and Cadmium Binding to Fish Gi : Estimates of ing of Meta

Richard C. Playle' and D. George Dixon Department of BioBogy, University sf LWaterBoo, Waterboo, ON N2% .?GI, Canada

and Kent Burnissn Nationai Water Research hsfifute, Environnsent Canada, Burl-lingfon, O N L7R 4A6, Canada

Playle, W.C., D.G. Bixon, and K. Burnison. 1993. Copper and cadmium binding to fish gills: estimates of metal-gill stability constants and model ling of metal accumulation. Can. J. Fish. Aquat. Sci. 58: 2678-2687.

Fathead minnows (Pimephales promebs) were exposed to 17 p g Cu %-! or 6 p g Cd - L-I in synthetic soft water in the presence of competing ligands. Measured gill metal concentrations correlated with free metal ion concentrations, not with total metal. Langrnuir isotherms were used to calculate conditional metal-gill equilibrium constants and the number of binding sites for each metal. Log K C ~ - ~ , I I was estimated to be 7.4 and the number of Cu binding sites on a set of gills (70 mg, wet weight) was -2 X 1 0-8 mol (-30 nmol . g wet weight-'). Log Kcd-gll~ was -8.6, and the number of Cd binding sites on minnow gills was -2 x 1 0-lo n-eol (-2 nmol . g wet weight-'). Stability constants for H+ and Ca interactions at metal-gill binding sites and for metat interactions with dissolved organic carbon (DOC) were estimated using these metal-gill constants. All stability constants were entered into the MlNEQL+ aquatic chemistry program, to predict metal accumulation on fish gills using metal, DOC, and Ca concentrations, and water pH. Calculated metal accumulation on gills correlated well with measured gill metal concentrations and with LC50 values. Our approach sf inserting biological data into an aquatic chemistry program is useful for modelling and predicting metal accumulation on gills and therefore toxicity to fish.

Des TGtes-de-boule (Pimephales promedas) ont 6t6 expos6s 2 du Cu ou du Cd en concentration de 1 7 pg - L-I ou 6 p g - L-l, respectivement, dans une eau douce synthktique, en pr6sence de ligands concurrents. Les concentrations mesur6es de metal sur les branchies etaient en corr6lation avec celles des ions metalliques tibres, mais pas avec celles du metal total. Bes isotherrnes de Langmuir ont servi 2 calcuier Ces constantes conditionnelles d'kquilibre metal-branchies ainsi que le nombre de sites de fixation dans le cas de chaque mktal. be log Kcu-branch,e a kt6 6valu6 5 7,4 et le nombre de sites de fixation du Cu sur une paire de branchies (76) mg poids frais) 6tait d'environ 2 X 1 mol (-30 nmos - g poids frais-I). Le log d(Cd-bra,,chlea kt6 6valu6 5 environ 8,6, et Je nombre de sites de fixation de Cd sur les branchies 6tait d'environ 2 x 1 0-lo mol (-2 nmol - g poids frais- I ) . Les constantes de stabiIit6 des interactions H+ et Ca sur les sites de fixation des m6taux avec le carbone organiyue dissous (COD,, ont 6t6 kvafu6es 2 partir de ces constantes m6taux-branchies. Toutes les constantes de stabilit6 ont 6t6 versees dans le programme de chimie de l'eau MINEOL' dont on s'est servi pour prevoir I'accumulation des metaux sur les branchies 2 partir des concentrations des rn6taux' du COB et du Ca ainsi qu'h partir du pH de l'eau. L'accurnulation calcul6e de m6taux sur ies branchies etait en bonne corr6lation avec la concentration de m6taux mesur6e sur les branchies ainsi qu'avec Bes valeurs de CLso. Nstre methode d'insertion de donrakes biologiques dans un programe de chimie de t'eau constitue une faqon utile de mod6liser et de pr6voir Ifaccumulation des m6taux sur les branchies, donc de la toxicit6 de ces m6taux pour les psissons.

Received September 1 1992 Accepted May 18/ 1 993 (JB635)

t has long been realized that metal-organism stability constants would be useful in interpreting and predicting the toxicity of waterborne metals to aquatic organisms

(Biesinger and Christensen 1972; Zitko and Carson 1976). To date, conditional stability constants for Cu have been determined for algae (Xue and Sigg 1990) and for eoamplexing agents released by algae (Van Den Berg et al. 1979; Xue and Sigg 1990) and zooplankton (Fish and Morel 11983). These Cu stability

R e p je 18 septembre 1992 Accept6 Be 18 mai 1993

constants fall in the range 1 0 ~ - 1 0 ~ ~ , indicating that 50% of the Cu binding sites would be occupied at aqueous Cu concen- trations o f I O - ~ - I O - ~ ' M.

For fish, Reid and McDonald (1991) detemined metal-gill stability constants for La, Cu, Cd, and Ca, using excised gills of rainbow trout. Their values, loz4-1 03.', were determined using radiolabelled megals in metal solutions. These values are low compared to those for algae, most likely because of the relatively high metal concentrations used. The conditional nature

'Author to whom correspondence should be addressed. Current : Department of ~ i ~ l ~ ~ ~ , Wilfrid Laurirr University of stability constants I ~ G ~ s them dependent on the metal Con-

75 University Ave. West, Waterloo. ON N2E 3C5, Canada. centrations used du~ing their determination: high metal

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concentrations yield low stability constants because even low affinity sites are able to bind excess metals. At low metal concentrations, only high affinity sites are able to bind the limited mount of metal, yielding higher stability constants.

Strength of metal binding is not the sole determinant sf metal interactions with aquatic organisms. Dissolved organic carbon (DOC), Ca concentration, and pH are important modifiers of metal bioaccumulation and toxicity in freshwater organisms (Yan et al. 1990; Meador 1991 ; Playle et al. 1992, 1993). Dis- solved organic carbon can complex metals, rendering them unavailable to interact with an organism, and calcium competes with metals for binding sites, as does H+. In addition, water pH determines the proportion of total, unbound metal that is in the cationic form, generally thought to be the toxic species sf metals (Morel 1983; Pagenlcopf 1983).

The objective of the present study was to determine metal-gill stability constants for low concentrations of Cu and Cd, and to use these constants, plus metal-DOC, Ca-gill, and H-gill constants, in a computer model to predict metal accumulation on fish gills. This approach has been used for algae (Xue and Sigg 1990) but has not yet been applied to fish. Metal accumulation on minnow gills in the presence of complexing ligands such as EDTA were detemined in our preceding paper (Playle et al. 1993). We used the lowest concentrations of Cu and Cd (17 and 6 wg - L-l, respectively) that were feasible to yield measurable metal deposition on fish gills, using graphite furnace atomic absorption spectroscopy. The gill accumulation data plus cal- culated free metd concentrations in the water were used to determine the metd-gill stability constants. Free metal con- centrations were calculated using MINEQL' (Schecher 1991), an aquatic chemistry computer program based on equilibrium constants. Finally, our estimates of the metd-gill stability con- stants were assessed by comparing calculated metal accumu- lation on gills with measured gill metal concentrations of mimows exposed to Cu and Cd in water fmm a series of Ontario lakes.

Materials and Methods

Ligand Experiments

Detailed descriptions of the ligand experiments are given in the preceding paper (Playle et al. 1993) and the results are summarized in Tables I and 2 of the present paper. In brief, acclimated adult fathead minnows (Pirnepkeakes p~-omelas) were exposed to 17 yg Cu Lpl (8.27 pM) or 6 kg Cd- LA' (0.05 pM) in synthetic soft water (Na, Ca -50 yM) at pH 6.2, 19°C. Fish were killed at the end of the 2-3-h exposures, their whole gill baskets were removed, weighed, then digested in 1 N H,SO, for 8 h at 80°C. The supernatant of the gill digests was diluted with deionized water, then analyzed for Cu and Cd using graphite furnace atomic absorption spectroscopy. Experiments were run in the presence or absence sf freeze-dried dissolved organic carbon (DOC, 0-8 mgo L-I) or the synthetic ligands EDTA, nitrilotmiacetic acid (NTA), ethylenediamine (EN), and citric, glutamic, oxdic, and salicylic acids (0-2580 pM).

Calculation of Gill Stability Constants

We used Eangmuir isotherms to calculate Cu- and Cd-gill stability constants (Kc,_,iu, KCd-gill). In order to make these calcu- lations, estimates sf free Cu2+ and Cd2+ concentrations in the synthetic soft water, with and without added ligands, were made

using the MINEBE+ program. Input values were 0.27 FM CU, 0.05 pM Cd, 70 p,M Na, 35 pM Ca, fixed pH 6.2, 19"C, and concentrations of the added ligands (0-2500 pM). Solution ionic strength was calculated by MHNEQL' (H- 1 X lW4), and metd-ligand log K values were supplied by the computer program (Table 1, 2). We ignored changes in water pH near the gills (Playle and Wood 1989; Lin and Randall 1990) because both Cu and Cd exist in the cation f o m at pH < 7; small pH changes near pH 4 were therefore inconsequential with respect to speciation of these two metals.

Because we had information on Ca and H" interactions at Cu and Cd gill-binding sites (Playle et al. 11992, 1993), it was possible to estimate Kc ,-,,,, and K,-,,,, once and Kc,,,,, were determined. These estimates were made in the following manner.

Metal-ligand equilibrium constants are defined as the concentration of the metal-ligand complex divided by the product of free metal and free ligand, once the reaction is complete. For ewanple, Cd and gill interactions can be defined as

[Cd-gill] (1) K ~ d - ~ i l l =

[Cd] . [gill]

Calcium interactions at the same sites on the gill can be defined by

[Ca-gill] (2) Kca-gil~ =

[Ca] [gill]

Remanging Equation 1, and substituting for [gill] in Equation 2, we get

[Ca-gill] [Cd] (3) K~a-~ill = [Ca] . [Cd-gill]

This equation can be used to calculate Kca,ill if we h o w either how many gill sites, or the proportion of sites, that are occupied by Ca and Cd. If half the gill sites are occupied by Ca, and half by Cd, then Equation 3 simplifies to

In the situation where the same mount of Cd accumulates on the gills in the presence of Ca as in its absence (i.e., the mount of Ca in solution does not significantly out-compete Cd for the gill binding sites) then

If the Crt in solution prevents all Cd from accumulating on the gills, then

Hydrogen ion interactions at rnetd binding sites can be cal- culated in the same manner, by substituting H+ for Ca. Copper

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TABLE 1 . Gill Cu concentrations (above background) of fahead minnows held for 2-3 h in soft water containhg 0.27 FM total Cu, plus ligmds. S u m m ~ z e d from Playle et al. (1992, 1993). EN = ethylenedimine.

Ligmd Free cu2' Gilt Cu log K Cu-ligmd (PW (pM; MINEQL' calculation) (yrnol . g- wet tissue) (from MHNEQL')

EBTA 0.15 0.25 0.30

NTA 0.25 2.5

EN 0.25 2.5

25 Citrate 0.25

2.5 25

Bxdate 0.25 2.5

25 Glutmate 0.25

2.5 25

Sdicylate 0.25 2.5

25 100 250

No ligands No ligmds,

no added Cu

TABLE 2. Gill Cd concentrations (above background) of fathead minnows held for 2-3 h in soft water c~ntaining 0.05 FM total Cd, plus ligmds. S u m ~ z e d from Playle et al. (1993). EN = ethylenedimine.

Ligmd Free cd2' Gili Cd l ~ g K Cd-Iigmd (PM) (a; MINEQL' calculation) (ram01 - g- wet tissue) (from M~~INEQL')

EDTA 0.05 0.20 0.25 0.5 5.0

No ligmds

No ligmds, no added Cd

can be substituted for Cd. Implicit in these calculations is the Lakewater Exposures and Toxicity Tests assum~tisn that there is ensueh metal and time for the nil1 sites

V b/

to becme saturated. Tn practice, the metal coneen$rations and All stability constants were entered into the MImQLc aquatic exposure times used in our expe6men$s adequately fulfilled chemistry program, to cdculate metal accumulation on fish gills. these assumptions (Playle et al. 1992, 1993). To test the model, we wanted to measure metal accumulation on

2680 Can. J. Fish. Aquaf. Sci., Vol. 50, 1995'

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0.1 0.2

free Cu+" (pMB

RG. 1. Plot of measured gill Cu against free 6u2' concentrations. Gill Cu concentrations were from fithead minnows exposed 2-3 h in synthetic soft water with 17 p.,g Cu . k-' (0.27 pM) and 0-250 pM of synthetic ligmds (data fmm PlayIe et al. 1992, 1993). Free eu2' was calculated using MINEQL+ and known concentrations of ligands and ions (Table 1). Gill Cu varied directly with free Cu2+ ( T = 0.858, P < 0.001). The equation of the line was y = 6.23~ -t- 0.86 ( y in pg Cu a g tissue-') 01- y = 0.098~ + 0.014 ( y in pmol . g tissue ').

gills when minnows were exposed to 17 gag Cu L-' and 6 gag Cd . L-I in natural soft water, as opposed to synthetic soft water. In addition, we wanted to conelate metal accumulation on gills with metal toxicity.

Wdter was collectzd from five lakes NE of Peterborsugh, Ontario, in June, 1991. The lakes were: Salmon L., Brooks Bay and Sharps Bay of Jack L., Anstruther L., Loaacks k., and Wolf L. (dl approximately N44"42', W78" 13 '). Twenty-L stainless steel containers (Spzu-tenburg Steel Products, Spartenburg, SC) were used to collect the surface water. These containers were then pressurized with N2 gas and the water was filtered through I .0 gam glass fibre filters ( M E , 142 i~m, Gelman Sciences Inc., Ann Arbor, MI) to remove algae. The filtered water was stored in 40-L stainless steel containers (Simgo hc., Toronto, ON), brought back to Waterloo, transferred to 4-L polyethylene con- tainers, and stored at 4OC. These water samples were used for combined Cu and Cd metal exposures to adult fathead minnows, conducted in the same manner as described for the synthetic softwater experiments. A synthetic soft water plus a 10-pM (total) mixture of glutamic, citric, and oxalic acids was also used (3.3 ~ L M of each), to mimic a 5 mg DOC - L-I solution (Campbell and Stokes 198%).

Toxicity tests using a gesmetiic series of concentrations of the Cu and Cd mixture were run in the lake waters, using juvenile fathead minnows. These fish were 91 rno old, and had been held in soft water for 1 wk before the experiment. During acclimation they were fed live brine shrimp and fine powdered fish food twice daily. Ninety-six h ILC50s were determined using duplicate exposures of five fish per 188-mlk solution in un-aerated poly- ethylene urine cups. Eighteen juveniles (blotted dry) weighed individually yielded an average weight of 1.6 & 0.5 mg. LC50 values were calculated using a trimmed Spemm-Kaber analysis.

Calculation of Metd-Gill Stability Constants

Copper Previously, we measured Cu accumulation on gills of fathead

minnows exposed to 17 y~g Cu a IL-' (8.27 FM) in synthetic soft water, in the presence or absence of synthetic ligands (Playle et al. 1992,1993). Because we h e w the water chemistry and the

0 0 0 20 40 60

free Cd++ (nM)

FIG. 2. Plot of measured gill Cd against free Cd:+ concentrations. Gill Cd concentrations were frorap fathead minnows exposed 2-3 h in synthetic soft water with 6 pg Cd L-' (0.05 pM) and 0-2500 pM of Iigands (data from Playle et al. 1993). Free Cd2' was calculated rasing MINEQLf and h a w n concentrations sf Bigands and ions (Table 2). Gill Cd varied directly with free Cd2' ( r = 0.732, P < 0.01 ). The equation of the line was y = 0.004~ + 0.163 (y in kg Cd . g tissue-') or y = 0 . 0 3 1 ~ + 1.45 1 ( y in nand - g tissue-').

ligand concentrations, we were able to calculate free Cu concen- trations using the MHNEQL+ program. Gill Cu eoncentmtions varied directly with free Cu2+, the major component of free Cu: gill Cu versus free Cu2+ gave a straight line (Fig. 1; r = 8.858, P < 0.001). Gill Cu concentrations did not correlate with total Cu, which was constant at 0.27 pM.

To calculate log Kc,-,i, and the nwrlber of Cu binding sites om the gills, a Laragmaair plot of Cu adsorption was constructed. Free Cu2+ (micromolm) divided by gill CU (micromsles Cu per gram wet tissue, minus background gill Cu, Table 1) was plotted against free Cu2' (micromolar). The equation for the line was y = 3 5 . 3 8 ~ + 1.28. The inverse of the slope of the line is the number of gill binding sites, = 8.03 ~ m s l g wet tissue-'. The inverse of the intercept = k'. (binding site number), therefore K= 26 L . yrnol-I = 26 X 106 E . mol-I, and log Kc,,ill = 7.4. Average total gill weight for the mimows was 0.07 g (wet tissue; Playle et al. 1993), tl~erefore total binding sites for Cu on the gills of our fathead minnows was about 2X 1W9 mol.

Cadmium. Accumulation of Cd on gills of fathead minnows exposed to

6 pg Cd . L-I (0.05 pM) in the presence or absence of synthetic ligands (from Playle et al. 1993) was plotted against free Cd2+ concentrations calculated using MINEQL* (Fig. 2). Although there were not as many data points as there were for Cu, gill Cd still varied directly with free Cd2+ (r = 0.732, P 9 0.01). It was difficult to obtain intermediate values for the curve, because so little Cd was used in the experiments (e.g., 0-50 nM is just one-fifth the scale for Cu in Fig. 1).

The log K,,,,,, and the number of binding sites for Cd were cdculated using a Langmuir plot. Gill Cd concentxations above background (Table 2) were used for these calculations. The Langmuir isotherm, calculated in nmoles, yielded the line y = 0 . 4 4 ~ + 1 .O&. The number of gill binding sites was the inverse of the slope, = 2.27 nmol g wet tissue-'. From the inverse of the intercept, K = 0.42 % nmol-' = 8.42 X 10% L. mol- ', and log dkl,d.,il, = 8.6. Average gill weight was 0.07 g wet tissue; thus the number of gill binding sites for Cd was about 2 X 10-'%08 per minnow.

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TABLE 3. Input data for the metal-gill interaction modd. Initial log K values were as calculated in the text. Final log K values were determined using MINEQL' to best fit, within the constraints of the initial log K values, calculated gill metal with measured gill metal ccsncentratisns for the synthetic soft water system.

Binding sites Complex Initial log K Final log K

Cu-gill 2 X lo-' mol per fish

Cd-gill 2 x I O- '' rnol per fish

Calculation of Ca, H', and DOC Stability Constants

C d c i m and H+ can compete with Cu and Cd for binding sites on gills. By using the log Kme,,,-gl,, values calculated above, and data presented in Playle et A. (1 992,1993), the stability constants for Ca and H+ were estimated.

Copper accumulation on minnow gills was not affected at pH 4.8 (15.8 gaM H+) compared to pH 6.3 (0.5 pM H'; Playle et al. 1992). Inserting Kc,,,,, = loT4, free Cu = 2.53 X lop7 M (MINEQL+ calculation), and H+ = 15-8 X los6 M into Equation 5 yields log KH-gl,, < 5.6. Copper accumulation was not affected by 2 000 p M Ca (Playle et d. 1992): using Equation 5, Cu = 2.53 X lW7 M, Ca - 2 X 10-%, and Kc ,-,,,, = lo7." yields log Kc,-?, < 3.5.

Cadmum accumulation on the gills was reduced at pH 4.8 (Playle et d. 1993). Inserting Kc,.,,,, = BOH.6, free Cd = 4.93 X 1W8 M (MINEQL' calculation), and H' = 15.8 X 1OP6 M into Equation 6 yields log KH,,, > 6.1. For Ca, 95 pM did not reduce Cd accumulation on gills, whereas 1 850 pM Ca did (Playle et al. 1993:s. Using Equations 5 and 6, respectively, log Kc,_,lll < 5.3, but >4.3, the mean of which is log Kc,.,,il = 5 .O.

Five to 6 mg DOC L-I prevented Cu accumulation on gills, by complexing all the 0.27 pM Cu in solution (Playle et al. 1993). Thus, there were at least 0.05 p-nol binding sites per mg DOC - L-l. These metal binding sites must have had a higher affinity for Cu than did the Cu binding sites on the gill, or the DOC would not have been able to prevent Cu accumulation on the gills. The Bog Kc,-= value for the 0.85 pmol of high affinity binding sites per mg - Lml DOC must therefore be about 1-2 log units higher than the log Kc,-,lI1 = 7.4 value. Our initial estimate of the stability constant was log Kc,-Doc = 8.4-9.4, which is in the mid-range of published values (Van Den Berg and Krmer 1979; Morel 1983). Note ha t direct attack of the Cu-gill complex by DOC is unlikely (adjunctive pathway of ligand exchange; see Hering and Morel 1990). More likely is the disjunctive pathway, that first involves dissociation of Cu from the Cu-gill complex, the free Cu then being able to react with DOC in solution.

Cadmium has consistently been reported to bind <IOX as well to DOC than does Cu (Alberts and Giesy 1983; Tuschall and Brezonik 1983). Thus, an initial estimate of the log Kc,,,, value of 10-100X Bess than for Cu would be 6.4-8.4. The log R for humic substances is 3 to 4 (DeWit et al. 19911, therefore we set log KH_,,, = 4. We had no idomation on the interactions sf Ca with DOC.

Model Development

Our purpose in determining metd-gill and other stability constants was to use these values in a computer model to predict metal accumulation on fish gills, and thereby predict metal toxicity. In essence, bringing biology into an aquatic chemistry program. The model we chose to use was MINEQLf (Schecher 1991), an aquatic chemistry program based on log K stability constants.

Our initial estimates of log K stability constants md the number of metal binding sites are s u m m ~ z e d in Table 3. To input these data into the MINEQL' program, three null com- ponents were defined (gillcu, gillcd, and DOC). We simulated two different sites for Cu and Cd binding on the gills (gillcu, gillcd) on the basis of our previous results (Playle et d. 1993) that indcated, for the low metd concentrations we used, no competition between Cu and Cd for gill binding sites. Metal, Ca, and H+ interactions with gillcu, gillcd, and DOC were defined in the program as 1 : 1 reactions (Table 3).

We used Na = 70 pM, Ca = 35 pM, Cd = 0.05 pMt Cu = 0.27 FM, fixed pH 6.2, and Z 9°C for water chemistry parmeters of the synthetic soft water. For log K values for which we just had upper or lower values, or a range, we chose final log Kvalues after optimizing to best fit the observed data in Playle et al. (1992, 1993). In essence, the model was able to mimic the greater effect of H' and Ca on Cd binding to gills, and the greater effect of DOC on Cu binding to gills. Of note in Table 3 are the final log Kmetal-BOC values, that we optimized to ]log Kc,-,, = 9.1 (in the model, Cu binds to DOC 50X better than it binds t s the gill), and log = 7.4 (in the model, Cd binds to DOC 50X less well than does Cu, and Cd binds to DOC I6 X less well than Cd binds to gills). We chose B mg . L-I DOC = 0.05 ~ m o l e binding sites, to allow enough binding sites for all the Cu and Cd in solution at 6 mg DOC L-* (e.g., no competition by the metals for the sites). In general, as long as a ligand is in excess relative to metals in solution, metals do not compete with each other (Morel 1 983).

Metal Accumulation and Toxicity to Fish Exposed to Cu and Cd in Soft Lake Water

Of the natural lake waters that we used in the Cu and Cd exposures, Salmon L., and Brooks and Sharps bays contained more Ca, more DOC, and had higher pH than did Anstruther, Loucks, and Wolf lakes (Table 4). The synthetic soft water plus a 18 FM (total) addition of glutmic, citric, and oxalic acids had water chemistry similar to that of synthetic soft water without

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TABLE 4. Lake water md synthetic soft water characteristics, and juvenile hthead minnow 96 h LC50s. Water characteristics were measured on water used for the metal deposition and LC58 experiments, except for conductivity, which was from a 1988 survey. SSW = synthetic soft water; + ligands = plus 10 pM (total) glutamic, citric, and oxalic acids. DOC = mg . L-', Ca and Na = FM, conductivity (Cond.) = pmho cm-l, and all metals = yg L-'. Mean 5 SEM (n) if n > 2. For LC50, values in parentheses = 95% CI.

96 h LC50

Water PH DOC Ca Na Cond. Cu Cd &I He Zn Cu Cd

Salmon Lake 8.05 k0.01 (12)

Brooks Bay 7.80k0.01 (12)

S h q s Bay 7.9050.02 (12)

AnstrutherLake 7.10iz0.83 (12)

Loucks Lake 7.04kO.O 1 (12)

Wolf Lake 6.91 k0.03 (12)

SSW + ligmds 6.60k0.03 (8)

SSW 6.49k0.01 (20)

added ligands. Minnow exposures in both these synthetic soft 0.8 waters were at pH -6.7, to match the pH of the softer sf the natural waters. h

Copper and Cd 96-h LC50 values for juvenile fathead minnows exposed in the various waters are given in Table 4. In general, LC50s for the Cu and Cd mixtures were lowest (metals were most toxic) in softer water* Metals were most toxic in the synthetic soft water, and were only slightly less toxic in the presence of the 10 pha ligand mixture. Metals were Heast toxic in water from Salmon L., with too few mortalities to accurately determine the LC50 (Table 4).

Cadmium accumulation on gills of adult minnows exposed to 6 pg Cd Lpl and 17 pg Car L-' was highest in Wolf L. water (significantly above background, and not significantly different from metal-exposed fish in synthetic soft water done; Fig. 3), and lowest in Salmon L. water (not significantly above back- ground). Fish exposed to the lake waters without added metals did not have gill Cd concentrations significantly above back- ground (0.18 ? 0.02 pg Cd - g wet tissue-'; +95% CCI, n = 17). Fish exposed to CQ and Cu in synthetic soft water plus 10 kEa/f (total) glutamic, citric, and oxdic acids had 0.64 2 0.23 (6) kg Cd . g wet tissue-l, and fish exposed to the metals in synthetic soft water without added ligands had gill Cd concentrations of 0.54 * 0.05 (9) kg Cd - g. These exposures were run at pH 6.6 and 6.7, respectively; their gill Cd concentrations were no? significantly different from those of fish exposed in synthetic soft water at pH 6.2.

Copper accumulation on gills of adult minnows was never above background in fish held in lake waters supplemented with Cu and Cd (Fig. 4), probably because DOC exceeded 5 mg . L-I in all lake waters (see Playle et al. 1993). In lake water without added metals, gill Cu was 0.80 +- 0.08 (1 71, also not significantly above background. Minnows exposed to Cu and Cd in synthetic soft water plus the 10 pM ligand mixture (pH 6.6) had gill Cu concentrations of 2.58 + 0.18 (6) pg Cd . g wet tissue-', while fish exposed to the metals in synthetic soft water without added bgmds (pH 6.7) had gill Cu concentrations of 2.16 + 0.30 (9) ~g Cu g-'. Neither concentration was significantly different from

Lake FIG. 3. Gill Cd concentrations of fathead minnows exposed 2-3 h in Cd- and Cu-supplemented lake water. Cadmium deposition on gills was highest in the softest water. S = Salmon L., B = Brooks Bay, SH = S h q s Bay, A= Anstruther L,, L = Loucb L., and W = Wolf L. Vertical lines on bars = 95% CI; n = 6 for each bar. Dashed horizontal lines represent the 95% CI sf gill Cd for minnows held in synthetic soft water in the absence of added metals (n = 38). The grey horizontal band represents the 95% CI of gill Cd for fish held 2-3 h in synthetic soft water with 6 pg Cd . L-' ( i ~ = 45). *, **, *** = significantly below metal-exposed gills (grey band), and +, ++, +++ = significantly above background gill metal concentration (dashed lines) for B e 0.05, 4 .01 , 90.001, respectively.

gills of minnows exposed to Cu in synthetic soft water at pH 6.2.

Model Testing

Final values for the log K stability constants (Table 3) were used to assess the usefulness sf the model in predicting gill metal concentrations and metal toxicity. In the model we used the

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Lake

FIG. 4. Gill CU concentrations of fathead minnows exposed 2-3 h in Cd- and Cu-supplemented lake water. There was no significant zcumu- lation of Cu on the gills, likely because DOC was 2 6 mg - L-' in each lake water. Dashed horizontal lines represent the 95% CI of gill Cu for minnows held in synthetic soft water alone (n = 38)- The grey horizontal band represents the 95% CI of gill Cu for fish held 2-3 h in synthetic soft water with 1'7 k g Cu - L-' (n = 4%). Lake names and other details are given in caption of Fig. 3.

FIG. 5. Measured and modelled gill Cd, for fathead minnows exposed 2-3 h to 6 k g Cd - L-I and 1'7 pg Cu - L-' in natural soft waters and synthetic soft water (SSW) with or without 10 FM added ligmds (at pH 6.7). Lakes are defined in the caption of Fig. 3. Gill Cd is expressed as a percentage of the mean gill Cd concentration (1 80%) for fish held in synthetic soft water plus Cd at pH 6.2 (grey bar, kg§% Ch). Background gill Cd = 0% (+95% CH, dashed line). Calculated gill Cd cornelated well with measured gill Cd (see text for details). Measured gill Cd values are given with their 95% CCH.

DOC, Ca, and Na values (molar), and pH (fixed) for the eight waters listed in Table 4. The model mimicked a system open to atmospheric CO,. Although we did not measure C032- in the lake waters, for the model we assumed its concentration was equal to that of Ca (Table 41, which yields approximately the same values as if the more rigorous Henderson-Hasselbdch equation were used to calculate HC8,-. For the comparison of measured and calculated gill Cd and Cu, measured values from Fig. 3 and 4 were converted to percentages, with background gill metal = 0%, and gill metal concentrations of fish exposed to metals in synthetic soft water (at pH 6.2) -- 100%. The calculated gill metal concentration in synthetic soft water (pH 6.2, assuming back- ground DOC = 1 rng . L-') was used as the model 100% value.

S B SH A h W SSW SSW Q

Lake Bigends

FIG. 6. Measured and modelled gill Cu for fathead aninnows exposed 2-3 h to 17 kg Cm L-' and 6 gcg Cd k-' in natural soft waters and in synthetic soft water (SSW) with or without added ligands. Lakes are defined in the caption of Fig. 3. Gill Cu is expressed as a percentage of the mean gill Cu concentration (1064%) for fish held in synthetic soft water plus Cm (grey bas). Background gill Cu = 0% (dashed line). Calculated and measured gill Cu correlated well (see text for details).

Modelled gill Cd agreed well with our measured values (Fig. 51, with the Bargest discrepancy for S h q s Bay. The eor- relation coefficient between measured and modelled gill Cd was 0.9008 (P < 0.011). Measured gill Cd versus the juvenile fathead minnow LC%& for Cd (Table 4) yielded a correlation coefficient of -0.897 (P < 0.01). Modelled gill Cd versus the LC50 values yielded a better correlation (r = -0.960; P < 0.001). These data indicate that model-calculated gill Cd concentrations could rea- sonably be used to predict acute Cd toxicity to fish.

Measured and modelled gill Cu also agreed well (Fig. 6; r = 0.926, P < 0.001), in spite of the fact that none of the gill Cu concentrations from the natural lakewaters was above back- ground (Fig. 4). Measured gill Cu versus the Bog of the juvenile fathead minnow LC%Os for Cu yielded r = -0.830 (P < 0.0%; gill Cu vs. linear LCSOs was not significant, r = -0.662, P r 0.05)- Modelled gill Cu had a slightly better correlation with the log LCSOs (a. -- -0.895, P < 0.01; the correlation with linear LCSOs was r = -0.727, P < 0.05). As was the case with Cd, modelled gill Cu concentration was a good indicator of acute Cu toxicity.

In the present study, we have better defined interactions of Cu and Cd with fish gills, as influenced by dissolved organic cabon (BOG), pH, and @a. We used Langmuir isotherms to estimate conditional metal-gill stability constants. These were log Kc ,-,,,, = 7.4 and log Kc ,-,,,, = 8.6. There were a b u t 10 X more of the weaker Cu sites on the gills (-2 X 1W9 mol per fish) than Cd sites (-2 X 10-1° mol per fish).

Intemtisns of Cu and Cd with the gill and with DOC, and competing reactions with Ca and H+, are illustrated conceptually in Fig. 7, along with their log K values. Cadmium binds to the gills better than does Cu, and is more affected by competition from Ca and M+. These irnteractisns are likely a consequence of active Cd uptake through high affinity Ca channels in addition to general surface binding. Cadmium reduces Ca uptake, mainly though effects on basolateral @aZ*-ATPase activity (Verbost et al. 1989). The data of Reid and McDonald (1988) support the existence of higher affinity binding sites for Cd compared to Gu. Calcium competition reduces Cd uptake at gills, and therefore

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the subsequent distribution to the rest of the fish body (WicMund and Wunn 1988). Transfer of metals through the basolateral membrane is probably by diffusion (Vesbost et al. 1989).

Copper, on the other hand, appears to bind to weaker sites on the gills, and possibly also enters through ion channels (Fig. '$1, but binds more strongly to DOC than does Cd (log 2YCu~,,, = 9.1 ; log Kcd.Doc = 7.4). The binding order for DOC reflects relatively stmng binding of Cu to Iigands. The effect of low concentrations of CU at the gills is an irhibition of Na' influx, mainly through effects on Na+-K+-ATPase (Lauren and McDonald 1 987). Relatively more inhibition of this effect was seen by carbonate complexation than by Ca (Lauren and McDonald 1986), which agrees with our results of less protective effects of Ca on Cu accumulation, compared to effects on gill Cd (Playle et al. 1992, 1993). Note that in the conceptual model illustrated in Fig. 7 the gill mucus layer is ignored. Gill mucus probably binds metals and slows metal access to the gill (Ptirt and Lock 1983). However, once a metal exceeds the complexing and sloughing ability of mucus, the metal will become available to react at the gill epihelium; this is the situation illustrated in Fig. 7.

There are few log Km,,-,,,, values with which to compare our values. Reid and McDonald (1991) found log K values of about 3.5,3.0,3.0, and 2.4 for La, Cd, Ca, and Cu, respectively. These values are low, probably because they used rraM concentrations of metals in heir experiments instead of environmental, sub-rraicromolx concentrations. Log Kvalues are sensitive to the metal concentration at which the experiments are run, a d it is difficult to extrapolate from a stability constant obtained at high meta1:ligand ratio to a stability constant for a low meta1:ligand ratio (Perdue and Lytle 1983). However, Reid and McDonald (1991) found that Cd bound to gills better than did Cu, which agrees with our results. Xue and Sigg (19964) determined log = 8.8 (at pH 6.0), which is -I0X higher than ours for fish at about the same pH, These workers were successful in model- ling Cu binding to algae using a computer program based on log R values, their input log Kc,.,,,, values, and the number of binding sites (2 pmol . g dry algae-') they determined for algae.

We inserted our metal-gill, H+-gill, Ca-gill, and metal-DOC stability constants (Fig. 7; Table 3) into MHNEQL+ (Schecher 199 1). The model adequately reflected measured metal accumu- lation on the gills (Fig. 5,6). In addition to Cu and Cd bound to gills, the MINEQL+ output included Ca and H+ bound to gill sites (an indication of competition for the sites), unoccupied gill sites. free metd in solution (an indication of the pool of metal available to bind at the sites), and metal bound to DOC (representing the extent of metal complexation). This simple model considered only 1: I metal-ligand interactions, and ignored Ca interactions at metal binding sites on DOC. At the low metal concentrations we used, 1 : I binding would be expected: Cabaniss and Shuman (1988a) showed that Cu-DOC interactions were >90% B:l binding. Not incorporating Ca-DOC interactions probably would have had little effect on the model, because both Cu and Cd would bind much more strongly to DOC than would Ca (Kemdorff and Schnitzer 1980; Cabmiss and Shuman 1988a; Daly et al. 1990). It should be noted that our stability constants represent average conditional stability constants, akin to con- tinuous multiligand binding models (e.g., Perdue and Lytle 1983; G r i m et al. 1991) or '6qqeaasiparticle" models (Sposito 1981).

Stability constants change with solution chemistry (pH, ionic strength) as well as with metd and Bigand concentration (Perdue and Lytle 1983; Cabaniss and Shuman 1988a). Van Den Berg and Krmer (1979) and Morel (1983) determined that there is

DOC

2"'

Cu - 7.4

# - 5.4

Car - 3.4

Cd - 8.6

H- 6.7

Car - 5.0

CU+ +

FIG. 7. Conceptual iillustmtion of Cu, Cd, Ca, and H' interactions at fish gills and with dissolved organic carbon (DOC). The numbers are the log K conditional stability constants for the interactions. In water (left), Cu and @d may become complexed by DOC, and also face competition from Ca and H+ for negative binding sites on gill surfaces. AS well as binding to the general gill surface, Cu may enter the gill though low affinity channels. Copper disrupts NaC uptake at the basolateral Na pump. The stability constant for Cd is higher than for Cu: if surface binding was solely responsible for metal reactions at the gills, Cd would be expected to bind less stro~agly than Cd. Uptake of Cd though high affinity Ca channels is likely the explanation of the higker-thm- expected K ~ d - ~ d l . Also illustrated is the inhibition of the basolateral @a pump by Cd.

about a one log unit decrease in stability constants as pH decreases by one unit. However, we used the approach of @Him et al. (1991) and Cabaniss and Shuman (1988a) where changes in stability constants are considered to be due solely to H+ competition, That is, the use of one mead-gill stability constant, with another H+ stability constant for competition, as opposed to different metd-gill constants for each pH. In spite of the qualifications and simplifications used in the model, relative amounts of Cu and Cd deposition on the gills were reasonably well predicted (Fig. 5 6 ) . Copper was complexed better by DOC than was Cd, and Cd binding to the gills was affected more by Ca than was Cu.

We calculated approximately 0.05 kmol metal binding sites per rng DOC. Reported number of Cu binding sites for DOC vary from about 10 kmol per mg DOC (Cabaniss and Shuman 1988a) to 0.2 pmol per mg DOC (Hering and Morel 1990). Our low number of binding sites is indicative of the low metal concen- trations we used in our experiments. The sites were relatively strong metal binding sites (log Kc ,,_,, , = 9.1, log kl,, -,,, = 7.4), which may represent metals binding with carboxyl groups (Taga et al. 199 I). Sunda and Hanson (1979) determined log Hil,u~I,,c = 9.0 for river water at pH 5.95, very similar to our Cu result. Hering and Morel (1990) used 0.0% pM Cu concentrations in their experiments, and found slightly stronger Cu binding to humic acid (log KCu.hum,c acid = 10.1). Holm and Curtiss (1990) used 0.05, 8.2, and 10 pM Cu in their experiments, and found log K= 10.1,8.5, and 5.5, respectively, for natural organic matter in ground water, demonstrating the dependence of stability constants on the mead concentration used in their determination. Grirnm et al. (1991) determined (mean) log Kc,-,,, = 4.2-4.9, a

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much lower stability constant than hose listed above, but they used experimental Cu concentrations 2 5 pM. Holm and Curtiss (1990) determined that complexes with natural organic matter should dominate Cu speciation at the low Cu concentrations they used. In agreement, our measured and modelled gill Cu values (Fig. 6 ) were mainly influenced by DOC concentration.

Cadmium has been shown to bind less well than Cu to DOC and fulvic and humic acids (e.g., Florence 1977; KemdorR and Schnitzer 1980; Alberts and Giesy 1983; Lund et al. 1990; Sahu and Banerjee 1990). This binding order is related to general metal-ligand binding strength. If Cu binds better to DOC and humic acids, then Cu toxicity would be expected to decrease in response to organic material more than does Cd toxicity (e.g., Winner 1985,1986 for Cu, Cd toxicity to Daphnkup~le~x). In our work, freeze-dried DOC protected against Cu deposition on minnow gills at concentrations 24.8 mg * L-' (Playle et al. 1993), and insignificant Cu accumulation was observed in natural lake waters containing 1 6 mg DOC L-' (Fig. 61, whereas Cd still bound to fish gills in the presence of 2 6 mg DOC L-' (Pig. 5). Because of these binding results, we suggest that toxicity of the metal mixtures in soft waters (Table 4) was due mainly to Cd.

In the model we ignored differences in DOC source or com- position, because concentration of DOC adequately determined gill Cu concentrations (Playle et al. 1993). Although DOC from different sources can bind or detoxify metals to varying degrees (e.g., Nor and Cheng 1986; Sahu and Banerjee 19901, DOC from similar sources (i.e. same size of watershed and vegetation types) generally binds metals in a simraila- manner (Oliver et al. 1983; Cabaniss and Shuman 1988b). Source of DOC may not be as important as pH, alkalinity, or ionic strength in determining metal binding to DOC (Cabaniss and Shuman 1988b), and these workers suggested that experimental effort should be expended in defining these effects, not on documenting variations in DOC binding properties with season or location.

Synthetic soft water plus 10 gJLP9.1 (total) citric, glutamic, and csxalic acids did not reduce Cu or Cd accumulation on fathead minnow gills, nor did it alter metal toxicity (Table 4). This mixture was meant to be a synthetic DOC analogue, similar to but simplified from those used by Sposito (1981) and by Campbell and Stokes (1985). Campbell and Stokes (1985) found that a 22 pM DOC analogue complexed -70% of a 6 pg Cu * L-I solution (pH 6). With 3 X that mount of Cu in our experiments, and half the ligmd concentration, only about 12% of the Cu would be expected to be complexed, not enough to reduce Cu accumulation on gills. However, the model did calculate some reduction in free Cu (to about 60% of total; Fig. 6). It is not surprising that the 10 pM (total) ligand solution did not reduce Cu or Cd bound to gills or reduce metal toxicity, a d it does not adequately reflect 5 mg DOC . E-' in our system.

The model should be expanded to include other metals. It may be possible to calculate stability constants from previously published ligand work, but some metal-gill stability constants have already been published. For example, Wilfinson et al. (1990) used sdmonid data to estimate Hog KA,_,,,, - 6.5. Con- ditional stability constants for organic material and Al are available: log KA,-,, = 6.2 (Urban et d. 1990), log KA~~f,,l,lc ,,, = 7.8 (Shuman 1992). Constants are also available for other metals such as Mn (log KMl,_mc = 3.8; Urban et al. 1990), Fe (log KFc-,, = 9.4; Urban et al. 1990), and Co (log Kco-humlc acid = 5-6; Van Loon et al. 1992). Weaker complexing metals such as Ni and Zn would likely bind to DOC about 1OX less well than

does Cu (Morel 1983). Lin and Benjamin (1992) indicated some of the confounding effects of ligand addition to a multiple-metd system. A ligand may complex a metal that is otherwise bound to a (gill) binding site, which would free that site (e.g., increase free gill sites) to which another metal may bind. Models such as ours will be useful to predict complex interactions of this type.

In an expanded model, competition for gill binding sites by Ca and H+ need to be considered for each metd using appropriate log K values. Metals will be less toxic in hard water, and very acidic conditions will keep metals off gills. Of course, H+ itself can be disruptive to gills. Simulations using our model indicated that H+ would protect against metal binding at gills before it would displace metal from DOC. By the time enough metal was displaced from DOC for the metal to interact at the gills (pH < 4), H+ itself would likely become toxic. This point illustrates the usefulness of the model in integrating multiple effects such as speciation, competition, and complexation, a situation where models relying on correlation (Yan et al. 1998; Meador 1991) are often inappropriate.

Output from the present model is toxicant bound to the gills, the target organ, which then must be translated into toxicity to the fish. This last step requires the most study. Together, the model output would give an indication of d l types of competition for gill binding sites, of which Ca interactions can be considered beneficial, and most others as potentially toxic. In our lake water experiments, measured and calculated gill Cu and Cd concen- trations did correlate with toxicity, but it is not known how much metal accumulation is necessary to affect fish survival on a long-term basis. Presumably, a constant accumulation (gill "dose") will be toxic, such as was found for whole-body Cd accumulation in Hyalekla aztecn! (Borgmann et d. 199 1).

In sumary , measurement of metal deposition on minnow gills has allowed us to estimate conditional stability constants of Cu and Cd binding to gills, the number of metd binding sites on the gills, plus interactions with Ca, H+, and DOC. This infomation, and simi%ay information for other metals, will be useful in predictive models for metal accumulation on gills, and will therefore be useful to predict metd toxicity to fish.

We thank DHS. Chris Wood and Gord McDonald, McMaster Uni- versity, HM%ton, Ontario, for the use of their graphite furnace for metal analyses. We thank Dr. David Lean for help collecting lake water samples. This research was funded by an Operating Grant (No. 8155) from the Natural Sciences and Engineering Research Council of Canada to D.G. Dixon. by a Research Grant from the Borset Research Centre, Ontario Ministry of the Environment, to D.G. Diwon, and by an E.B. Eastburn Postdoctoral Fellowship to R. Playle.

References

ALBERTS. J.J., AND J.P. GHESY. 1983. Conditional stability constants of trace metals and naturally occu~ing humic materials: Application in equilibrium models and verification with field ddta, p. 333-348.111 R.F. Chistman and E.T. Gjessing 1ed.l Aquatic and terrestrial humic materials. Ann Ahor Science, Ann Arbor7 MI.

BHESINGER, K.E.. m~ G.M. CHRISTENSEN. 8972. Effects of various metals on survival, growth, reproduction, and metabolism of Daphalia mkagna. J. Fish. Res. Board Can. 29: 1691-1700.

BORGMAYN, U., W.B. NOWOOD, . ~ N D I.M. BABIBAD. 1991. Relationship between chronic toxicity and bioaccumulation of cadmium in Hyulellu aztecka. Can. J. Fish. Aquat. Sci. 48: 1055-1060.

CAB.~NISS, S.E., m~ M.S. S ~ I I I M A N . 1988a. Copper binding by dissolved organic matter: I. Suwannee River fulvic acid equilibria. Gemhim. Cosmochim. Acta 52: 185-193.

Can. 9. Fish. Aquut. Sci., Vol. 50, I9%3

Can

. J. F

ish.

Aqu

at. S

ci. D

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oade

d fr

om w

ww

.nrc

rese

arch

pres

s.co

m b

y C

ON

CO

RD

IA U

NIV

on

04/0

8/13

For

pers

onal

use

onl

y.

Cmmrss, S.E., AND M.S. SAUMAN. 198%. Copper binding by dissolved organic matter: IT. Variation in type and source of organic matter. Geochim. Cosmochim. Acta 52: 195-200.

CAAIPHIELL, P.G.C., aha, P.M. STOKES. 1985. Acidification and toxicity of metals to aquatic biota. Can. J. Fish. Aquat. Sci. 42: 2034-2049.

BALY, H.R., M.J. JONES, B.T. WART, AND 1.C. CAMPBELL. 1990. Copper toxicity toParufya austmliensis: In. Influence of dissolved organic matter. Environ. Toxicol. Chem. 9: 101 3-1018.

DEWIT, J.C.M., M.M. NEDERLOF, W.H. VAN WHEMSDLBK, AYD L.K. KOOPAL. 1% 1. Determination of H9 and metal ion affinity distributions for humic substances. Water Air Soil Pollut. 57-58: 339-349.

FISH. W.. AND F.M.M. MOREL. 1983. Characterization of organic copper- complexing agents released by Dapknia magna. Can. J. Fish. Aquat. Sci. 40: 1270-1277.

FLORENCE, T.M. 1977. Trace metal species in fresh waters. Water Research 11: 68 1-687.

GWM, DM., L.V. ~ZABRAGA, L.A. C A ~ I R A , AND W. SUSETYO. 1991. Con- tinuous multiligand distribution model nsed to predict the stability constant of Cu(lI) metal complexation with humic material from fluorescence quenching data. Environ. Sci. Technol. 25: 1427-143 1.

HERING, J.G., AND F.M.M. MOREL. 1990. Kinetics of trace metal complexation: Bigand-exchange reactions. Environ. Sci. Technol. 24: 242-252.

HOLM, T.R., AND C.D. CURTBSS 111. 1990. Copper complexation by natural organic matter in ground water, p. 508-518. In D.C. Melchior and R.L. Bassett Led.] Chemical modeling of aqueous systems II. American Chemical Society, Washington, K.

KERNBORPT. H., AND M. SC~ITZER. 1980. Sorption of metals on humic acid. Geochim. Cosmochim. Acta 44: 1701-1708.

CAU&N, B.J., AND B.G. MCDONALD. 1984. Influence sf water hardness. pH, and alkalinity on the mechanisms of copper toxicity in juvenile rainbow trout, S a h o gairdneri. Can. J. Fish. Aquat. Sci. 43: 1488-1496.

LAU&N, D.J., AND B.G. MCDONALD. 1987. Acclimation to copper by rainbow trout, Sdmo goirdneri: biochemistry. Can. J. Fish. Aquat. Sci. 44: 105-1 11.

%IN, 6.-F., AND M.M. BENJAMIN. 1992. The effects of strongly complexing ligands on the adsorptive partitioning of metal ions. Water Res. 26: 397-407.

Lm, H., AND D.J. RANDALL. 1990. The effect of varying water pH on the acidification of expired water in rainbow trout. J. Exp. Biol. 149: 149-160.

LLJND, W., I.A. HELEAK, AND H.M. SEIP. 1990. Studies of the complexation properties of aquatic humic material by differential pulse polarognphy. Sci. Total Environ. 92: 269-28 1.

MEADOR, J.P. 1991. The interaction of pH, dissolved organic carbon, and total copper in the determination of ionic copper and toxicity. Aquat. Toxicol. 19: 13-32.

M~RHL, P.h&M,B983.Priplcifles_of aquatic_chemistry- John Wiley and Sons, Toronto, ON. 446 p.

NOR, Y.M.. AND H.H. CBENG. 1986. Chemical speciation and bioavailability of copper: uptake and accumulation by Eichomia. Environ. Toxicol. Chem. 5: 941-947.

OLIVER, B.G., EM. THU~BAV, AND R.L. MALCOLM. 1983. The contribution of lumic substances to the acidity of colored natmal waters. Geochim. Cosmwhim. Acta 47: 203 1-2035.

PAGENKOPF, G.K. 1983. Gill surface interaction model for trace-metal toxicity to fishes: role sf complexation, pH, and water hardness. Environ. Sci. Technol. 17: 342-347.

PART, P., AND R.A.C. LOCK. 1983. Diffusion of calcium, cadmium and mercury in a mucous solution from rainbow trout. Comp. Biochern. Physiol. 76C: 259-263.

PERDUE, E.M., ~ V D C.R. LYTLE. 1983. A critical examination of metal-ligand complexation models: application to defined multiligand mixtures, p. 295-313. Zn R.F. Christman and Egil T. Gjessing [ed.] Aquatic and terrestrial humic materials. Ann Arbor Science, Ann Arbor, MI.

PLAYLF;, R.C.. AND C.M. WOOD. 1989. Water chemistry changes in the gill micro- environment of rainbow trout: experimental observations and theory. J . Comp. Pkysiol. B 159: 527-537.

~ A Y L E , R.C.. R.W. GENSBMER, AND D.G. DIXON. 1992. Copper accumulation on gills of fathead minnows: influence of water hardness, complexation and pH of the gill micro-environment. Environ. Toxicol. Chem. 1 1 : 38 1-39 1.

RAYLE, R.C., D.G. DIXON, AND K. BURNISON. 1993. Copper and cadmium binding to fish gills: modification by dissolved organic carbon and synthetic ligands. Can. J. Fish. Aquat. Sci. 50: 2667-2677.

REID, S.D., AND D.G. MCDONALD. 8988. Effects of cadmium, copper, and low pH on ion fluxes in the rainbow trout, Salrno guirdneri. Can. J. Fish. Aquat. Sci. 45: 244-253.

REID, S.D., AND D.G. MCBQNALD. 1991. Metal binding activity of the gills sf rainbow trout (Oncorhynchles mykiss). Can. J. Fish. Aquat. Sci. 48: 1061-1068.

SAHU, S., AND D.K. BANENEE. 1990. Complexation properties sf typical soil and peat humic acids with copper(I0 and cadmium(1I). Int. J. Environ. Anal. Chem. 42: 35-44.

SCHECHEW, W. 199 1. MLNEQLf version 2.1. Environmental Research Software. SHUMAN, M.S. 1992. Dissociation pathways and species distribution of alumi-

num bound to an aquatic fulvic acid. Environ. Sci. Technok. 26: 593-598. SPOSITO, G. 1981. Trace metals in contaminated waters. Environ. Sci. Technol.

15: 396-403. SIJNDA, W.G., AND P.J. HANSON. 1979. Chemical speciation of copper in river

water: effect of total copper, pH, carbonate, and dissolved organic matter, p. 347-180. Zn E.A. Jenne Led.] Chemical modeling in aqueous systems. ACS Symposiuin Series 93, American Chemical Society, Washington. DC.

TAGA, M., S. TANAKA, AND M. RIKUSHMA. 1998. Evaluation of copper(II)- binding ability of humic acids in peat using sulphopropyl-sephadex C-25 cation exchanger. Anal. Chim. Acta 244: 28 B -287.

TVSCBALL, S.R. JR., AND P.L. BREZONIK. 1983. Complexation of heavy metals by aquatic humus: a comparative study of five analytical n~ethods, p. 27 5-294. Zn R.F. Christman and E.T. Gjessiny [ed.] Aquatic and terrestrial humk materials. Ann Arbor Science, Ann Arbor, MI.

URBAN, N.W., E. GORHAM, J.K. U ~ ~ E K W O O D , F.B. MARTIN, AND J.G. OGDEN III. 1990. Geochemical processes controlling concentrations of Al, Fe, and Mn in Nova Scotia lakes. Limnol. Qceanogr. 35: 151 6-1534.

VAN DEN BERG, C.M.G., AND J.W. KUMER. 8979. Conditional stability constants for copper ions with liganh in natmal waters, p. 11 5-1 32. Hn E.A. Jenne [ed.] Chemical modeling in aqueous systems. American Chemical Society, Washington, DC.

VAN DEN BERG, C.M.G., P.T.S. WONG, AND Y .K. CHAU. 1979. Measurement of complexing materials excreted from algae and their ability to ameliorate copper toxicity. J. Fish. Res. Boad Can. 36: 901-905.

VAV COON, L.R., S. GWANACHER, AND H. HARDLTF. 1992. Equilibrium dialysis- ligand exchange: a novel method for determining conditional stability constants of radioaruclide-humic acid complexes. Anal. Chim. Acta 248: 235-246.

VERBOST, P.M., J. VAN ROOU, G . FLIK, R.A.C. COCK, AND S.E. WBNDBLAAR BONGA. 1989. The movement of cadmium through freshwater trout branchial epithelium and its interference with calcium transport. J. Exp. Bid. 145: 185-197.

WICKLUND, A., AND P. RUNN. 3988. Calcium effects on cadmium uptake, redistni- bution, and elimination in minnows, Plzoxi~us phmimm, acclimated to different calcium concentrations. Aquat. Toxicol. 13: 109-122.

WILKINSON, K.J., P.G.C. CAMPBELL, AND P. COUTURE. 1990. Effect of fluoride complexation on aluminum toxicity towards juvenile Atlantic salmon (Sal~no ,~alnP-). Can. J. Fish. Aquat. Sci. 47: 1446-1452.

WINNER, W.W. 1985. Bioaccumdation and toxicity of copper as affected by inter- actions between humic acid and water hardness. Water Res. 19: 449 -455.

WINNER, R.W. 1986. Interactive effects of water hardness and kutnic acid on the chronic toxicity of cadmium to Daphnl'apleEex. Aquat. Toxicsl. $: 28 1-293.

Xm. H.-B., AND L. SIGG. 1990. Binding of Cu(1I) to algae in a metal buffer. Water Res. 24: 2 129-1 136.

YAV, N.D., G.L. MACKIE, AND P.J. DILLON. 1990. Cadmium concentrations of crustacean zooplarakton of acidified and wonacidified Canadian shield lakes. Environ. Sci. Technol. 24: I 365-1 352.

ZITKO, V., AND W.G. CARSON. 1976. A mechanism of the effects of water hardness on the lethality of heavy metals to fish. Chemosphere 5: 299-303.

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