 and NallC03
Added, CO.
Bubbling
HC1 Added
to Lower pH
and Alkalinity
Set 4 — 10/83 (Effects of Humlcs and Suspended Clay)
Ambient Lake 7.93 1370 51.2
Superior Water 8.15 1370 51.2
8.03 1370 51.2
Clay Added 8.12 1040 20.9
to Turbidity 8.36 1040 20.9
of 70 NTU 8.22 1040 20.9
9 -1
x 10 M
7. 14
7.16
7.15
7.90
7.99
7.94
8.50
8.64
8.56
7.16
7.17
7.16
299
299
299
1050
1050
1050
2140
2140
2140
331
331
331
33.6
33.6
33.6
15.7
15.7
15.7
12.5
12.5
12.5
67.4
67.4
67.4
LC50 LC50 LC50 LC50 LC50
[CuOH+] [Cu(0H)2°] |CuC03°l [Cu(C03)2 1 lCu-Org]
9-1 9-1 9-1 9-1 9 -
10 M x 10 M x 10 M x 10 M x 10 M
79.6
115
93.9
54.3
113
75.5
646
915
755
40.5
81.0
55.4
291
0b
131
9. 17
9.59
9.37
0. 189
0.207
0.198
243
2 54
248
1.71
1.88
1.79
11.3
0.00
6.03
32.1
39.6
35.4
3.81
5.81
4.65
805
993
889
30.8
46.9
37.7
160
oa
62.8
78.1
108
90.5
37.0
70.5
49.6
1940
2650
2240
294
547
390
0.
0
20.8
21.3
21.0
0.461
0.483
0.471
98.3
101
99.6
0.114
0.120
0.117
144
141
142
93.0
1 54
116
12.2
33.3
18.9
676
1110
838
7.20
19.3
11.1
530
0.00
330
58. 1
101
73.8
11.7
35.3
18.8
475
816
603
8.8
26.2
14.2
457
39.5
309
-------
Table 3-2 (continued)
u a
pH
LC50
Diss. Cu
9 -1
x 10 M
LC^
[Cu 1
9 -1
x 10 M
LC50
ICuOH J
9 -1
x 10 M
LC50
lCu(OH)2°]
9 -1
x 10 M
LC50
lCuC03°]
9 -1
x 10 M
LC50
[Cu(C03)2 ]
9 -1
x 10 M
LC50
(Cu-Orgl
9 -1
x 10 M
Humlcs Added
7.91
3920
19.5
33.6
4.17
262
2.88
3630
( 5 mg/L TOC)
8.10
3920
19.5
52.0
10.0
406
6.89
3430
8.00
3920
19.5
40.9
6.19
319
4.27
3540
Humics and Clay
8.17
3860
18.1
56.3
12.7
481
10.5
3280
Added (5 mg/L
8.28
3860
18.1
72.5
21.1
617
17.3
3120
TOC, 70 NTU)
8.22
3860
18.1
63.3
16.1
538
13.2
3210
Set 5 — (Variable Humlc Concentrations)
Ambient Lake
7.94
1160
50.1
92.3
12.3
703
8.00
293
do Superior Water
8. 12
1160
50.1
140
28.2
1060
18.3
0a
8.02
1160
50.1
112
17.8
847
11.6
131
Humics Added
7.91
1610
35.2
60.6
7.53
4 54
4.76
1050
( 1.25 mg/L
8.11
1610
35.2
95.9
18.9
713
22.7
736
TOC)
8.00
1610
35.2
74.5
11.4
557
7.15
924
Humics Added
7.94
3120
35.2
64.8
8.62
503
5.86
2 500
( 2.50 mg/L
8.06
3120
35.2
85.6
15.0
661
10.1
2310
TOC)
8.00
3120
35.2
73.8
11.2
571
7.56
2420
Humics Added
7.95
3860
18.0
33.9
4.63
257
2.98
3540
( 5.0 mg/L
8.11
3860
18.0
49.1
9.65
369
6.16
3410
TOC)
8.02
3860
18.0
40.1
6.47
303
4.16
3490
Set 6 -- 5/84 (Alkali
nity Effects
at Ambient pH)
HCL Added,
7.84
724
82.9
123
13.0
374
1.32
130
CO- Reduced
7.84
724
82.9
123
13.0
374
1.32
130
z
7.84
724
82.9
123
13.0
374
1.32
130
-------
Table 3-2 (continued)
LC50 LC5Q LC50 LCSO LC50 LC50 LC50
Diss. Cu [Cu ) [CuOH ) (Cu(OH)2°) [CuC03°] (Cu(C03)2 ] [Cu-Org]
a 9-1 9-1 9-1 9-1 9-1 9-1 9-1
pH x 10 M x 10 M x 10 M x 10 M x 10 M x 10 M x 10 M
Ambient Lake
7.87
1240
62.2
97.7
11.1
729
6.93
333
Water
7.99
1240
62.2
129
19.4
965
12.1
52.1
7.93
1240
62.2
111
14.2
827
8.91
217
(K+, Na+)HC03, C02
7.96
724
6.54
10.9
1.45
562
57.1
86.2
Added to
8. OA
724
6.54
13.2
2.11
677
82.9
0a
6x Ambient
8.00
724
6.54
11.9
1.72
609
67.1
27.5
Alkalinity
CO
I
For each test water, the top pH listed is the average pH value determined with a combination pH electrode that was calibrated in
one buffer at pH 7. The middle pH is the average pH value determined with a single glass electrode referenced against a double
junction Ag/AgCl reference electrode and calibrated in two buffers (pH 7 and pH 10). The bottom pH is the average of the top pH
and middle pH.
k The sum of calculated concentrations of all major Inorganic species exceeded the dissolved copper.
c The sum of calculated concentrations of all major inorganic species equaled the dissolved copper.
-------
Table 3-3. The average pH, 96 hr in t|rms of dissolved copper (moles/L), percentages of
dissolved copper made up by Cu , CuOH , Cu(OH) CuCO^', Cu(CO^)2°, respectively,
the sum of the inorganic percentages, and the theoretical copper organic percentages
a
pH
LC50
Cu Diss..
x 10 M
+ 2
% Cu
% CuOH+
% Cu(0H)2°
% CuC0^°
1 Cu(C03)2 2
X Inorganic
Cu-Org
Set 1 — 4/83 (Ca+,
Mg,, Na+
Effects)
Ambient Lake
8. 10
441
2.08
5.42
1 .05
40.8
0.662
49.9
50. 1
Superior Water
8.16
441
2.03
6.21
1 .38
46.7
0.871
57 .1
42.9
8.13
441
2.03
5. 78
1 .19
43.8
0.762
53.5
4b. 5
2 x 10~3 M CaCl
8.01
1720
4.96
9.01
1 .34
b2 .2
0.994
78.b
2 1 .<~
co Added
8.09
1720
4.96
10.8
1 .94
75.0
1.44
94. J
5. n
o
8.05
1720
4.96
9.77
1 .60
68.0
I .19
85.6
14.4
2 x 10~3 M MgCl^
8.01
630
4.10
7.63
1 . 14
53.2
0.833
66.9
33. 1
Added
8.10
630
4.10
9.38
1 .73
65.2
1 .2b
01.7
la. J
8.05
630
4. 10
8.41
1 .39
58.4
1 .01
73.3
2o. /
-3
b
2 x 10 M NaCl
8.10
1100
5.30
13.2
2.49
96.4
1.71
119
%
Added
8. 14
1100
5.30
14.5
3.00
105
2.05
131
Is
0h
8.12
1100
5.30
13.8
2.73
101
1.86
125
ub
Set 2 -- 6/83 (pH Effects at
Ambient Alkalinity)
C(>2 Bubbling
6.63
110
13.5
1 .22
0.0079
9.27
0.005
24.0
/ b. 0
6.53
no
13.5
0.964
0.0050
7.32
0.003
21.a
78.2
6.58
110
13.5
1 .07
0.0062
8.06
0.004
22.1
77.3
CO Bubbli ng
7 .30
299
10.0
4.21
0.129
33.8
0.093
48.3
51./
L
7.40
299
10.0
5.35
0.205
42.5
0.146
58.2
41 . H
7.35
299
10.0
4.72
0.160
37 .1
0.112
52.1
47.9
-------
»
Table 3-3 (cont inued)
pH
LC50
Cu Diss^.
x 10 M
X Cu
+ 2
% CuOH
Ambient pH
8,
.02
787
3.
.98
8
.81
8,
.10
787
3.
.98
10
.6
8,
.06
787
3.
.98
9
.61
NaOH Added
8,
.65
1120
0.
,753
7
.11
8.
.81
1120
0.
753
10.
. 3
8.
.72
1120
0.
753
8
.38
Set 3 — 7/83 (pH Effects at 3X Ambient Alkalinity)
u> KHC0- and
| NaHCO^ Added
CO2 Bubbling
KHCO and
NaHC03 Added
CO^ Bubbling
KHCO- and
NaHC03 Added
HCI Added to
Lower pH and
Alkalinity
Ambient Lake
Superior Water
7. 14
299
11.2
3.07
7.16
299
11.2
3.21
7.15
299
11 .2
3.13
7.90
1050
1 .94
3.06
7.99
1050
1 .94
3.77
7.94
1050
1.94
3.37
8.50
2140
0.583
3.65
8.64
2140
0.583
5.05
8.56
2140
0.583
4.23
7.16
331
20.4
6.28
7.17
331
20.4
6.44
7.164
331
20.4
6.34
> of Humics
and !
Suspended Clays)
7 .93
1370
3.74
6.79
8.15
1370
3. 74
11.2
8.03
1380
3.74
8.47
X Cu(0H)2° X CuCO^" X CutCOj)^ ^ X Inorganic X Cu-Org
1 .41
2.05
1.68
66. 7
80.2
72.0
0.912
1 .1
1 .07
81.9
98.2
88. A
18. 1
I .05
II .6
4.85
10. 1
6.74
57.7
81.7
67 .4
3.62
7.23
4.95
74.0
1 10
88. J
2b^0
b
0
11.7
0.063
0.069
0.066
81.3
84.9
82.9
0.571
0.629
0.599
96.2
100
98.0
3.79
o.ooc
2 .02
0.362
0.553
0.443
76.7
94.6
84.7
2.93
4.47
3.59
84.8
105
94.0
V
5.96
1 .73
3.29
2.32
90.7
124
105
13.7
25.6
18.2
110
150
130
0
0.139
0.146
0.142
29.7
30.5
30.1
0.034
0.036
0.035
56.5
57.5
57.0
<~3.5
42.5
43.0
9.891
2.43
1 .38
49.3
81 .0
61 .2
0.526
1.41
0.810
61.3
100
/ 5.9
30.7 t
o.oo1
24. I
-------
Table 3-3 (continued)
LC50
Cu Diss^.
pH3 x 10 M X Cu % CuOH+
Clay Added to 8.12 1040 2.01 5.59
70 NTU 8.36 1040 2.01 9.71
8.22 1040 2.01 7.10
Humics Added 7.91 3920 0.497 0.857
(5 mg/L TOC) 8.10 3920 0.497 1.33
8.00 3920 0.497 1.04
Clay and Humics 8.17 3860 0.469 1.46
Added (70 NTU, 8.28 3860 0.469 1.88
V (5 mg/L TOC) 8.22 3860 0.469 1.64
*—»
ro
Set 5 — 11/83 (Effects of Variable Humics Concentrations
Ambient Lake 7.94
Superior Water 8.12
8.02
Humics Added 7.91
( 1.25 mg/L 8.11
8.00
Humics Added 7.94
( 2.50 mg/L 8.06
8.00
Humics Added 7.95
( 5.0 mg/L 8.11
TOC) 8.02
1160 4.32 7.96
1160 4.32 12.1
1160 4.32 9.66
1610 2.19 3.76
1610 2.19 5.96
3120 1.13 2.37
3120 1.13 2.08
3120 1.13 2.74
3120 1.13 2.37
3860 0.466 0.878
3860 0.466 1.27
3860 0.466 1.04
% C^OHj^" % CuCO.^" % Cu(CO.^)2 ^ X Inorganic X Cu-Org
1.13
3.39
1 .81
45.7
78.5
58.0
0.850
2.52
1 .37
56. 1
96.2
70.3
43.9
3.6
29. 7
0.106
0.255
0.158
6.68
10.4
8.14
0.073
0.17b
0.109
7. 37
12.6
9.95
92. b
0 / .4
9U. 1
0. 329
0.54 7
0.417
12.5
lb.O
13.9
0.272
0.446
9. 342
15.0
19.3
16.8
65.0
60. /
6 J. 2
1 .06
2.43
1 .53
60.6
91 .4
73.0
0.690
1 .56
1 .00
74.7
112
69. 7
25.3
oa
11.3
0.468
1.17
0.359
28.2
44.3
18.3
0.29b
0.727
0.242
34.9
54.3
22.4
65. 1
45. 7
77 .b
0.276
0.481
0.359
16.1
21 .2
18.3
0.186
0.324
0.242
19.8
2 5.9
22.4
60.2
74.1
//. 6
0.120
0.250
0.168
6.66
0.56
7.85
0.077
0.159
0.108
8.21
11.7
9 .b4
91.6
66. 3
90.4
-------
Table 3-3 (continued)
LC50
PH3
Cu Diss,
x 10 M
% Cu+2
% CuOH+
% Cu(OH)^°
% CuCO^"
X Cu(C03)2 2
X Inorganic
X Cu-Org
Set 6 — 5/84 (Alkalinity Effects at Ambient
pH)
HCL Added to
7.84
724
11.5
16.9
1 .80
51.6
0. 182
«2.0
ltt.0
reduce alkalinity
7.84
724
11.5
16.9
I .80
51.6
0.182
82.0
18. U
to below ambient,
7.84
724
11.5
16.9
1 .80
51.6
0. 182
82.0
18.0
and CO^ removed
to maintain
ambient pH
Ambient Lake
7.87
1240
5.02
7.88
0.895
58.8
0.559
73. 1
2b.9
^ H-0
7.99
1240
5.02
10.4
1 .56
77.8
0.976
95.8
4.2
1 £
M
U>
7.93
1240
5.02
8.95
1 .15
66.7
0.719
82.5
17.5
(K*, Na+)HC03, C02
Added to 3x
7.92
8.04
614
614
1.71
1.71
2.79
3.68
0.350
0.606
76.2
101
3.09
5.43
tt4.4
113
15.b
a
0
Ambient
8.00
614
1.71
3.18
0.448
86.3
4.04
95.9
4.1
Alkalinity
(K+, Na+)HCO CO
Added to 6x
7.96
8.04
724
724
0.904
0.904
1.51
1.82
0.200
0.291
77.6
93.5
7.88
11 .4
B. 1
108
11.9
oa
Ambient
8.00
724
0.904
1 .64
0.238
84.1
9.27
96.2
3.a
Alkalinity
For each te9t water, the top pH listed is teh average pH value determined with a combination pH electrode that was
calibrated in one buffer at pH 7. The middle pH is the average pH value determined with a single.glass electrode
referenced against an Ag/AgCl double junction reference electrode and calibrated in two buffers (pH 1 and pH 10). The
bottom pH is the average of the top pH and middle pH value.
The sum of calculated major inorganic percentages exceeded 1002 of dissolved copper.
The sum of calculated major inorganic percentages was equal to 100X of dissolved copper.
-------
There will, in some cases, be given several different interpretations
of the data dependent upon whether or not a theoretical copper-organic frac-
tion is assuned to be present. The steeper than expected non-Nernstian
electrode response slopes, the associated large differences between dissolved
copper and the calculated sum of all known major inorganic complexes in some
test waters, and the relatively high stability constants for the formation of
many copper organic complexes has led to our speculation that there may be a
substantial copper-organic fraction and/or unknown copper-inorganic fraction
in test waters with ambient or lower lake water pH and/or alkalinity as will
be discussed in greater detail in Section 3.10. Despite substantial evidence
for the presence of a significant copper-organic fraction in some test waters,
the calculated concentration of the fraction and the proportion of dissolved
copper it represents may be subject to large error since the calculations
depend upon the difference between dissolved copper and the sum of the cal-
culated concentrations of all known major copper-inorganic complexes. The
reason is that each of the calculated concentrations of major inorganic copper
+ 2
species may be subject to substantial error due to errors in the Cu and pH
determinations and errors in the stability constants used in the calculations.
The possible large errors in the calculations of the concentrations and pro-
portions of theoretical copper-organic complexes along with the uncertainty in
their existence should be kept in mind in reading the discussion of the six
sets of copper toxicity tests given below.
3.1 SET 1 - SEPARATE EFFECTS OF MgCl , CaCl AND NaCl ADDITIONS ON COPPER
TOXICITY IN LAKE SUPERIOR WATER
The main objective of the first set of copper toxicity tests was to
determine the relative effects of the two principal components of hardness,
+ 2+2
Ca and Mg , on copper toxicity when added separately to Lake Superior
water. In addition, the effect of adding Na+ on copper toxicity in Lake
Superior water was also studied to determine if there might, be any toxicologi-
cal problems associated with using NaHCO^'or Na^CO^ to increase alkalinity
independently of hardness in later experiments.
The set of four tests consisted of 96 hour LC50 determinations and 7 day
growth studies on fathead minnow larvae in four types of test water: ambient
3-14
-------
Lake Superior water and separate 2 x 10 M concentrations of CaCl MgCl and
+ 2 +2
NaCl in Lake Superior water. The chloride salts of Ca and Mg were added
to the test waters instead of bicarbonate salts, so that increases in hardness
could be obtained without the associated increases in alkalinity normally
encountered in natural waters. Because the stability constants for copper
-3
chloride complexes are low, the addition of 4 x 10 M chloride should not
have any significant effect on copper speciation.
The LC50 values in terms of total, dissolved, and free (Cu+^) copper that
-3
were determined in lake waters with separate 2 x 10 M concentrations of
CaCl^, MgCl^ and NaCl were all greater than those determined in ambient lake
water, but the effects of CaC^ and of NaCl were substantially greater than
the effect of MgCl_ (Table 3-1). For example, the ratios of the dissolved
-3
copper LC50s in 2 x 10 M CaCl-, MgCl„ and NaCl to that in ambient lake water
+2
were, respectively, 3.9, 1.4, and 2.5 (Table 3-1). The ratios of Cu LC50s
-3
in 2 x 10 M CaC^, MgCl^, and NaCl to that in ambient lake water were,
respectively, 9.5, 2.9, and 6.5. Figure 3-1 contains four plots of the nega-
+ 2 . .
tive logarithm of the Cu activity versus the negative logarithm of the
dissolved copper, one for each of the four test waters. The arrows on the Y
and X axes show the relative positions of the negative logarithms of the LC50
values in terms of the Cu ^ activity and dissolved copper, respectively, for
the four test waters .
2 +2
The results of the Ca+ and Mg tests (in which hardness in both waters
was approximately 250 mg/L as CaCO which is approximately five times greater
+2
than in ambient lake water) suggest that Ca is much more responsible on an
equivalent basis for the correlations between observed reductions in copper
+ 2 +2
toxicity and hardness than Mg . Therefore, separate measurements of Ca and
+2
Mg may be of greater value in predicting copper toxicity than hardness
+2 +2
determinations, particularly if the Ca /Mg ratio changes substantially with
time in the same natural water or with different natural waters.
3-15
-------
7 -Log Cud 6
2 X 10~3 M Ca
2XIO"3m Naj
v2X IO"3M M g *2
o LW
0
Figure 3-1 The separate effects of Ca+2, Mg+2 and Na+ on the negative logarithms
Cu+2 and dissolved copper LC50s and in Lake Superior water
3-16
-------
The large increase in the LC50 values in lake water with the addition of
-3 +
2 x 10 M NaCl (which had no effect on hardness) indicated that Na could
-3
reduce copper toxicity on an equivalent basis (2 x 10 N NaCl compared to 4 x
-3 +2 +
10 N CaC^) to an even greater extent than Ca . Therefore, the Na test
indicated that neither NaHCO^ nor Na^CO^ could be used for increasing the
alkalinity in future tests, unless an additional chemical substituent which
could nullify the effect of Na+ on copper toxicity was added with the NaHCO
-3 +
or ^200^. The 2 x 10 M concentration of Na which was used xn the experi-
ment was much greater than would normally be found in any natural freshwater.
Therefore, the effect of more typical concentrations of Na on copper toxicity
+ 2 +2
would probably be small compared to Ca and Mg , even in relatively soft
waters .
-3
In going from the ambient lake water to the 2 x 10 M CaCl^, the calcu-
lated proportions (percentage/100) at the LC50 point of dissolved copper due
to Cu+^ and the other major inorganic copper species increase, however, the
calculated proportion of dissolved copper due to the theoretical copper-
organic fraction decrease (Table 3-3) possibly due to competition between the
+ 2 +2
Ca and Cu for organic ligands. Therefore, by referring to the first
grouped term of equation (1-22), it can be seen that it is possible that the
increase in the dissolved copper LC50 in going from ambient lake water to 2 x
-3 ...
10 M CaCl^ lake water is at least partially due to the positive contribu-
tions of the negative proportional changes of possibly toxic organic compounds
significantly offsetting the negative contributions of the positive propor-
tional changes of any toxic inorganic copper species.
The increase in the dissolved copper LC50 in going from ambient lake
-3
water to 2 x 10 M CaCl^ lake water appears to also be at least partially due
to decreases in the toxicity/unit concentration of one or more toxic copper
species as represented by the second grouped term of equation (1-22). The
postulate is supported by the chemical speciation calculations listed in Table
+ 2
3-2. The calculated concentrations at the LC50 point of Cu and all of the
other known major inorganic copper species are 5 to 10 times greater in the 2
-3
x 10 M CaCl^ lake water than in the ambient lake water. Also, the calcu-
lated concentration of the theoretical copper-organic fraction is greater in
3-17
-------
the 2 x 10 ^ M CaCl^ lake water than in the ambient lake water. Therefore, if
the calculations are even just qualitatively correct in showing that all of
the concentrations of major inorganic species and the concentration of the
copper-organic fraction increase, the toxicity/unit concentration (e.g., T\ in
equation 1-11) of one or more toxic copper species would have to decrease in
order for equation (1-14) to be fulfilled. The reason is that if all of the
toxicities/unit concentration remained constant in going from ambient lake
-3
water to the 2x10 M CaCl^ lake water, all of the overall fractional con-
tributions f^ of each species to toxicity would increase as can be seen from
equation (1-15) because the concentration of each of the major inorganic
species and the copper-organic fraction appear to increase. Therefore, if all
of the fractional contributions increased, all of the f^ would be positive
such that equation (1-14) could not be fulfilled.
Although the calculations of the concentrations of major inorganic copper
species may have some significant quantitative error, the calculated increases
(e.g., 5-10 times) are so large that it is unlikely that any of the concentra-
tions of the major inorganic species decreased. Of course, the calculated
increase in the concentration of the copper-organic fraction was not as large,
and is probably subject to much larger error. Therefore, it is possible that
-3
the concentration of the copper-organic fraction was lower in the 2 x 10 h
CaC^ lake water than in the ambient lake water, and that the copper-organic
fraction is toxic. However, even if that is the case, if it is assumed that
the toxicities/unit concentration remain constant, equation (1-17) would apply
so that the product of the toxicity/unit concentration times the decrease in
concentration (if any) of the copper-organic fraction would have to equal trie
sum of the product of the toxicity/unit concentration times the increase in
concentration of each major inorganic copper species. Although possible, it
is not probable. The reason is that the calculated increase in the total
concentration of the inorganic copper fraction in going from ambient lake
_3 —^
water to 2 x 10 M CaC^ is large ( + 1.24 x 10 M) compared to the calculated
total concentrations of the copper-organic fraction in either water (0.21 x
10 M in ambient, 0.25 x 10 in 2 x 10 M CaCl^). Therefore even if there
is a relatively large error in the calculated concentrations of the copper-
organic fraction and there is an actual decrease in the concentration of the
3-18
-------
copper-organic fraction in going from ambient lake water to the 2 x 10 M
CaCl2 instead of the calculated increase, it is not likely that the magnitude
of such a decrease would be comparable to the increase in the total concen-
tration of the inorganic copper fraction. Therefore, in order for equation
(1-17) to hold, which is based on an assumption that all of the toxicities/
unit concentration remain constant, the toxicity/unit concentration of the
copper-organic fraction vrould have to be much greater than the concentration
change weighted average of the toxicities/unit concentration of the inorganic
copper species. That is unlikely particularly since the concentration of
+2
Cu , which is postulated to be a very toxic species, increases substantially
and therefore potentially contributes significantly to the concentration
change weighted average of the toxicity/unit concentration of the inorganic
fraction.
The discussions above support the postulate that the increase in the
-3
dissolved copper LC50 in going from ambient lake water to 2 x 10 M CaCl^
lake water is due to a decrease in the proportion of a theoretical toxic
copper-organic fraction and/or to a decrease in the toxicity/unit concentra-
tion of one or more toxic copper species. Although it is impossible from the
present data set to determine the relative contributions of each possible
cause to the increase in the LC50, it is certain that the increase in the LC50
cannot be contributed to by proportional changes of the known major inorganic
species since they all increase. Furthermore, the contribution, if any, of
any proportional decreases of copper-organic complexes appears to be small.
Therefore, it appears probable that a decrease in the toxicity/unit concentra-
tion of one or more toxic copper species does occur which contributes signifi-
cantly to the increase in the LC50.
If Ca+^ does significantly decrease the toxicity/unit concentration of
one or more toxic copper species as was implied by the chemical speciation
calculations, it would be necessary in developing the chemical speciation
method of predicting LC50 values to determine several sets of toxicities/unit
concentration, one set for each of several values of hardness normally
encountered in natural waters. The advantage in that case of developing a
more empirical toxicities factors method is clear since any changes in the
3-19
-------
toxicity/unit concentration of toxic copper species would affect, and there-
fore be reflected in, the observed functional dependence of experimental LC50
values on hardness. However, if substantial amounts of toxic copper-organic
or other unknown toxic copper complexes are also present it would be extremely
difficult, and perhaps impossible, to develop either a chemical speciation
method or a more empirical toxicity factors method iAiich could be used to
accurately predict LC50 values in waters.
3.2 SET 2 - THE EFFECTS OF pH ON COPPER TOXICITY IN LAKE SUPERIOR WATER AT
AMBIENT ALKALINITY '
The objective of the second set of copper toxicity experiments was to
determine the effect of pH on copper toxicity in Lake Superior water at
ambient alkalinity and hardness. The four types of test waters used were:
ambient Lake Superior water with an average pH of 8.06 and Lake Superior
waters with the pH altered to an average pH of 6.58, 7.35, and 8.72, respec-
tively. The pH was decreased from the ambient pH without decreasing alka-
linity by bubbling CO^ into the test waters. The amount of NaOH required to
raise the pH to 8.72 from the ambient pH of 8.06 was not sufficient to cause
any significant increase in alkalinity over that of the ambient Lake Superior
water .
The LC50 values in terms of total and dissolved copper listed in Table
3-1 for set 2 test waters are similar and increase monotonically by a factor
of more than 10 with increasing pH over the entire range of pH tested from
+ 2
6.58 to 8.72. Jtowever, the LC50 values in terms of Cu activity and con-
centration increase by a factor of approximately 2 from pH.6.58 to 7.35, level
off between pH 7.35 and 8.06, and decrease by a factor of approximately 4 from
+2
pH 8.06 to 8.72. The somewhat unusual functional dependence of the Cu LC50s
on pH compared to the monotonic changes in the dissolved copper LC50 may be
reflective of a gradual reduction in the contribution of toxicity/unit con-
centration changes to the monotonic increase in the dissolved copper LC50s
compared to the contributions of proportional changes. Figure 3-2 contains
+2
four plots of the negative logarithm of the Cu activity versus the negative
logarithm of the dissolved copper, one for each of the four test waters in set
2. The arrows on the Y and X axes show the relative positions of the negative
3-20
-------
H=6.5 8
PH=7.3 5
7-
o- -
3-
CM
9-
7 -Log Cud 6 5
+2
Figure 3-2 The effect of pH on the Negative Logarithms of Cu and dissolved copper
LC50s in Lake Superior Water at 3X ambient alkalinity
3-21
-------
+ 2
logarithms of the LC50 values in terms of the Cu activity and dissolved
copper, respectively, for the four test waters.
As can be seen from the calculations for set 2 listed in Table 3-3, the
proportions at the LC50 point of dissolved copper due to copper hydroxy and
copper carbonate species increase with increasing pH over the range 6.58 to
+ 2
8.06. However, the calculated proportions of dissolved copper due to Cu
decreases and that due to the theoretical copper-inorganic fraction greatly
-2
decreases possibly due to the increased competition between CO. , OH , and
+ 2
organic ligands for Cu . The signs of the proportional changes between pH
8.06 and 8.72 are somevrtiat different in some cases, so they will be discussed
separately. The observed proportional changes between pH 6.58 and 6.06 along
with the first grouped term of equation 1-22 suggest that the increase in the
dissolved copper LC50s over that pH range is at least partially due to the
+ 2
positive contributions of the negative proportional changes of Cu . It may
also be due to possibly toxic copper-organic complexes significantly off-
setting the negative contributions of the positive proportional changes of
copper hydroxy and copper carbonate species.
It is also probable that reductions in the toxicity/unit concentration of
one or more toxic copper species, as represented by the second grouped term of
equation 1-22, do occur and contribute significantly to the increase in the
LC50 values. The reasoning is analogous to that for the set 1 experiments
although the support for the postulate is not quite as great as that for the
Ca ^ effect. The calculated concentrations listed in Table 3-2 for set 2
tests show the following. The calculated concentration at the LC50 point of
+ 2
Cu increases from pH 6.58 to 7.35 by a factor of over 3 and then remains
relatively constant from pH 7.35 to 8.06. The calculated concentration of the
theoretical copper-organic fraction increases by a factor of less than 2 from
pH 6.58 to 7.35, and then decreases by a factor of less than 2 from pH 7.35 to
8.06. The calculated concentrations at the LC50 point of all of the other
known major inorganic.copper species (e.g., copper hydroxy and copper carbon-
ate species) increase by a factor of well over 10 to well over 100 from pH
6.58 to 8.06. If the concentrations of all the major inorganic copper species
increase or remain constant and the concentration of the theoretical copper
3-22
-------
organic fraction increases or the copper-organic fraction is non-toxic from pH
6.58 to 8.06, there would have to be a decrease in the toxicity/unit con-
centration of one or more toxic species in order for equation (1-14) to be
fulfilled. Of course, the calculated concentration of the copper organic
fraction at the LC50 point decreases from pH 7.35 to 8.06, and the actual
concentration of the organic fraction may decrease from pH 6.58 to 8.06
despite the increase in the calculated concentration for pH 6.58 to 7.35.
That is because as discussed earlier, there may be large errors associated
with the calculation of the concentration of the copper-organic fraction.
Howver, even if that is the case, if it is assumed that the toxicities/unit
remain constant, equation (1—17) vrould apply so that the product of the
toxicity/unit concentration times the decrease in concentration (if any) of
the copper-organic fraction would have to equal the sun of products of the
toxicity/unit concentration times the increase in concentration for each major
inorganic copper species. Again, as was the case for Ca+^ addition, the
increase in the calculated total concentration of the inorganic copper frac-
tion from pH 6.58 to 8.06 (+6.71 * 10 ' M) is large compared to the calculated
concentrations of the theoretical copper-organic fraction at pHs 6.58, 7.35
and 8.06 (0.85 x 10 1.41 x 10 0.91 x 10 ' M, respectively). Therefore,
even if there is a relatively large error in the calculated concentrations of
the copper-organic fraction and the concentration of the copper-organic frac-
tion decreases with increasing pH from 6.58 to 8.06, it is unlikely that the
magnitude of any such decrease would be comparable to the increase in the
concentration of the total inorganic copper fraction. Therefore again, in
order for equation (1-17) to hold which is based on the assumption that all of
the toxicities/unit concentration remain constant, the toxicity/unit concen-
tration (or more accurately the concentration change weighted average
toxicity/unit concentration) of the copper-organic fraction would have to be
much greater than the concentration change weighted average toxicity/unit
concentration of the copper inorganic fraction. That is again, unlikely,
+2
particularly from pH 6.58 to 7.35 where the apparently toxic Cu concentra-
tion significantly increases and therefore potentially contributes signifi-
cantly to the concentration change weighted average toxicity/unit concentra-
tion of the inorganic copper fraction. The argument is not as strong for pH
+ 2
7.35 to 8.06 since the calculated concentration of Cu at the LC5G point
3-23
-------
remains almost constant and therefore
concentration change weighted average
inorganic copper fraction.
contributes little (if any)
toxicity/unit concentration
to
of
the
the
The discussions above concerning the pH range from 6.58 to 8.06 indicate
that increases in the dissolved copper LC50 over that range are due to nega-
+ 2
tive proportional changes of Cu and possibly toxic copper-organic complexes,
and/or to decreases in the toxicity/unit concentration of one or more toxic
copper species. Again, as with the Ca+^ effect, it is not possible from the
present set of data to determine the relative contributions of the two pos-
sible causes to the increase of dissolved copper LC50s with increasing pH.
+ 2 +2
Unlike for the Ca effect, the proportion of an inorganic species (Cu )
decreases with increasing pH and therefore could contribute to the increase
+2
in the dissolved copper LC50s. Furthermore, the species involved (Cu ) has
been postulated to be toxic and its proportion decreases substantially.
+ 2
Therefore, the proportional decrease in Cu from pH 6.58 to 8.06 could con-
tribute significantly to the increase in the dissolved copper LC50s . However,
the proportional decrease in Cu+^ cannot completely account for the observed
increase. The reason is that despite the negative proportional change of
+ 2 +2
Cu , the concentration of Cu increases from pH 6.58 to 7,35 and remains
almost constant from pH 7.35 to 8.06, whereas the concentrations of all of the
other known major inorganic complexes greatly increase. Therefore, it would
be impossible for equation (1-14) to be fulfilled, unless there was a signifi-
cant decrease in the concentration of toxic copper-organic complexes and/or
in the toxicities/unit concentration of one or more toxic copper species.
Furthermore, because any decrease in copper organic complex concentrations
appears to be small compared to the increases in the concentrations of the
inorganic complexes, it appears likely that decreases in the toxicity/unit
concentration of one or more toxic copper species do occur in going from pH
6.58 to 8.06, and do contribute significantly to the increase in the dissolved
copper LC50.
Although it was shown that it is not possible for the decrease in the
+ 2
proportion of Cu from pH 6.58 to 8.06 to completely account for the incre.
in dissolved copper LC50s, it is conceivable that the negative proportional
3-24
-------
decrease in Cu+^ could account for all, or at least most, of the observed
increase in the dissolved copper LC50s from pH 8.06 to 8.65. The reason is
*2
that in this case, the concentration of Cu does significantly decrease along
with the negative proportional change. Therefore, equation (1-14) could
+2
conceivably be fulfilled by the decrease in Cu concentration alone without
having to assume that the toxicities/unit concentration of one or more species
decreases, and/or that the concentrations of some toxic copper-organic
complexes (if any) decrease. If that was true, it can be shown from equation
(1-17) which would then apply to the pH range, that the toxicity/unit concen-
tration of Cu+^ would have to be >14 times the concentration change weighted
average toxicity/unit concentration of the remaining dissolved copper.
o o
Furthermore, since the increase in concentration of Cu(0H)„ , CuCO. and
+2
Cu(C0_) -2 are all much greater than the decrease in the Cu concentration,
+2
the toxicity/unit concentration of Cu would have to be much greater than for
any of those species.
If increasing pH does decrease the toxicity/unit concentration of one or
more toxic copper species, it would probably not be possible to develop a
chemical speciation method of estimating LC50s. The reason is that in order
to calculate the toxicities/unit concentration of copper hydroxy species, it
is necessary to vary the relative proportion of those species compared to
which can only be done, through varying the pH. However, if by varying the pH,
the toxicities/unit concentration also change, it would not be possible to
determine thera. Therefore, unless the toxicities/unit concentration of all of
the major copper hydroxy species are negligible, a chemical speciation method
could not be developed. However, any such effects of pH on the toxicities/
unit concentration should not interfere in the development of a toxicity
factors method, since they would affect and therefore be reflected by the
dependence of LC50 values on pH. If, however, there are significant amounts
of toxic copper-organic or other unknown toxic complexes present, it would be
extremely difficult to develop either a chemical speciation or an accurate
toxicity factors method.
3-2 5
-------
3.3 SET 3 - THE EFFECTS OF pH ON COPPER TOXICITY IN LAKE SUPERIOR WATER WITH
THE ALKALINITY INCREASED TO APPROXIMATELY THREE TIMES THE AMBIENT
ALKALINITY
There were three major objectives of the third set of copper toxicity
test9. The first objective was to determine the effect of pH on copper toxic-
ity in waters with an intermediate alkalinity. The second objective was to
determine the effect of alkalinity on copper toxicity at a constant, lower
than ambient lake water pH. The third objective was to determine the effect
of alkalinity on the pH dependence of copper toxicity by comparing the results
of the tests run in the third set at intermediate alkalinity to those of the
tests run in the second set at the lower ambient lake water alkalinity.
The first three test waters in the third set of copper toxicity tests
were all Lake Superior waters with alkalinities increased to approximately
_3
three times (e.g., 3 x 10 eg,/L) that of ambient lake water and pHs
adjusted to an average pH of 7.15, 7.94 and 8.56, respectively. The pH 7.94
is similar to that of ambient lake water. The alkalinities of the first three
test waters were increased without increasing the ambient hardness of lake
water by adding a ratio of NaHCO^ to KHCO^ which did not appear to have any
effect on copper toxicity based on preliminary static experiments with various
ratios of Na+/K+ added. The K+ apparently increases copper toxicity alone and
therefore can offset the decrease in copper toxicity by Na alone. The b.56
pH was the unaltered steady state pH of the water vAien the alkalinity was
increased to three times that of ambient lake water. The pHs of 7.94 and 7.15
were obtained without lowering the alkalinity by bubbling CO2 into the test
waters. In addition to the three test waters run in the third set at three
times the ambient alkalinity, a fourth test water was run at approximately
-3
one-half (e.g., -0.5 x 10 eq/L) the alkalinity of ambient lake water and at
a lower pH (7.15) than ambient lake water.
The effect of increasing pH on the LC50 values in the first three test
waters in set 3 which were at intermediate alkalinity can be seen from Table
3-1. The LC50 values in terms of total and dissolved copper were similar at
any given pH, and increased monotonically by a factor of over 7-fold with
increasing pH over the entire range tested from pH 7.15 to 8.56. However, the
3-26
-------
LC50 values in terms of the Cu activity and concentration decreased mono-
tonically by a factor of approximately 3-fold with increasing pH over the same
+ 2
range. Figure 3-3A contains four plots of the negative logarithm of the Cu
activity versus the negative logarithm of the dissolved copper, one for each
of the four test waters in set 3. The arrows on the Y and X axes show the
relative positions of the negative logarithms of the LC50 values in terms of
+ 2
the Cu activity and dissolved copper, respectively, for the four test
waters.
Plots of -log LC50DjSS Cu and -log LC50^Cu+2^ versus pH for both set 2
and set 3 test waters are presented in Figure 3-3B. The plots of -log
LC50^Tf,_ „ vs pH for sets 2 and 3 are qualitatively similar in that both show
DISS Cu
monotonic increases with increasing pH and a reduction in the slope at higher
pHs . The reduction in the slopes at higher pH may be due to a reduction in
factors other than inorganic proportional changes contributing to the increase
in the LC50s such as possibly a reduction in the contribution of toxicities/
unit concentration changes. It could also be due to increases in the propor-
-2 +
tion of CuCOH)^ and/or Cu(C0^)2 copper complexes as compared to CuOH
and/or CuCO^ complexes if either or both of the di complexes has a greater
toxicity/unit concentration than the corresponding mono complex.
The slope of the -log LC50DISg vs pH plot for set 3 waters run at
approximately 3 x 10 ^ eq/L alkalinity is steeper at all pHs tested than the
-3
slope of the plot for set 2 waters run at 1 x 10 eq/L alkalinity. There-
fore, the two curves diverge and there is a gradual increase in the ratio of
-3
the dissolved copper LC50 at 3 x 10 eq/L alkalinity to the dissolved copper
_3
LC50 at 1 x 10 eq/L alkalinity with increasing pH ranging from approximately
1.25 at pH 7.15 to approximately 2.0 at pH 8.56. That is not surprising since
much of the available literature suggests that the toxicity/unit concentration
+2 + ° -2
of Cu and of CuOH are far greater than that of CuCO^ or CuCCO^^ • How-
ever, it is somevAiat surprising that the apparent effect of alkalinity on the
pH dependence of dissolved copper LC50s is not even greater than was observed.
Furthermore, it appears that the observed effect may have been greater than
the actual effect of alkalinity due to a possible decrease in the sensitivity
of the organisms to copper in set 3 compared to set 2. Although a dissolved
3-2 7
-------
pH=7.1 6
A LK=26.4 MG/L
MG/L
=7.94
_A LK=I48 MG/L
p H—8.D 6
i\LK = l50 MG/L
7 -Log C-Un 6
Figure 3-3A The effect of pH on the negative Logarithms of Cu
LC509 in the Lake Superior water at 3X ambient alkalinity
3-28
+2
and dissoLved copper
-------
5
'
L 4. C W Ul p IL u 11 LUC 1VC
superior water at ambient and 3X ambient
3-29
-------
copper LC50 was not determined in ambient lake water in set 3, dissolved
copper LC50s in ambient lake water steadily increase from set 1 to set 2 to
set 4 which suggests that the organisms sensitivity to copper may have
decreased in going from set 2 to set 3. If so, part of the increase in dis-
-3
solved copper LC50s in going from set 2 at 1 x 10 eq/L alkalinity to set 3
-3
at 3x10 eq/L at a given pH may have been due to a decrease in sensitivity
causing a decrease in the toxicities/unit concentration of toxic species.
One of the reasons why the effect of alkalinity is not as great as
expected may be due to substantial organic complexation in the lower alkalin-
ity waters of set 2. For example, the dissolved copper LC50s of the second
test water of set 2 and the first test water of set 3 are the same. Although
0 -2
the concentrations of CuCO, and Cu(C0.) are much greater in the set 3
+ 2
test water than in the set 2 test water, the concentrations of Cu and of
CuOH are similar in the two waters. The reason appears to be that a somewhat
comparable amount of copper is bound to organic ligands in the lower alkalin-
* -2
ity water (Table 3-3) as was in the form of CuCO^ and CutCO^^ in the
higher alkalinity water. That, along with the identical dissolved copper
LC50s and the comparable Cu+^ and CuOH+ concentrations, suggest that the
toxicity/unit concentration of the copper-organic fraction may be somewhat
e
comparable to that of CuCO . Therefore, organic complexation in the lower
J e
alkalinity waters could have a similar effect on copper toxicity as CuCO^
formation in the higher alkalinity and therefore could mask to some extent the
effect of alkalinity.
+2
The monotonic decrease in the Cu LC50s with increasing pH as opposed to
+ 2
the non-monotonic dependence of Cu LC50s on pH in set 2 as shown in Figure
+2
3-3B, may be reflective of a greater contribution of decreases in the Cu
proportion to the increases in the dissolved copper LC50s in set 3 than in set
2. In set 2, other factors such as decreases in the proportion of toxic
copper-organic complexes and/or decreases in toxicities/unit concentration may
have contributed substantially to the increases in the dissolved copper LC50s.
In particular, the near Nernstian slopes of the electrode response and the
small or negative differences between the dissolved copper and the sum of
calculated concentrations of known inorganic copper species (both of which are
3-30
-------
shown in Table 3-3) indicate that any copper-organic fraction present is
probably very small. Therefore, it is unlikely that decreases in the propor-
tion of toxic copper-organic complexes (if present) would contribute signifi-
cantly to the observed increases in the dissolved copper LC50s. Of course,
there might be decreases in the toxicity/unit concentration of toxic copper
species which contribute significantly to the dissolved copper LC50 increase.
However, there is not as much evidence to support that as there was in set 2,
+ 2
because equation (1-14) can conceivably be fulfilled by decreases in the Cu
concentration alone over the entire pH range tested.
If the toxicities/unit concentration are assumed to remain constant with
increasing pH from pH 7.15 to 8.56, equation (1-17) will apply. If in addi-
tion the copper-organic fraction is assumed to be negligible, it can be shown
that in order for equation (1-17) to hold over the pH range of 7.15 to 8.56,
+2 +2
the product of the Cu toxicity/unit concentration times the decrease in Cu
concentration would have to equal the sura of the products of the toxicity/unit
concentration times the increase in concentration for each of the other inor-
ganic copper species. By substituting values from Table 3-2 into equation
+ 2
(1-17), it can be shown that the toxicity/unit concentration of the Cu would
have to be greater than 88 times the concentration change weighted average
toxicity/unit concentration of the rest of the dissolved copper if the toxici-
ties/unit concentration remained constant. Furthermore, because the increases
+ -2
in the concentration of CuOH , CutOH^* and Cu^O^^ are larger and the
increase in the concentration of CuCO-0 is much larger than the decrease in
+2 . +2
the Cu concentration, the toxicity/unit concentration of the Cu would have
+ -2
to be greater than those for CuOH , CutOH)^ and CutCO^^ ar|d much greater
than for CuCO^".
Since we want to determine the effects of alkalinity on copper toxicity,
it is interesting to note that even in the minimum case with the toxicities/
unit concentration of the copper hydroxy species assumed to be zero, the
+ 2
toxicity/unit concentration of Cu is still calculated to be at least 82
times greater than the concentration change average weighted toxicity/unit
-2
concentration of CuCO^0 and CutCO^0^ • That does not mean of course that
3-31
-------
-2
the fractional contribution of CuCO^° and/or Cu(CO^) to toxicity are neces-
sarily negligible since the fractional contribution is equivalent to the
toxicity/unit concentration times the concentration. Therefore if the ratio
-2 +2
of the sum of CuCO^° and CutCO^^ concentrations to the Cu concentration
is large such as in alkaline waters, they may still contribute significantly
to toxicity.
The results from the fourth test water of set 3 run at one-half ambient
alkalinity and a pH of 7.16 compared to the results from the first test water
of set 3 run at three times ambient alkalinity and an almost identical pH of
7.15, indicate the following. In going from test water 1 to test water 4, the
Cu+^ LC50 approximately doubles, the concentrations of CuOH+, Cu(OH) and
_2
Cu-Org increase and the concentrations of CuCO^0 and CuCCOj^ decrease. If
the concentration changes (Table 3-3) in going from test water 1 to test water
4 are substituted into equation (1-17) assuming that the toxic ities/unit con-
centration remain constant, the following equation results
<3'4 * 10"8' Tcu'2 * (1'2 * Itf8> TCuOH* ' <2'7 * 10"1
-------
3.4 SET 4 - THE SEPARATE AND JOINT EFFECTS OF HUMICS AND SUSPENDED CLAY ON
COPPER TOXICITY IN LAKE SUPERIOR WATER
The objective of the fourth set of copper toxicity tests was to determine
the separate and joint effects of humics and suspended clays on copper toxic-
ity in Lake Superior water. The four test waters used in set 4 were: ambient
Lake Superior water with negligible amounts of humics and suspended clay,
ambient lake water with Lake Superior shore red clays added in suspension
until a turbidity of 70 NTU was obtained, ambient lake water with humics pur-
chased from Aldrich added to a nominal TOC of 5 mg/L, and ambient lake water
with both the clay and humics added at the previous separate levels. The
addition of the clay suspension increased the average pH of the ambient lake
water between 0.15 and 0.20 pH units possibly due to the exchange of H+ with
cations from the clay, but did not have any significant effect on any other
general water quality parameters. The addition of the humics did not have any
significant effect on the pH or any other general water quality parameter.
Table (3-1) shows that in going from ambient lake water to the suspended
clay water, the total copper LC50 increases approximately 50%, whereas the
+ 2
dissolved copper LC50 decreases approximately 25% and the Cu LC50 decreases
by over 50%. Figure 3-4 contains four plots of the negative logarithm of the
+ 2
Cu activity versus the negative logarithm of the dissolved copper, one for
each of the four test waters in set 4. The arrows on the y and x axes show
the relative positions of the negative logarithms of the LC50 values in terms
of the Cu ^ activity and dissolved copper, respectively, for the four test
wat ers.
Since the suspended clay can adsorb copper to some extent, the increase
in the total copper LC50 is not surprising. However, since the clay suspen-
sion increased the pH of the ambient lake water, the decrease in the dissolved
copper LC50 was somewhat unexpected because the dissolved copper LC50s
increased with increasing pH in set 2 and set 3 tests. Furthermore, in look-
ing at the chemical speciation calculations in Table (3-2), it can be seen
+2
that the concentrations of Cu at the LC50 point and all of the other known
-2
major inorganic species except CuCCO^j ar>d CuCOH^* decrease substantially
in going from the ambient lake water to the suspended clay lake water, and
3-33
-------
HA -0.0 MG/L
7--
LW
LAY
8-
O
Figure 3-4 The separate and joinc effects of humics and suspended clay on the
negative togarithn of Cu+^ and dissolved copper LC50s
3-34
-------
that the calculated concentration of the theoretical copper-organic fraction
also decreases slightly. The calculated concentrations of Cu(OH) stays
_2
approximately the same and the calculated concentration of CuCCO^^
increases, but to a small extent compared to the decreases in concentrations
of Cu+^, CuOH+ and CuCO^'. Therefore, if the toxicities/unit concentration
are assumed to remain constant and if any copper-clay complexes formed are
-2
assumed to be non-toxic, the toxicity/unit concentration of CuCCO^^ would
have to be many times greater than the concentration change weighted average
toxicity/unit concentration of the rest of the dissolved copper for equations
+ 2
(1-14) and (1-17) to hold. Furthermore, since the decrease in Cu concentra-
_2
tion is approximately 10 times the increase in Cu(C0,)_ , the toxicity/unit
-2
concentration of Cu(C0.) would have to be over 10 times greater than for
+2
Cu for equation (1-17) to hold. That is unlikely considering that the
results of previous tests have suggested that the toxicity/unit concentration
-2 +2
of CutCO^^' i-s lower than for Cu
The above discussion suggests that in order for equation (1-14) to hold
in going from ambient lake water to suspended clay lake water, either the
toxicity/unit concentration of one or more toxic copper species must increase,
or some of the suspended copper-clay complexes must be toxic, or both.
Table (3-1) shows that in going from the ambient lake water to the humics
lake water, both the total and dissolved copper LC50s increase by a factor of
+2
approximately three fold, whereas the Cu LC50 decreases by over 60%. Humics
have a relatively large binding affinity for copper and do not appear to
affect general water quality parameters. Therefore, the rather large
increases in the total and dissolved copper LC50s in going from ambient lake
water to the humics lake water is not surprising. However, again as with the
+2 + 0
clay, the concentrations of Cu , CuOH and CuCO^' decrease substantially and
in this case, the concentrations of CutOH)^0 and Cu(,C0^)^' decrease as well.
Therefore, in order for equation (1-14) to be fulfilled in going from ambient
lake water to humic lake water, either the toxicity/unit concentration of one
or more toxic copper species must increase or at least some of the copper-
humic complexes must significantly contribute to toxicity, or both.
3-35
-------
Although Che hunics addition does not significantly effect the hardness
determination possibly because of the EDTA titration stripping any humic bound
Ca+^ or Mg+^, the humics may substantially reduce the concentration of Ca+^
and/or Mg+^ which could increase the toxicities/unit concentrations of the
toxic copper species (83). Also copper-hunic complexes could possibly contri-
bute substantially to toxicity but again through, some external mechanism
since dialysis experiments (82) have shown that virtually all of the cadmium-
humic complexes have molecular weights exceeding 1000 and are, therefore,
unlikely to be transported across gill membranes.
Table (3-1) shows that in going from ambient lake water to the humics
plus suspended clay water, the increase in the total copper LC50 is almost
equal to the sum of the separate increases in going to the suspended clay
water and in going to the hunics water. The dissolved copper LC50 increases
by a factor of approximately three , whereas the Cu+ LC50 decreases by
approximately 60%. The concentrations of all of the major inorganic forms of
-2
copper decrease except for CutCO^^ which increases slightly. Therefore, by
an argument analogous to that given above, it is unlikely that equation (1-14)
could be fulfilled without the toxicity/ unit concentration of one or more
toxic copper species increasing and/or copper-hianic complexes contributing
significantly to toxicity.
In going from the humics water to the humics plus suspended clay water,
+ 2
the dissolved copper LC50 and the Cu LC50 remained almost constant despite
the increases in the concentrations of all of the other inorganic copper
species. This suggests that while the combined effect of humics and suspended
clays on copper toxicity in terms of total copper is almost additive, the
+ 2
combined effects in terms of dissolved copper and Cu are at best, no more
than the effects of humics alone. That is sometftat surprising since it
appears that the sole addition of suspended clay to lake water does lead to an
increase in the toxicity/unit concentration of toxic copper species and/or to
the formation of toxic copper-clay complexes as discussed earlier. However,
the presence of substantial amounts of humics appears to almost completely
nullify such suspended clay effects and reduce the effects to that of
adsorption alone.
3-36
-------
3.5 SET 5 - THE EFFECTS OF VARIABLE HUMIC CONCENTRATIONS ON COPPER TOXICITY
IN LAKE SUPERIOR WATER
The objective of set 5 was to determine the effect on copper toxicity of
humic concentrations that were lower than the 5.0 mg/L used in set 4, because
+2
the reduction in the Cu LC50 observed in set 4 for that humic concentration
may have possibly been due to the himics stresssing the organism. The four
test waters of set 5 consisted of ambient lake water and lake water to which
hunics were added to a nominal TOC of 1.25 mg/L, 2.50 mg/L and 5.0 mg/L.
Table (3-1) shows that both total and dissolved copper LC50s increase
mono tonic ally by over a factor of 3 with increasing hunic concentration to
5 mg/L nominal TOC with the dissolved copper LC50 averaging approximately o0%
+ 2
of the total copper LC50. The Cu LC50 decreases approximately 33% in going
from ambient lake water to the 1.25 mg/L nominal TOC water, remains constant
in going from the 1.25 mg/L nominal TOC water to the 2.50 mg/L nominal TOC
water, and then decreases by approximately 50% in going from the 2.50 mg/L
nominal TOC water to the 5.0 mg/L nominal TOC water. In going from the ambi-
ent lake water to any of the humic waters, the concentration at the LC50 point
+ 2
of Cu , and all of the other major inorganic species decrease. Therefore, in
order for equation (1-14) to be fulfilled, either the toxicities/unit concen-
tration of one or more toxic copper species must increase, or at least some of
the copper-humic complexes must contribute significantly to toxicity, or both.
+ 2
Figure 3-5A contains four plots of the negative logarithm of the Cu activity
versus the negative logarithm of the dissolved copper for the four test waters
of set 5. The arrows of the Y and X axes show the relative positions of the
+ 2 . .
negative logarithms of the LC50 values in terms of the Cu activity and dis-
solved copper, respectively.
There is some evidence to support che postulate that the effects of the
lower himic concentrations on copper toxicity may also be primarily through
increases in toxicities/unit concentration rather than through contributions
of copper-hunic complexes to toxicity. The reason is that despite an over
100% increase in the large copper-organic fraction in going from the 1.lb mg/L
nominal TOC water to the 2.50 mg/L TOC water, the concentrations at the LC50
+2
point of Cu and all of the other known major inorganic copper species remain
3-3 7
-------
HA =1.25 MG/L
' sf HA=2.5 MG/L
LW
HA =5.0 MG/L
CM
o
Figure 3-5A. The,effects of variable humic concentrations on the negative
logarithms of Cu and dissolved copper LC50s in Lake Superior water
3-38
-------
relatively constant, as can be seen in Table 3-2. That suggests that the
copper-organic fraction does not contribute significantly•to toxicity. It
also suggests that the toxicities/unit concentration remain relatively con-
stant in going from the 1.25 mg/L nominal TOC water to the 2.50 mg/L nominal
TOC water even though they appear to possibly increase in going from the
ambient lake water to either the 1.25 mg/L nominal TOC water or the 2.50 mg/L
nominal TOC water.
The results of the set 5 copper toxicity tests suggest that humics may
increase the toxicity/unit concentration of one or more toxic copper species.
As mentioned previously, humics could possibly increase toxicities/unit con-
+ 2 +2
centration through the binding of Ca and/or Mg which appear to possibly
decrease toxicities/unit concentration. However, if that is the case, the
results of set 5 suggest that the 1.25 mg/L nominal TOC is capable of binding
+2 +2
most of the Ca and Mg present because there does not appear to be any
increase in the toxicities/unit concentration in going from the 1.25 mg/L
nominal TOC water to the 2.50 mg/L nominal TOC water.
The effects of humics added to a nominal TOC of 2.5 mg/L, 10 mg/L, and 'lb
mg/L on cadmium toxicity in Lake Superior water are depicted in figure 3-5B.
+ 2
The figure contains four plots of the negative logarithm of the Cd activity
versus the negative logarithm of dissolved cadmium, one for Lake Superior
water and three for the three waters with added hunics. The arrows on the Y
+ 2
and X axes show the relative positions of LC50 values in terms of the Cd
activity and dissolved cadmiian, respectively, for the four test waters. The
figure shows that whereas increased in humic concentration increase dissolved
+ 2
cadmiun LC50s, they decrease Cd LC50s. Therefore, the effects of increasing
humic concentration on cadmium toxicity are qualitatively identical to those
on copper toxicity which are depicted in figure 3-5a.
By analogy to the earlier discussion on the effects of humics on copper
toxicity, the results of the cadmium toxicity tests suggest that hunics
increase the toxicity/unit concentration of toxic cadmium species and/or that
at least some cadmium-humic complexes are toxic. There were no apparent signs
of stress on the organisms except in the 20 mg/L humics water where a statis-
3-39
-------
HA=2.5 MG/L
6-
LW
7.3---
3 :
O
8-
Figure 3-5B The effect of variable humic concentrations on the negative logarithms
of Cd+2 and dissolved cadmium LC50s in Lake Superior water
3-40
-------
tically significant decrease in the average 7 day growth weight in control
water was observed. However, the organisms could still possibly be stressed
at the lower hvanic levels which could increase the tox ic i t ie s/uni t concentra-
tion of toxic cadmium species. Although it is possible that some of the
cadmium-hunic complexes could exert toxicity directly, the mode of such action
would have to be external because almost all cadmium-humic complexes appear to
have molecular weights well above those which could be transported across
membranes into the organism (82).
3.6. THE EFFECTS OF ALKALINITY ON COPPER TOXICITY IN LAKER SUPERIOR WATER AT
AMBIENT pH
The objective of set 6 of the copper toxicity tests was to determine the
effect of various levels of alkalinity on the toxicity of copper to fathead
minnow larvae in Lake Superior water at ambient pH. The four test waters of
set 6 consisted of Lake Superior water with the alkalinity reduced to approxi-
mately 38% of the ambient alkalinity, Lake Superior water with ambient alka-
linity, and Lake Superior water with the alkalinity increased to approximately
3.5 times and to approximately 7 times the ambient alkalinity, respectively.
The pH of all four test waters were maintained as close to that of the ambient
pH of Lake Superior water as possible. The alkalinity of the first test water
was reduced to below the ambient alkalinity by the addition of hCL. The ph of
the low alkalinity water was maintained close to ambient pH by aerating the
headbox with air from which a substantial portion of the CO^ had been removed
by a commercial CO^ scrubber. The alkalinity of the third and fourth test
waters was increased over that of ambient lake water by adding a ratio of
NaHCO^ to KHCO^ which did not appear to have any effect on copper toxicity
based on preliminary static toxicity tests conducted on fathead minnow larvae
+ +
in waters containing various ratios of Na and K chlorides. The pHs of the
higher alkalinity systems were maintained close to the ambient lake water ph
by bubbling the headbox with CO^.
It can be seen from Table 3-1 that large changes in alkalinity have an
unexpectedly small effect on copper toxicity in terms of total or dissolved
copper. Although the LC50 values in terms of total and dissolved copper for
ambient Lake Superior water are almost twice that for the other three test
3-41
-------
waters, the LC50 values for the other test waters are almost identical despite
large differences in the alkalinity. In contrast, the LC50 values in terms of
+ 2
the Cu activity and concentration decrease monotonically with increasing
alkalinity. Figure 3-6 contains four plots of the negative logarithm of the
+ 2
Cu activity versus the negative logarithm of dissolved copper, one for each
of the four test waters of set 6. The arrows on the Y and X axes show the
relative positions of the negative logarithms of the LC50 values in terms of
+2
the Cu activity and dissolved copper, respectively.
The results of set 6 are consistent with the results of test waters 1 and
4 of set 3 because the LC50 values in terms of total and dissolved copper were
similar for those test waters even though the alkalinity of the two test
waters were much different. Furthermore, just as with set 6, the LC50 values
+2
in terms of the Cu activity and concentration decreased substantially with
increasing alkalinity.
The concentrations of Cu+^, CuOH+, and CuCOH)^ decrease with increasing
alkalinity of the test waters in set 6 whereas the concentrations of CuCO °
_2
and CuCCO^^ increase. Therefore, equation (1-14) can be fulfilled without
changes in the toxicity/unit concentration of toxic copper species occuring.
However, the results for set 6 and for test waters 1 and 4 of set 3 are com-
pletely inconsistent with a model that assunes both constant toxicities/unit
concentration and a much higher concentration change weighted average
toxicity/unit concentration for Cu ^ and CuOH+ than for CuCO^0 and Cu(CO^)^
as will be discussed below.
It was previously shown in Section 3.3 that the toxicity/unit concen-
+2
tration of Cu could be no greater than 3.5 times that of the concentration
• -2
change weighted average toxicity/unit concentration of CuCO^" and CuCCO^)^
if the toxicities/unit concentration remain constant in going from test water
1 of set 3 to test water 4 of set 3. Likewise, if the test waters of set 6
are compared assuming constant toxicities/unit concentration and using equa-
tion (1-17), it can be shown that the concentration change weighted average
+ 2 +
toxicit y/umt concentration of Cu and CuOH is not substantially greater
than that for CuCO^° and CuCCO^)^ • For example, if the toxicities/unit
3^42
-------
ALK = I' 7 MG/L
ALK= 42 MG/L
-------
concentration are assisted to remain constant in going from test water 1 of
set 6 to test water 4 of set 6, it can be shown from equation (1-17) that the
+ 2
concentration change weighted average toxicity/unit concentration of Cu and
CuOH+ is, at most, less than 2 times greater than that for CuCO^" and
Cu(C03)2~2.
In summary, it is possible that changes in alkalinity at constant pH do
not affect the toxicities/unit concentration of toxic copper species. how-
ever, if the toxicities/unit concentration of toxic copper species do remain
constant for waters of different alkalinity but the same pH, it does not
appear that there are large differences between the calculated concentration
change weighted average toxicities/unit concentration of Cu+^ and CuOH+ and
_2
that of CuCO^" and CuCCO^^ • That means that the carbonate complexes could
possibly contribute significantly to copper toxicity in waters with substan-
tial alkalinity. However, recently completed static studies by Benoit and
Mattson (83) have indicated that the anion of Na+ and K+ salts may have a
substantial impact on the toxicity of the salt. Therefore, the ratio of
NaHCO^ to KHCO^ that we used to raise alkalinity in these waters which was
based on a no effect ratio of Na+ and K+ chloride salts, may not have been the
proper ratio to use to completely nullify the opposite effects of Na* and K+
on toxicity vtaen added as bicarbonate salts.
3.7 FIELD STUDY ON THE NAUGATUCK RIVER, CONNECTICUT
Static 96 hour LC50s and EC50s in terms of total, dissolved and free
*2
copper (Cu ) were determined for one day old fathead minnow larvae in water
samples taken from several different sites along the Naugatuck River. Static
tests were run in Lake Superior water for comparative purposes, water taken
from a relatively unpolluted upstream site (Nl), and water taken from several
downstream sites containing increasing quantities of industrial and municipal
sewage effluents (N4-A, N5, N6, N7). Figure 2-2 is a map showing the location
of the various sites along the Naugatuck River.
The LC50s and EC50s in terms of total, dissolved and free copper (Cu**
activity) for the different site waters are listed in Table 3-4 based on meas-
3-44
-------
uremenCs performed at Che beginning of each test. In going from Che Lake
Superior water Co che relacively unpolluted NaugaCuck water (Nl) to the
increasingly polluted NaugaCuck waters (N4A, N5, N6, N7), the LC50 values in
terms of noC only tocal and dissolved copper, buC also Che Cu+^ acCivicy, gen-
erally increase. The increases in Che cocal and dissolved copper LC50s are
not surprising because the binding capacity for copper of waters containing
municipal sewage effluents are generally much greater than that of similar
waCers conCaining liccle or no municipal sewage effluent. However, Che pro-
+ 2
porCional increases in Che Cu LC50 values are ofcen also substantial. Thac
indicates that the decrease in copper toxicity in going downstream is not just
due to any increase in the binding capacity of the water because if that was
~ 2 / • x
the case, the increases in the Cu LC50 values (if any) would be much
smaller. Therefore, there appear to be factors introduced downstream which
can not only bind copper, but reduce the toxicity of the non-bound bioavail-
able fraction. One of those factors may be hardness because hardness values
increase monotonical1y in going downscream from N-l (Hardness = 36 mg/L) Co
N-7 (Hardness = 90 mg/L), and because hardness has been previously shown Co
reduce copper coxiciCy.
Alchough che LC50 values in cerms of Che CoCal, dissolved and free copper
follow general Crends , Chere are several anomalous values. In parCicular, Che
+2
Cu LC50 value for che N-4A wacer is lower Chan in Che N-l waCer even chough
ic is much higher Chan for Che N-l waCer in all of Che ocher downscream waCer
samples. That anomality may have been at least partially due to possible
+ 2
errors in the initial determinations of the Cu activities in (N-4A) test
+ 2 +2
waters. If the Cu LC50 value for the N-4A test waCer is based on Cu
deCerminaCions ac Che end of Che 96 hour determination, it is at least com-
+2
parable to the Cu LC50 in N-l water. The only other anomalies concern the
N-7 waCer for which che CoCal, dissolved and free (Cu+^) copper LC50s are
lower Chan for Che N-6 waCer upstream. Thac may have been due to diluCion
from Great Brooke which intersects the Naugautuck River jusC upsCream from N-7
and v*iich may have a lower binding capaciCy Chan waCer from Sceele Brook which
inCersecCs Che NaugauCuck River jusC upstream from N-6.
3-45
-------
TABLE 3-4
Static 96 hour acute values for I day old fatheaad minnow larvae in water taken
from various sites along the Naugatuck R. in Connecticut
(the parentheses contain 95% confidence intervals)
Hard and A1 k in ^
Li
LC50s in ug/L
EC50s in ug/L
+ 2X +2
Water pH Hard Alk Total Dissolved (Cu ) Total Dissolved (Cu )
Lake
Super ior
7 . 74
52
55
55
(38-81)
52
1 .82
( 1 .02-3.23)
55
(38-81)
52
(0
1 .54
.92-2.56)
N-l
7.44
36
38
180
(166-195)
171
3.66
(3.14-4.26)
173
(153-196)
164
(2,
3.40
.61-4.41)
N-4A
7.50
55
42
322
(249-415)
232
1.96 J
3.00
201
167
0.60
N-5
7 .48
68
40
511
(439-594)
363
16. 1
(12-21.7)
229
153
3.30
N-6
7.33
82
40
>998
>449
>20. 1
265
133
3.46
N-7
7 .28
90
43
689
(555-854)
42 7
14.8
(11.5-18.9)
282
(257-310)
1/5
3.02
-------
The EC50s in terns of total, dissolved and free copper are also listed
on Table 3-4. The EC50s are the chemical concentrations required to adversely
effect the mobility of 50% of the organisms. In contrast to the LC50 values,
+2
the EC50 values in terms of total, dissolved, and free (Cu ) copper remain
relatively constant in going from the N-l water upstream to the various waters
downstream (N4-A, N5, N6, N7). Again, the only major anomaly is the Cu ^ EC50
+ 2
for the N4-A water which, as previously discussed for the Cu LC50, may be
based on erroneous activity measurements. The relative constancy of EC50
values compared to the increasing LC50 values going downstream may possibly be
due to some kinetics effect that decreases the bioavailable fraction through-
out the 96 hour tests (90).
The above postulate is based on the following reasoning. During a 96
hour test, the exposure of an organism to a given chemical at the gill mem-
brane is given by
wh e r e
Q(t) = flow of water across the gills as a function of time
^BAF^O = concentration of the bioavailable fraction of the chemical
It should require less total exposure for an organism to develop and
continue to exhibit mobility impairment than for it to be killed. If the
kinetics of binding are slow enough such that the exposure of the organism is
sufficient to cause mobility impairment before a substantial reduction in the
bioavailable fraction occurs, but fast enough to lower the bioavailable frac-
tion of the chemical substantially before lethal amounts of exposure occur,
the EC50s should be less dependent upon the binding capacity of the water than
the LC 50s.
+2
The postulate is partially but not completely supported by Cu deter-
minations in waters bracketing the LC50 point after the termination of the
96 hr .
Total Exposure =
3-47
-------
+ 2
tests. The Cu activity should be at least somewhat proportional to the
+ 2
bioavailable fraction. The Cu activities in N-5 and N-6 test waters which
bracket the LC50 point decreased by close to 50% and 252 during the duration
of the test. Those are the two waters which show the largest increase in LC50
+ 2
values from those upstream. However, the Cu activity in one of the N-7 test
waters which bracketed the LC50 increased during the test even though the LC50
for the N-7 test waters was also substantially higher than for the upstream
waters N-l and N4A.
The results of these tests, particularly for the N-4A, N-5, N-6 and N-7
test waters, should be used with caution since there was substantial algae
growth in all of those test chambers at the end of the tests. Such algae
growth can lead to substantial pH and D.O. fluctuations during the tests which
can obviously greatly effect the bioavailable fraction and total exposure of
the organisms. Such factors can effect not only acute values in terras of the
free copper, but also acute values in terms of the total and dissolved copper
as well. Therefore, such tests should be conducted in the future in such a
way as to minimize algae growth. Possible solutions include the use of
renewal or flowthrough tests instead of static tests or the use of some alga-
cide which is non-toxic to the test organisms. The use of renewal or flow-
through tests could also have the advantage of minimizing any kinetic effects
+ 2
(as previously discussed) which may occur. Future Cu activity determina-
tions should also take into account possible absorptions of copper onto the
walls of the elctrode chamber.
3.8 POSSIBLE VARIATIONS IN LARVAE SENSITIVITY TO COPPER TOXICITY OVER TIME
In developing either a chemical speciation method or a toxicity factors
method for estimating LC50s, it is necessary to use data from multiple sets of
toxicity tests. Therefore, it is necessary to determine if the sensitivity of
larvae to copper toxicity changes over time. If the sensitivity of the larvae
does change, it is then necessary to normalize all data from each set to some
chosen standard test water before the data from different sets can be analyzed
together to develop either a chemical speciation or toxicity factors method.
3-48
-------
Of the six sets of copper toxicity tests run to date, five included as
one test water, ambient Lake Superior water, to determine if any changes in
larvae sensitivity and/or the basic diluent water occurred over time. Set 3
was the only one which did not include as a test water, ambient Lake Superior
water. Figure 3-7 depicts the change in the negative logarithm of the dis-
+ 2
solved copper and Cu LC50s in ambient Lake Superior water between different
sets as a function of the date on which each set was started. Table 3-1 and
Figure 3-7 show that in going from the ambient lake water of set 1 to the
ambient lake water of set 4, the dissolved copper LC50 increases monotonically
+ 2
by over a factor of 3 and that the Cu LC50 increases monotonically by over
a factor of 6. The chemical speciation calculations in Table (3-2) show that
the calculated concentrations at the LC50 point of all of the other major
inorganic copper species greatly increase as well and that the calculated
concentration of the theoretical copper-organic fraction increases despite a
proportional decrease. Therefore, in order for equation (1-14) to hold in
going from the ambient lake water in set 1 to the ambient lake water in set 4,
it would be necessary for the' toxicity/unit concentration of one or more toxic
copper species to decrease, assuming that the calculation showing the concen-
tration of the copper-organic fraction increasing is at least qualitatively
correct. Of. course, as always, there is a chance that the concentration of
the copper-organic fraction actually decreases in going from the set 1 to the
set 4 ambient lake water because the calculation of the copper-organic
fraction is subject to large error. If that was the case, it would be
theoretically possible for equation (1-14) to be fulfilled without any change
in the toxic ities/unit concentration. However, it is unlikely that any
decrease in the concentration of the copper-organic fraction would be large
enough to offset the relatively large increases in the concentrations of all
of the major inorganic copper species.
The above discussion indicates that the sensitivity of fathead minnow
larvae to copper toxicity changes over time, and that such changes may be
reflected in apparent changes in the toxicity/unit concentration of one or
more toxic copper species. If this is the case, it will probably be necessary
in the future to run one ambient lake test water per each set of tests to use
as a reference water, to designate one such ambient lake test water as the
3-49
-------
Hog u*>(cudl#B
3
o
3
u
o
m
o
-j
o»
o
J
Months
Figure 3-7. The variation in the negative logarithms of Cu and dissolved
copper LC50s in ambient Lake Superior water between different sets over time
3-50
-------
standard test water, and to normalize the data from each set to that of the
+2
standard water by multiplying all Cu concentration data in the set by the
+ 2 +2
ratio of the Cu LC50 in the reference water for the set to the Cu LC50 in
the standard water.
3.9 DIFFERENCES BETWEEN STEADY STATE AND EQUILIBRIUM CONCENTRATIONS
The accuracy of equations (1-1) to (1-7) for calculating the steady state
concentrations of copper hydroxy and copper carbonate species from the steady
• • +2
state activity of Cu will depend to a large extent on how close those
... +2
chemical species are to being in equilibrium with the Cu at steady state,
+ 2 +2
regardless of how far the Cu steady state activity may be from the Cu
equilibrium activity. The reason is that the equations are based on an
+ +2
assumption of equilibrium between the given species (e.g., CuOH ) and Cu ,
+2 . ...
not upon the assumption that the Cu is in equilibrium with the entire system
or that the entire system is at equilibrium. Therefore, any difference
+2
between the steady state and equilibrium Cu activity should represent only
the worst possible error in the calculation of the steady state concentrations
of the copper hydroxy and copper carbonate species.
+ 2 ...
Table 3-5 contains ratios of equilibrium to steady state Cu activities
for duplicate sets of four waters representing the various types of waters
considered in the six sets of copper toxicity tests; ambient Lake Superior
water, lake water with elevated alkalinity and pH, lake water with added
humics, and lake water with lowered alkalinity and pH. The table lists by
column the type of water, the duration of the experiment which was generally
+ 2
longer than the time required to reach equilibrium, the ratio of the Cu
+2
activity in the Teflon bottle at the end of the experiment to the Cu activ-
ity in the flow-through chamber before the transfer to the Teflon bottle, the
+2 • • +2
ratio of the equilibrium Cu activity to the steady state Cu activity, and
the difference between the pH in the flow-through chamber before transfer to
the Teflon bottle and the pH in the Teflon bottle at the end of the experi-
+2
ment. Absorption is taken into account by computing the ratio of Cu
activities listed in column 5 from the product of the ratios listed in columns
3 and 4.
3-51
-------
TABLE 3-5
+2 ... +2
Comparison of Equilibrium Cu Activities to Steady State Cu Activities
Time
(Cu+2)
Bott le
(Cu+2)
Flow
(CUT}Flow
(CuT)Boltle
(Cu+2>
bqui 1
(Cu 2)ss
(pH)
(pH)
F1 ow
Bot t Le
Ambient Lake H^O
Ambient Lake H^O
5 hrs
7 hrs
0.95
0.88
1 .03
1 .02
0.98
0.90
0.04
0.03
Alk =
Alk =
, PH =
, PH =
7 hrs, O.N.
7 hrs
0.89
0.91
0.96
a
0.86
0.91
-0.03
0.01
llumics (2.5 mg/L
TOC)
Humics (2.5 mg/L
TOC)
12 hrs, O.N.
10 hrs
0.78, 0.78
0. 79
1 .06
0.98
0.83, 0.83
0.77
0.11
-0. 19
Alk =
Alk =
, pH =
, pH -
10 hrs, O.N.
7 hrs, O.N.
0.69, 0.68'
0.79, 0.70
l,12t
1.15
0.78, 0.76
0.80, 0.80
0.08
0.02
3
k Sample misplaced
Overn ight
-------
The equilibrium to steady state Cu ratios listed in column 5 of Table
+2
3-5 show that the difference between the equilibrium and steady state Cu
activities in ambient lake water and in lake water with elevated alkalinity
and pH is at most less than 10% and 15%, respectively. The results are con-
sistent with the good agreement in most cases for waters with elevated
alkalinity and/or pH between the measured dissolved copper and the sura of
calculated concentrations of major inorganic species based on the use of
+ 2
measured steady state Cu activities in equations (1-1) to (1-7).
+2
The ratios of equilibrium to steady state Cu activities listed in
column 5 of Table 3-5 show that the difference between the equilibrium and
+ 2
steady state Cu activities in lake water with added humics, and in lake
water with lowered alkalinity and pH is, at most, less than 25%. That is
somewhat greater than the difference in ambient lake water or in lake water
with elevated alkalinity and pH. Actual differences between equilibrium and
~2
steady state Cu activities of greater than 15% could conceivably introduce
some significant error in using equilibrium equations (1-1) to (1-7) to cal-
culate steady state concentrations of chemical species from the steady state
Cu ^ activity. However, as stated previously, the differences indicated by
Table 3-5 represent only worse case situations because the species whose con-
centrations are calculated from equation (1-1) to (1-7), may be closer to
+2 +2 .
equilibrium with Cu than the Cu is to equilibrium with the system as a
whole.
3.10 THEORETICAL COPPER-ORGANIC COMPLEXATION BASED ON THE SLOPES OF THE
ELECTRODE RESPONSE
The accuracy and utility of using the cupric ion electrode to determine
Cu ^ activities in natural waters has been questioned by several groups (46,
74, 84-86). Ions which appear to interfere with the electrode include
Ag (72), Hg ^(72), Fe ^(72, 84) and CI (72, 74, 84). In addition, the elec-
+ 2
trode is susceptible to oxidative dissolution of Cu in aqueous solution
(particularly in acidic solutions) and apparently becomes slowly oxidized when
stored in air (73-75). Humics may also possibly interfere with the electrode
either by coating the membrane (86) or by changing the standard potential of
the electrode (46).
3-53
-------
Although interferences by Ag+, Hg+^, or Fe+^ are possible in some natural
waters, the concentration of all of those cations in Lake Superior water,
which was used as the diluent water for all in lab toxicity tests, are rela-
tively low (57) and therefore should not have posed a significant problem.
The use of a double junction reference electrode should have protected the
solutions in which measurements were made from Ag+ contamination. Although
CI concentrations in Lake Superior water are also low, several tests within
set 1 of the toxicity experiments involved adding CaCl , MgCl„ or NaCl to a
-3 - -
nominal concentration of 4x10 M CI . However, those levels of CI are still
well below (eg > lOx less) those levels reportedly required to adversely
+ 2
affect electrode performance (72, 87-89). Although the Cu determination in
the NaCl solution.appeared to be too high, the CI concentration in that
+2
solution was less than in the CaCl^ or MgCl^ solutions *rtiere the Cu deter-
minations were in close agreement with conputer estimates.
Although we observed evidence of Cu ^ dissolution off the membrane, it
did not appear to be a significant problem except in solutions with a pH below
+ 2
6. Furthermore, the Cu activities in the lab flow through chambers were
measured in a flow-through electrode chamber designed to carry solution
exposed to the electrode rapidly from the electrode and to continuously impact
the electrode membrane with fresh solution. Therefore, it is unlikely that
+ 2
the electrode measured significant quantities of any Cu built up from
dissolution off the electrode membrane. The electrode was routinely polished
daily and was polished more frequently if measurements within duplicate tanks
on the same day varied by more than 2mV. Therefore, it is also unlikely that
any tarnishing of the electrode membrane through solution oxidation, air
+2
oxidation, or coating by hianics significantly affected Cu determinations.
The suggestion that humics could possibly substantially alter the standard
potential of the electrode pair is difficult to prove or disprove and has not,
as yet, been confirmed or disproven by other groups. Finally, we saw little
evidence of the cupric ion electrode responding to either copper hydroxy or
copper carbonate complexes as was reported by Wagemann 1980 (85). Although
+2
the Cu determination in the high alkalinity and pH water of set 3 appeared
to be high, measurements in other high alkalinity and/or pH waters appear to
agree closely with computer estimated values.
3-54
-------
If the potential E of a cupric ion electrode/double junction reference
electrode pair is plotted versus the log of the dissolved copper, the slope of
the plot at any given point can be shown, assuming Nernstian behavior to be
given by
- N - N(CU+2) (3-1>
dl°8CuDISS DISS
where
N = Nerstian slope
D * CuDISS/Cu*2
+ 2
If the dissolved copper consists only of Cu , Copper hydroxy and copper
carbonate species, D will only be a function of pH and alkalinity so that
dE - N _ N(Cu+2)Yi_D\ ldp\ dAlk (3-2)
%
U
dlogCuDISS pfp^llk dCuDISS v A >pH dCuDISS
Since JpH and dAlk are both negative and
dCuDISS dCuDISS Vat"fclk
and/ qD \ are both positive, the grouped expression
V^H
within the brackets of the second term of equation (3-2) is negative and the
overall second term of equation (3-2) is positive. Therefore, the slope of
the electrode response, even in organic free water, is expected to be greater
than Nernstian, provided that the added copper is sufficient to significantly
decrease the pH and/or alkalinity of the water.
Table 3-6 lists by column for each of the test waters for each of the six
sets of copper toxicity experiments the chemical substituents added to the
test waters, the pH, alkalinity and hardness of the test waters, the computed
+2
free copper (Cu concentration) percentages of dissolved copper based on
+ 2
electrode determinations of Cu activities, the REDEQL computed free copper
percentages based on the dissolved copper assuming no organic complexation of
copper, the ratio of the conputed free copper percentages based on electrode
determinations to those based on REDEQL assuming no organic complexation, the
percentage difference between dissolved copper,and the sum of all major
3-55
-------
TABLE 3-6
Comparison of Computed Free Copper Percentage of Dissolved Copper
at the LC50 Point Based on Electrode Determinations to REDEQL
Computed Free Copper Percentages Based on Inorganic Speciation Alone
for the Six Sets of Copper Toxicity Tests
PH
Alk
Hard
% Cu + 2
Elect rode
% Cu
REDEQL
Rat io
Test s
% Cu-Org
Slope
mV/Log Unit
Set 1 - 4/83 (Ca*2, Mg*2, Na* Effects)
ui
i
cr>
Ambient Hardness
2x10_^M CaCl
2xl0J?M MgCl
2x10 M NaCl
8.13 42.6
8.05 43.1
8.06 43.3
8.12 43.3
47.0
243
255
47.2
2.03
4.96
4. 10
5.30
Set 2 - 6/83 (pH Effects at Ambient Alkalinity)
3. 26
4.88
4.83
3.60
0.62
1.02
0.85
1.47
46. 5
14.4
26. 7
0
40. 5
33.7
36.5
35.9
CO^ Bubbling
CO Bubbling
Ambient pll
NaOH Added
6. 58
7.35
8.06
8. 72
48.9
44.5
42.8
47.4
49.2
46.2
45. 1
45.2
13.5
10.0
3.98
0. 75
51.8
16.5
3.82
0. 75
0. 26
0.61
1.04
1.00
77.3
47.9
11.6
11.7
42. 1
39. 1
37. 5
34. 1
Set 3 - 7/83 (pH Effects at 3X Ambient Alkalinity)
(Na ,K )HC0
CO,
(Na+,K*)HCO , CO
(Na+,K+)HC03-
Hcl Added
7.15
155
46.2
11.2
9.91
1.13
2.02
27.6
7.94
148
45.0
1.94
1. 77
1. 10
5.98
28.0
8. 56
150
45.0
0. 58
0. 38
1.52
0
30. 3
7. 16
26.4
45.0
20.4
32.2
0.63
43.5
33.8
-------
TABLE 3-6 (continued)
CO
I
cn
% Cu+2
% Cu+2
Test s
SI ope
PH
Alk
Hard
Electrode
REDEQL
Rat io
% Cu-Org
mV/Log Unit
Set 4 - 10/83 (Effect
s of
Clay, Humics)
Ambient hake Water
8.03
40.8
42. 5
3. 74
4. 24
0.88
24. 1
38. 1
Suspended CLay
(70 NTL)
8.22
46.5
48.9
2.01
2.46
0.82
29. 7
39.3
Hum i c s (5 mg/L TOC)
8.00
44.2
45.5
0. 50
4. 20
0. 12
90. 1
60. 7
Clay and Humics
8. 22
48. 5
46.8
0.47
2.36
0. 20
83.2
56.4
Set 5 - 11/83 (Effect
s of
Variable
Humics
Concent rat ions)
Ambient Lake Water
8.02
43.0
45.0
4. 32
4. 17
1.04
11.3
36.0
Humics (1.25 mg/L
TOC )
8.00
42.2
45.0
2.19
4. 36
0.50
57.4
45.1
Humics (2.50 mg/L
TOC)
8.00
43.8
45. 5
1.13
4.28
0.26
77.6
54. 3
Humics (5.0 mg/L
TOC)
8.02
42.8
45.0
0.47
4. 17
0. 11
90.4
57.0
Set 6 - 5/84 Alkalini
ty Effects at
Amb ient
pH)
HC1 Added, C02
Removed
7.84
17.0
44.0
11.5
12.0
0.96
18.0
41.0
Ambient lake Water
(Na ,K )HCO
, ~ +x 3-
(Na ,K )HC0
7.93
42.0
44.0
5.02
5.20
0.97
17.5
35.3
8.00
8.00
161
318
44.0
45.0
1.71
0.90
1.49
0. 79
1. 14
1. 14
4. 1
3.8
30. 1
29.6
-------
+2
inorganic copper concentrations based on electrode determinations of Cu
activities (designated % Cu-Org) , and the linear regression derived slopes of
the electrode response.
The slopes of the electrode response listed in Table 3-6 are all far
greater than Nernstian except the ones listed for waters with higher than
ambient lake water alkalinity. The slopes of the electrode response in waters
with higher than ambient alkalinity are closer to Nernstian presumably because
the added buffering capacity makes and dAlk small so that the slope
dCu_T_. dCu___.
DISS DISS
given by equation (3-2) reduces to approximately Nernstian. However, even
though greater than Nerstian slopes are expected in test waters run at the
lower alkalinities, the magnitude of the observed greater than Nernstian
slopes cannot be explained alone by the relatively small and often negligible
observed decreases in pH and alkalinity with increasing dissolved copper
within the low upper range of copper used for our toxicity tests. Therefore,
we have postulated the existence of one or more organic ligands which have a
+ 2
relatively strong affinity for Cu , but which have low activities such that
+ 2
coraplexing with Cu substantially lowers the ligand activity. In such a
case, the slope of the electrode response would be given by
+ \ »fesA ik
t^'Alk^DISS dCuDlSS \»V
dlozCu ,„- l\ rtnH / . dCu \C»Alk/ dCu \ pL./ dCu
u Alk \ DISS \ i' DISS
AlkpH
where
L = activity of i'*1 ligand
and the slope would depend not only on decreases in the pH and alkalinity, but
also decreases in organic ligand activities with increasing dissolved copper.
The ratios of the electrode determined free copper percentages to the
REDEQL determined free copper percentages at the LC50 point, assuming no
organic complexation are listed in column 7 of Table 3-5. The ratios for 12
of the 24 test waters of the six sets of copper toxicity experiments are
within the range 0.85-1.15. Of the 12 test waters outside of this range, six
3-58
-------
exhibit very low ratios and contain added hianics and/or clay which can complex
substantial amounts of copper beyond that predicted by REDEQL based on
inorganic speciation alone. Of the remaining six test waters outside the
range, one is the ambient lake water of set 1, and three are waters with lower
than ambient pH and/or alkalinity all of which exhibit low ratios. Copper-
organic complexation could possibly be more extensive in waters with lower
alkalinity and/or pH due to a reduction in competition between the organic
-2 +2
ligands, OH , and CO. for Cu even though lower pH waters may have
+ 2 +
increased competition between Cu and II for organic ligands. The two
remaining waters (the NaCl water of set 1 and the high alkalinity and pH water
of set 3) had high ratios of approximately 1.53, which we cannot explain.
The non-humic waters listed in Table 3-6 with the greatest non-Nernstian
slopes also generally have the lowest electrode/REDEQL free copper percentage
ratios and the highest theoretical copper-organic fraction percentages. How-
ever, those observed correlations are the result of the affect6 of a steeper
than Nernstian slope on related computations, and would be observed regardless
of whether or not a copper-organic fraction contributed significantly to the
non-Nernstian slope. Therefore, we attempted to determine if a significant
copper-organic fraction existed in Lake Superior water despite the low organic
content of the water. In doing so, we performed two sets of titration experi-
ments .
The first set of titrations was designed to observe the effects of
reduced organic content on the greater than Nernstian slope of the electrode
-3 2
response. The first set of experiments consisted of 10 M Cu(NO^) titrations
of Lake Superior water, UV irradiated Lake Superior water, Burdick and Jackson
HPLC grade water, UV irradiated Burdick and Jackson water, and Burdick and
Jackson water passed through a HPLC colunn. Figure 3-8 contains plots of the
+ 2
negative logarithm of the Cu activity versus the negative logarithm of the
total copper for the five titrations comprising the first set.
Figure 3-8 shows that it is possible for waters with supposedly lower
organic contents than Lake Superior water such as HPLC grade Burdick and
~2
Jackson water to exhibit even greater non-Nernstian slopes and lower Cu /Cu^
3-59
-------
6-
CvJ
+
Z)
O
CD
O
8-
a U V-B->- J
~ U V-L AKE
0 L A K E
4 HPL C-B+J
* B + J
6 -Log Cut 5
Figure 3-8 The effects of UV irradiation and/or HPLC clean up on copper titrations
of Lake Superior water and 3urdick * Jackson vater
3-60
-------
ratios at each point of a titration curve. It is unclear whether the addition
of N'aHCO^ to the Burdick and Jackson water to increase the alkalinity to that
of lake water also introduced organic contamination to the water. btowever ,
the organic content of the Burdick and Jackson water even after the addition
of NaHCO. should have been less than the average lppm TOC of Lake Superior
.J -3
water since the amount of NaHCO. added (1x10 M) was small. The large,
+2
steeper than Nernstian slope and relatively low Cu /Cu^ ratios for the
Burdick and Jackson water titration could have been due to some factor other
than copper-organic complexation such as electrode interference. . However, it
is unlikely since Figure 3-8 shows that UV irradiation decreases the slope of
the Burdick and Jackson water to that comparable to lake water, and increases
+ 2
the free Cu /Cu^ ratios to above those for lake water presumably through
photochemical degradation of organics. Passing the Burdick and Jackson water
down a C—18 reverse 'phase HPLC column to remove organics also decreases the
+2
slope of the titration curve and increases Cu /Cu^ ratios, although not to
the extent that UV irradiation did as can be seen in Figure 3-8.
Although it appears that copper-organic complexation is responsible for
most of the non-Nernstian characteristics of the Burdick and Jackson water
titration curve, it is not certain. That is because neither UV irradiation
nor HPLC results in 1inear-Nernstian titration curves despite the elimination
of pH change contributions to the non-Nernstian behavior [by performing con-
stant pH (8.00) titrations] and the negligible contribution of alkalinity
changes (as demonstrated by the small differences between pre-titration and
post-titration alkalin ities). • Nevertheless, both UV irradiation and HPLC
greatly reduce the non-Nernstian characteristics of the Burdick and Jackson
water titration curve so that the failure to produce completely Nernstian
curves may be due to insufficient organic removal.
The above discussion showed that it might be possible for waters with
comparable or even lower organic content to exhibit even greater non-Nernstian
behavior than Lake Superior water through copper-organic complexation. How-
ever, figure 3-8 shows that the UV irradiation of Lake Superior water had
virtually no effect on the titration curve except at the lowest section of the
curve. That is contrary to what would be expected if substantial amounts of
3-61
-------
copper-organic complexation occurred in Lake Superior water. Unfortunately,
the UV system used was not set up for continuous UV irradiation, so UV
irradiation was performed during three successive siphonings of the water
through the photochemical reactor cell. Since there was no operational TOC
analyzer available, it was not possible to determine if the resultant duration
and intensity of the UV irradiation was sufficient to appreciably reduce the
organic content of the Lake Superior water. Although by using the same pro-
cedure and equipment, UV irradiation did appear to appreciably reduce the
organic content of the Burdick and Jackson water, the organics in Lake
Superior water, particularly the naturally occurring higher molecular weight
ones, may be more resistant to UV degradation. Therefore, it is unclear
whether the failure of UV irradiation to appreciably affect the titration
curve for Lake Superior water was due to an insufficient reduction in the
organic content of the water, or to factors other than copper-organic complex-
ation such as electrode interferences being primarily responsible for the
non-Nernstian characteristics of the titration curve.
The failure of the first set of titrations to conclusively demonstrate
the presence of significant copper-organic complexation in Lake Superior
water led to the performance of a second set of six titrations designed to
+2
observe the effects of ionic strength, cations competitive with Cu for
+ 2+2
organic ligands such as Ca and Zn , alkalinity and pH on the titration
curves of Lake Superior water. The second set of titrations consisted of 10 J
3
Cu(N0 )„ titrations of Lake Superior water and seperate Lake Superior waters
-3 -A
with the following nominal concentrations: 2x10 Ca(N0 ) , 10 M Zn(N0 ) ,
-3 -3 -3
6x10 M NaNO^ and 4x10 M NaHCO^. Titrations of the 4x10 M NaHCO^ waters
were performed at pHs of 8.00 and 8.50. Figure 3-9 contains plots of the
+2
negative logarithm of Cu versus the negative logarithm of total copper for
each of the six titrations of the second set.
The plots in figure 3-9 are consistent with the postulate that signifi-
cant copper-organic complexation does occur in Lake Superior water. For
-4
example, the slopes of the titration curves for lake water containing 10 M Zn
_3
W°3>2 or 2x10 M CaCNO^^ are somevrtiat closer to Nerstian (e.g., for the type
of plots in figure 3-8 and 3-9, a Nernstian curve is linear with a slope of 1)
3-62
-------
- 3 -h2
° 2 x 10 m La
-4 _ +2
a IX 10 M Zn
»Lake Water
#6XIo"M Na+
pH=8.00
V\LK=250 MG/L
p H=8.5 0
'A L K =2 5 0
6 -Log Cut 5
T
Figure 3-9 The effects of Ca+2, Zn+2t Na""", pH and alkalinity on copper titrations
°f Lake Superior water
3-63
-------
than Lake Superior water, although both curves are curvilinear. The titration
-3
curves for lake water to which 4x10 M NaHCO^ was added were actually linear
with slopes much closer to Nernstian than Lake Superior water. The effects of
+2+2
Ca , Zn and NaHCO. on titration curves for Lake Superior water cannot be
-3
due to increases in the ionic strength because the addition of 6x10 M NaNO^
has very little effect on the titration curve as can be seen in figure 3-9.
The observation that CatNO^^ and ZnCNO^)^ affect the slope of the titration
curves *Aiereas NaNO^ does not supports the postulate that significant copper-
organic complexation is present in Lake Superior water. The reason is that
+2 +2 +
Ca and Zn both exhibit much higher affinities for organics than does Na
+2
and therefore, could conceivably effectively compete with Cu for organic
ligands.
The effect of NaHCO^ and the associated increase in alkalinity on the
slope of the titration curves is probably due to increased competition of
-2 +2
CO^ with organic ligands for Cu , which may lead to a reduction in copper-
organic complexation. Likewise, the increase in pH to 8.50 probably leads to
+ 2
increased competition between OH and organic ligands for Cu . Table 3-7
shows that relatively good agreement can be obtained between computed free
copper percentages based on electrode determinations and computed free copper
percentages based on REDEQL for the Lake Sueprior water titration, the 4x10 JM
-3
NaHCO^ lake water titration at pH 8.00 and the 4x10 M NaHCO^ lake water
titration at pH 8.50 if variable glycine concentrations are input into the
REDEQL program. The variable glycine concentrations are indicated in Table
3-7 footnotes, and all are either comparable or well below that which would
correspond to the reported average organic nitrogen (0.1 ppm) and TOC (1 ppm)
for Lake Superior water (92, 57). The fact that decreasing assuned glycine
concentrations with increasing copper are required to maintain reasonable
agreement between electrode and REDEQL computations may correspond to the
gradual saturation of binding sites and/or to increased negative interactions
between bound and unbound sites of copper binding organics in Lake Superior
water. Although the use of glycine is not meant to accurately model the
binding capacity of Lake Superior water, it does show that an organic with a
somewhat intermediate binding affinity for copper compared to many other
organic ligands and at concentrations at or well below that corresponding to
3-64
-------
TABLE 3-7
Comparison of Computed Free Copper Percentage of Total Copper Based on
Electrode Determinations to REDEQL Computed Free Copper Percentages Based
on Assumed Variable Glycine Concentrations
-Log Cu
% Cu
Electrode
X Cu*"2
REDEQL
Ratio
Ambient Lake Water
6.49
6.33
6.03
5.73
5.47
5.15
4.88
0.90
1.75
2.36
3.10
4.09
4.39
4.92
4.19
0.73'
4.19'
1.84C
4.19
2.38C
4.19°
2.95*
4.19
3-65!
4.19'
3.00*
4.19*
4.09
a
a
a
0.21
1 .23
0.42
0.95
0.56
0.99
0.74
1.05
0.90
1.12
1.05
10
1 7
20
4x10 3M NaHCOj, PH = 8.00
6.57
6.29
6.01
5.69
5.43
5.13
4.83
0.36
0.47
0.58
0.65
0.81
0.93
1.05
0.47
0.70
0.79
0.84
0.92
0.94
0.94
0.77
0.67
0.73
0.77
0.88
0.y9
1 . 12
,-3
4x10 M NaHCOj, pH * 8.50
6.57
6.30
6.01
5.69
5.14
4.83
13
16
19
21
30
30
0.10
0.19
0.22
0.24
0.26
0.24
1 .30
0 .04
0.8b
0.88
1.15
1.11
k Glycine Concentration
- LOG [Gly] - 5.10
d - LOG [Gly] = 5.40
- LOG [Gly] = 5.58
f - LOG [Gly] = 5.70
- LOG [Gly] - 6.00
^ - LOG [Gly] * 6.30
- LOG [Gly] c 6.60
= 0
3-65
-------
the average organic nitrogen and TOC of Lake Superior water, can bind substan-
tial amounts of copper. Furthermore, the good agreement between electrode and
-3
REDEQL computations in Table 3-7 for the 4x10 M NaHCO^i particularly the one
at pH 8.50 shows that increased alkalinity and/or pH can effectively modify
non-Nernstian characteristics caused by copper-organic coraplexation (in this
case, the assumed copper-glycine complexation).
Although neither the set 1 nor set 2 titrations conclusively show the
presence of significant copper-organic complexation in Lake Superior water,
they do support that postulate with the exception of the failure of UV irradi-
ation to significantly affect the Lake Superior titration curve, which may
have been due to an insufficient reduction of organic content.
3.11 MULTIPLE LINEAR REGRESSIONS
Despite apparent variations in the toxicity/unit concentration of toxic
copper species, an attempt was made to compute the toxic ities/unit concen-
tration of the five major inorganic copper species (Tj.TjjT^.T^.Tj) and an
average toxicity/unit concentration for the copper-organic fraction (T,) by
+2
regressing Y = 1/Cu ^50 on Xl' *2' ' ' ' X6 e<1uation (1-10) where Xj, Xj,
. . . X^ are related to the computed concentrations at the LC50 point of the
various copper species by the definitions given for equation (1-10). The
first attempt was made using the calculated concentrations of all five major
inorganic species and the theoretical copper-organic fraction for 14 of the 24
test waters. Of the six toxic ities/unit concentration calculated, four were
statistically different from zero at p = 0.05. However, the calculated
+ / \ -2
toxic it les/umt concentration of CuOH and Cu(CO^)^ > which were two of the
four statistically different from zero, were negative which has no physical
meaning.
The multiple linear regression was then repeated by regressing Y on only
those variables related to inorganic species and using data only from the 9
of the 14 test waters used originally which had negligible theoretical copper-
-organic fractions. Although the resulting regression equation accounted for
greater than 98% of the variance in Y, only two of the five computed toxici-
3-66
-------
ties/unit concentration were statistically different from zero at p ¦ 0.05.
Furthermore, of the two statistically significant toxicities/ unit concentra-
_2
tion, the one for CuCCO^^ was again negative.
The unsuccessful attempt to compute toxicities/unit concentration from
multiple linear regression may be primarily due to the variation and depend-
ence of toxicities/unit concentration on such factors as pH, and perhaps time,
as reflected by the variation in larvae sensitivity with time depicted in
Figure 3-7.
3-67
-------
4. CONCLUSIONS AND RECOMMENDATIONS
In this chapter, our conclusions with respect to the feasibility of
developing both a chemical speciation method and a toxicity factors method for
predicting copper toxicity in site waters will be presented along with
recommendations for future research. However, a summary of discussions on the
apparent dependence of toxicities/unit concentration on the values of general
water quality parameters and/or time, and on the possible presence of toxic
copper-organic complexes in Lake Superior water will be presented first.
4.1 THE APPARENT DEPENDENCE OF TOXICITIES/UNIT CONCENTRATION ON THE VALUES OF
GENERAL WATER QUALITY PARAMETERS AND/OR TIME
An analysis of the results of set 1 of the copper toxicity tests in
Section 3.1 indicated that the toxicities/unit concentration of at least some
toxic copper species probably decrease with increasing hardness. The increase
-3
in the dissolved copper LC50 in going from ambient lake water to the 2 x 10
+2
M CaC^ water cannot be contributed to by any proportional changes in Cu or
any other major inorganic copper species because the proportion of all such
species increase (see Table 3-3). The increase in the dissolved copper LC50
may be partially due to a decrease in the proportion of a possibly toxic
+2 +2
copper-organic fraction perhaps caused by competition between Ca and Cu
for organic ligands. However, it appears likely that the increase in the
dissolved copper LC50 is due primarily to the reduction in the toxicity/unit
concentration of one or more toxic copper species with increasing hardness.
-3
The reason is that in going from the ambient lake water to the 2 x 10 M
CaCl., water, the combined increase in the concentrations at the LC50 point of
+2
Cu and all other major inorganic copper species is much greater than the
calculated copper-organic concentrations in either water (see Table 3-2).
Therefore, it is unlikely that any reduction in the concentrations of copper
organic complexes could satisfy equation (1-14) alone without some reduction
in the toxicities/unit concentration of one or more toxic copper species also
occurring.
4-1
-------
An analysis of the results of set 2 of the copper toxicity tests in
Section 3.2 indicated that the toxicities/unit concentration of at least some
copper species also decreased with increasing pH, at least over the range of
pH 6.6 to pH 8.0. The increase in the dissolved copper LC50 with increasing
pH over that range appears to be at least partially due to the decrease in the
+2
proportion of Cu . It may also be due to the decrease in the copper organic
-2
proportion which may be caused by the increased competition between CO. ,
+2
OH , and organic ligands for Cu . However, the increase in the dissolved
copper LC50 is also probably at least partially due to a reduction in the
toxicity/unit concentration of one or more toxic copper species with
increasing pH. The reason is that despite the decrease in the proportion of
+2 +2
the Cu , the concentration of Cu at the LC50 point remains relatively
constant, whereas the concentrations of all of the other major inorganic
species greatly increase with increasing pH (see Table 3-2). Therefore, it is
unlikely that any decrease in the concentration of the much smaller copper-
organic fraction over that pH range could satisfy equation (1-14) alone
without some reduction in the toxicities/unit concentration of one or more
toxic copper species also occurring.
An analysis of the results of set 4 and set 5 in Section 3.4 and 3.5
indicated that the presence of added humics either increased the toxicity/unit
concentration of one or more toxic copper species, or formed toxic complexes
with copper, or both. The reason is that in going from ambient lake water to
any of the waters with added humics, the concentration at the LC50 point of
+2
Cu and all other major inorganic copper species decrease. Therefore, in
order for equation (1-14) to be fulfilled, the toxicity/unit concentration of
one or more toxic copper species must increase in the presence of humics
and/or the humics must form at least some toxic complexes with copper.
It has been shown that changes,in the proportions and concentrations of
inorganic copper species can often not account for the differences between
LC50 values for different test waters. However, it is conceivable, at least
from the standpoint of fulfilling equation (1-14), that such changes could
fully account for the differences between the LC50 values for the two highest
pH waters of set 2, the three waters of set 3 at higher than ambient
alkalinity, and the three highest alkalinity test waters of set 6. There does
not appear to be any significant copper-organic complexation in any of those
waters, nor is there any substantial evidence to indicate that the toxicities/
4-2
-------
unit concentration of copper species vary between the waters. However, our
attempts to determine values for the toxicities/unit concentration of the
inorganic copper species through multiple linear regression on the data for
those test waters was unsuccessful as discussed in Section 3.11. One possible
explanation is that the toxicities/unit concentration of at least some of the
copper species may have varied between the waters as a function of pH and/or
alkalinity. However, the failure could have also been due to a possible
change in the sensitivity of the organisms to copper toxicity with time.
The postulate that the sensitivity of the organisms to copper toxicity is
changing with time is based on Figure 3-7. Figure 3-7 shows that in going
from the ambient lake water in set 1 to that of set 3 and set 4, the dissolved
+2
copper LC50s increase monotonically by a factor of over 3, while the Cu
LC50s increase monotonically by a factor of over 5. Therefore, it appears
that the sensitivity of fathead minnow larvae to copper toxicity was
decreasing with time, which would be reflected by an apparent decrease in the
toxicities/unit concentration of copper species. However, it should be noted
that the ambient lake water of set 1, in which the lowest dissolved copper
+2
LC50 and Cu LC50 were observed, was also calculated to have by far the
highest copper-organic fraction among the ambient lake waters. Therefore, it
is possible that variations in the copper-organic fraction in ambient lake
water could account for some of the variations in the dissolved copper and
Cu+2 LC50s.
A.2 POTENTIALLY TOXIC COPPER-ORGANIC COMPLEXES
There was substantial evidence presented in Section 3.10 to support,
though not conclusively, the postulate that substantial copper-organic
complexation can occur in Lake Superior water at or below ambient pH and/or
alkalinity despite the relatively low organic content of Lake Superior water
of approximately 1 ppm TOC. The evidence included the magnitude of the
steeper than Nernstian slopes of the electrode response in ambient lake water
and lake water with lowered pH and/or alkalinity (see Table 3-6), the copper
titrations of Lake Superior water discussed in Section 3.10, and the computer
calculations of copper-organic complexation presented in Table 3-7 which used
glycine as a substitute for organic nitrogen in Lake Superior water.
4-3
-------
Of course, there is still a possibility that the observed differences
between dissolved copper and the sum of inorganic copper concentrations based
+2
on electrode determinations of Cu activity in lake waters with ambient or
lower pHs and/or alkalinities, are due to interferences with the electrode.
However, the major known chemical interferences with the electrode do not
appear to be at high enough concentrations in Lake Superior water to cause a
+2
significant effect. The agreement between electrode determinations of Cu
+2
concentrations and REDEQL computations of Cu concentrations based on
inorganic speciation alone for some ambient lake waters and almost all lake
waters with elevated pH and/or alkalinity is good with ratios varying from
0.85 to 1.15, as can be seen in Table 3-6. Also, the slopes of the electrode
response in such waters, particularly the ones with elevated pH and/or
alkalinity, are generally close to Nernstian as can also be seen from
Table 3-6.
The possibility that the large differences between the slopes of the
electrode response in waters with lower than ambient pH and/or alkalinity and
in waters with higher than ambient pH and/or alkalinity was due to either an
ionic strength or pH effect was also investigated. However, the addition of
-3
6 x 10 M NaNO^ to ambient lake water had virtually no effect on the non-
Nernstian response of the electrode in lake water (see Section 3.10), which
indicated that large increases in the ionic strength without an associated
increase in alkalinity does not modify non-Nernstian responses. Also, even
though the slopes of the electrode response are greatly dependent upon pH in
Lake Superior water, pH does not appear to directly effect the slope of the
electrode response since Nernstian slopes can be obtained at both low pH in
0.01 M acetate buffer or at higher than ambient pH in lake water.
The above discussion indicates that the large differences in the slope of
the electrode response in lake water as a function of pH and alkalinity may be
due to the dependence of copper-organic complexation on those variables. In
particular, the amount of copper-organic complexation may decrease with
increasing pH and/or alkalinity due to the increased competition between OH ,
-2 +2
CO, , and organic ligands for Cu . Although increasing pH will decrease the
+ +2
competition of H with Cu for organic ligands, the copper-organic complex-
ation does appear to decrease with increasing pH, at least at pHs above 6.6.
4-4
-------
Therefore, the decrease in the competition between H and Cu for organic
ligand with increasing pH above 6.6 does not appear to be sufficient to offset
-2 +2
the increased competition between OH , CO^ , and organic ligands for Cu
Although there is substantial evidence to support the postulate that
significant amounts of copper-organic complexation can occur in lake waters
with.ambient or lower alkalinities and/or pHs, it is not known whether such
copper-organic complexes contribute significantly to toxicity. The reason is
that in each case in which the assumption of the presence of toxic
copper-organic complexes can help explain the differences between the
toxicities of two test waters, an assumption that the toxicities/unit
concentration of the inorganic species change can explain the differences
equally well. The reverse is not always true because in some cases it is
apparent that neither changes in the proportion of inorganic copper species
nor in copper-organic species can completely account for the differences in
the LC50 values between two test waters. That can be seen from the earlier
discussion on the apparent decrease in the toxicities/unit concentration of
copper species with increasing hardness and/or pH.
4.3 CONCLUSIONS WITH RESPECT TO THE FEASIBILITY OF DEVELOPING A CHEMICAL
SPECIATION METHOD FOR PREDICTING COPPER TOXICITY IN SITE WATERS
The results of our work, as summarized in Sections 4.1 and 4.2, indicate
that the feasibility of developing a chemical speciation method for predicting
LC50 values in site water from non-site toxicity data bases and the average
site specific values of general water quality parameters is very low. The
results have shown that differences in the LC50s between two test waters can
often not be fully explained by changes in the proportion or concentration of
inorganic copper species alone. Therefore, it appears that changes in the
toxicities/unit concentration of at least some toxic copper species and/or
changes in the proportions or concentrations of toxic copper-organic complexes
also occur between such waters. Unfortunately, the chemical speciation method
is not well suited to handle either possibility.
As previously discussed in Section 3.1, the apparent decreases in the
toxicities/unit concentration of toxic copper species with increasing hardness
4-5
-------
could possibly be handled by determining several sets of toxicities/unit
concentration, one for each of several values of hardness covering the range
of hardness normally encountered in natural waters. However, such procedures
would be subject to large error by the restricted ranges of hardnesses used
for the determination of each set of toxicities/unit concentration. Although
the problem with hardness could possibly be approximately handled, the
apparent decreases in the toxicities/unit concentration of copper species with
increasing pH, at least in lake waters with ambient or lower alkalinity,
cannot be handled by the chemical speciation method unless the toxicities/unit
concentration of all of the copper hydroxy species are negligible which is
unlikely. The reason, as discussed in Section 3.2, is that to determine the
toxicities/unit concentration of copper hydroxy species, it is necessary to
+2
vary the proportion of such species with respect to Cu which can only be
done by changing the pH. However, if in changing the pH, the toxicities
conentration of the copper hydroxy species are also changed, the values of the
toxicities/unit concentration cannot be determined.
Although apparent changes in the toxicities/unit concentration of copper
species may instead, in some cases, be at least partially due to changes in
the concentrations of toxic copper-organic complexes, the presence of toxic
copper-organic complexes is potentially equally detrimental to the development
of a chemical speciation method. The reason is that if more than a few
copper-organic complexes contribute significantly to toxicity, it would
probably be impossible to identify all such complexes, calculate the
associated ligand activities and determine the toxicities/unit concentration
of all of the complexes. Even if that could be done for Lake Superior water,
the types of copper-organic complexes in another water would probably be much
different. The same problem will also be encountered if copper-humic or
copper-clay complexes are toxic.
As previously discussed, the differences between LC50 values for some of
the test waters could be explained by changes in the proportions or
concentrations of inorganic copper species alone without postulating that the
differences were also due to changes in toxicities/unit concentration and/or
in the proportions or concentrations of toxic copper-organic species.
4-6
-------
However, our attempts to determine the toxicities/unit concentration of the
major inorganic copper species from a multiple linear regression on the data
for those test waters were unsuccessful. It is not known whether the failure
to make such determinations was due to any changes in the toxicities/unit
concentration with pH, alkalinity, or time. It is also possible that the
failure was due to the inability of a simple additive non-synergistic,
non-antagonistic model as represented by equation (1-10) to explain toxicity.
However, there is insufficient data available to attempt to formulate a more
complex model.
4.4 CONCLUSIONS WITH RESPECT TO THE FEASIBILITY OF DEVELOPING A TOXICITY
FACTORS METHOD FOR PREDICTING COPPER TOXICITY IN SITE WATERS
The feasibility of developing a toxicity factors method for predicting
LC50 values in site waters from non-site toxicity data bases and the average
site specific values of general water quality parameters is greater than for
developing a chemical speciation method. The reason is primarily because any
changes in the toxicities/unit concentration of toxic copper species with the
change in a general water quality parameter should affect, and be reflected
by, the functional dependence of copper toxicity (e.g., LC50 values) on the
general water quality parameter causing the change (e.g., pH). However,
despite the greater feasibility of developing a toxicity factor method, there
may be major potential problems with its development as well.
The major potential problems with the development of a toxicity factors
approach are the possibilities that numerous copper-organic complexes and/or
copper-humlc complexes and/or copper-clay complexes are toxic. If that is the
case, it would be impossible to accurately predict toxicity in a site water
from general water quality parameters alone, since the types and
characteristics of such complexes would vary greatly from site to site.
However, even if such complexes are toxic, they do not appear to be toxic
enough to make criteria based on total or dissolved copper LC50 values
non-protective. That can be seen from the fact that dissolved copper LC50
values increase with increasing organic and/or suspended solids
concentrations. Therefore, if such complexes are shown to exert significant
toxicity, total or dissolved copper should probably be the measured form of
4-7
-------
copper chosen on which to base LC50s in the toxicity factors method. Although
those measurable forms of copper are probably much greater than the
bioavailable fraction, criteria based on them will probably be protective in
the wide variety of site waters which can be encountered.
If it is shown that the copper-humic and copper-clay complexes are
non-toxic, the accuracy of the toxicity factors methods in predicting toxicity
in site waters could probably be greatly improved by basing LC50s on a
measurable form of copper which could approximate the bioavailable fraction
better than dissolved or total copper, such as perhaps >1,000 molecular weight
exclusion ultrafiltration. For example, if humics and suspended clay affect
copper toxicity primarily by decreasing the bioavailable fraction through
binding and adsorption, basing LC50s on some measurable form of copper which
closely approximates the bioavailable fraction should make such LC50s much
less dependent on humic and suspended clay concentrations than total or
dissolved copper LC50s. In addition, if the lower molecular weight
copper-organic complexes are also found to be relatively non-toxic compared to
inorganic species, it might be advantageous to base LC50s on the sum of all
major inorganic coppercomplex concentrations rather than 1,000 molecular
weight exclusion since such LC50s would be much less dependent of both high
and low molecular weight organic content. However, LC50s should probably not
+2 +2
be based on Cu alone due to the apparent non-monotonic dependence of Cu
LC50s on pH in water with moderately low alkalinities as is shown in Figure
3-B.
A.5 RECOMMENDATIONS
The results of copper toxicity tests so far have often shown that
differences in copper toxicity between different test waters cannot be fully
explained by changes in the proportions or concentration of inorganic copper
species alone, and that changes in the toxicities/unit concentration of toxic
copper species and/or changes in the proportions or concentrations of toxic
copper-organic complexes probably also generally occur. Although, as
discussed previously, it is not possible from the present data to determine
the relative contributions of changes in toxicities/unit concentration and
changes in the proportions or concentrations of toxic copper-organic complexes
4-8
-------
Co toxicity changes, either case is extremely detrimental to the development
of a chemical speciation method. Furthermore, it appears likely that changes
in several of the general water quality parameters do change the
toxicities/unit concentration of at least some of the toxic copper species.
Also, attempts to determine the toxicities/unit concentrations of copper
species, from multiple linear regression uaing data from test waters in which
no obvious changes in the toxicities/unit concentration had occurred, were
unsuccessful. Therefore, it is strongly recommended that efforts to develop a
chemical speciation method be dropped in favor of developing a toxicity
factors method. That does not necessarily mean that all work with the curpic
ion selective electrode should be dropped. As mentioned previously, some
linear combinations of the concentrations of copper species calculated from
electrode measured Cu activities could be a useful measurable form of copper
on which to base LC50s in the development of a toxicity factors method.
If copper-humic and copper-clay complexes are shown to be relatively
non-toxic, efforts could be concentrated on the development of a toxicity
factors method which uses, if possible, a measurable form of copper which will
minimize or hopefully eliminate the dependence of the LC50 values on huraic and
suspended clay concentrations. However, realistically, even if such a method
is developed, it may not be well received if the measurable form of copper
chosen is much more difficult to determine than dissolved copper. Therefore,
as a possible alternative, EPA should consider using dissolved copper as the
measurable form of copper on which a toxicity factors method should be based.
Of course, dissolved copper LC50s are very dependent upon humic and suspended
clay concentrations. Furthermore, the types and characteristics of humics and
clay vary so much, it would be very difficult to try to model such effects.
Therefore, if EPA decides to use dissolved copper as the measurable form of
copper on which to base the toxicity factors method, it should be realized
that any resultant criteria derived would probably be extremely conservative.
Nevertheless, if current equations relating total and dissolved copper LC50s
to hardness can be extended to multivariable equations relating total or
dissolved copper LC50s to other general water quality parameters as well such
as pH, any resultant criteria should be less conservative than they are now.
4-9
-------
The effects of alkalinity on copper toxicity appear to be small compared
to hardness as can be seen from a comparison of the results of set 1 and set 6
of the copper toxicity experiments. Therefore, since hardness and alkalinity
are so closely correlated in natural waters, it might be useful to break
hardness into carbonate hardness which would include any effect of alkalinity,
and non-carbonate hardness instead of considering alkalinity as a separate
variable.
The development of a multivariate equation relating LC50s to hardness and
pH or possibly carbonate hardness, non-carbonate hardness, and pH should be
based on the results of toxicity tests run in water which have typical
combinations of carbonate hardness, non-carbonate hardness, and pH values. In
general, waters will have a hardness equal to or greater than alkalinity so
that most waters will have some non-carbonate hardness. In addition, most
waters will have a pH somewhat lower than that expected if they were in
equilibrium with the atmosphere. Therefore, the tests should be run in waters
at several different hardnesses equal to the alkalinity covering the range of
alkalinity normally encountered in natural waters, and at pHs equivalent to
equilibrium with the atmosphere. In addition, tests should be run in waters
+2 +2
with non-carbonate hardness added in the form of Ca and Mg chloride and
sulfate salts, and pH lowered with CO2 bubbling. The base waters in which the
hardness is equal to the alkalinity and the pH equivalent to that of
equilibrium with the atmosphere should be made with mixtures of calcium and
+2 +2
magnesium bicarbonates at a Ca /Mg ratio typical of natural waters and
then aerated. Furthermore, since recent work by Benoit and Mattson (83) have
indicated that chloride and sulfate salts may exert somewhat different effects
on toxicities, the non-carbonate hardness should be added in ratios of
- -2 - -2
CI /SO, salts similar to the CI /SO, ratios typically seen in natural
waters. Because the effects of Na and K are not well known, the use of Na
and K+ salts should be avoided. The use of such salts is actually not
necessary if hardness is divided into carbonate and non-carbonate hardness and
the effects of alkalinity is considered jointly with hardness in the form of
the carbonate hardness. Typical combinations of carbonate hardness,
non-carbonate hardness, and the pH can be determined from national surveys of
freshwaters such as the U.S. Geological Survey entitled Quality of Rivers of
+2 +2 — -2
the United States, 1975 Water Year. Typical ratios of Ca /Mg and CI /SO^
can also be determined from such sources.
4-10
-------
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4-13
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52.
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Tahata - Needs reference.
R. Andrew, of U.S. EPA Environmental Research Laboratorry at Duluth.
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Research, Inc. Cambridge, MA. (1979).
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81. D. Benoit, V. Mattson, and D. Olson. "A Continuous Flow Mini-Diluter
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+ 2 +2
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TECHNICAL REPORT DATA
(Please read Instructions on the reverse before completing)
1. REPORT NO.
EPA/600/3-86/023
2.
9
3. RECIPIENT S ACCESSION NO.
4. TITLE AND SUBTITLE
The Effects of Variable Hardness, pH, Alkalinity, Sus-
5 REPORT DATE
March 1986
pended Clay, and Humics on
Aquatic Toxicity of Copper.
the Chemical Speciation and
6. PERFORMING ORGANIZATION COOE
7. AUTHOR(S)
Henry Nelson*, Duane Benoit
Mattson, Jim Lindberg.
, Russell Erickson, Vince
8. PERFORMING ORGANIZATION REPORT NO.
9 PERFORMING ORGANIZATION NAME AND ADDRESS
* Scinece Applications Int'l.
10. PROGRAM ELEMENT NO.
11. CONTRACT/GRANT NO.
12. SPONSORING AGENCY NAME ANO ADDRESS
U.S. Environmental Protection Agency
13. TYPE OF REPORT AND PERIOD COVERED
Environmental Research Laboratory-Duluth
6201 Congdon Boulevard
Duluth, MS1 55804
14. SPONSORING AGENCY CODE
EPA/600/03
IS. SUPPLEMENTARY NOTES
16 ABSTRACT
The effects of variable hardness, pH, alkalinity, humics, and suspended
clay on the chemical speciation of copper and its toxicity to fathead
minnow larvae in Lake Superior water were investigated. Two proposed methods
(toxicity factors and chemical speciation) for predicting LC50 values in
specific natural waters from laboratory toxicity data and the average
site-specific values of general water quality parameters were evaluated.
The accuracy of the cupric ion selective electrode in determining CU+2
activities in ambient and chemically altered Hake Superior water was also
determined.
17.
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Release to public.
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Unclassified
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132
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