The bioavailability of copper in wastewater to Lemna minor with biological and electrochemical measures of complexation

Wat. Res. Vol. 28, No. 12, pp. 2457-2467, 1994

Pergamon

0043-1354(94)E0084-J

Copyright © 1994 Elsevier Science Ltd Printed in Great Britain. All rights reserved 0043-1354/94 $7.00 + 0.00

THE BIOAVAILABILITY OF COPPER IN WASTEWATER TO L E M N A M I N O R WITH BIOLOGICAL A N D ELECTROCHEMICAL MEASURES OF COMPLEXATION JAMESA. BUCKLEY* Department of Civil Engineering, University of Washington, Seattle, WA 98195, U.S.A. (First received June 1993; accepted in revised form March 1994) Abstraet--Lemna minor (duckweed) was grown in treated domestic wastewater containing added Cu to study the relationship between complexation and bioavailability. Growth was not inhibited until total Cu exceeded 0.079-0.119 mg/1 or internal Cu exceeded 81 to 235/~g/g. Cu2÷ was detectable in the wastewater when total Cu measured ~>0.378mg/1. There was a significant (P < 0.05) relationship between Cu bioconcentration and total Cu in wastewater that was attributed, in part, to heterotrophy of duckweed. Results indicated that, in addition to the bioavailable Cu2+ species, some complexes of Cu formed with ligands in wastewater may also be bioavailable. Measurement of the Cu Complexing Capacity (CC) of the wastewater by Ion-SelectiveElectrode (ISE) gave values of 0.26 to 0.29 mg/1and indicated the presence of ligands with a continuum of binding strengths which eliminated two of three commonly used methods for calculating CC. The biologically-measured CuCC based on growth was 0.077-0.125mg/1 when calculated by two out of three measures of Effective Concentration (EC) and was two to three times more sensitive than the ISE method for measurement of CC. ECs based on tissue concentration of Cu rather than solution concentration were 81-471 #g/g and have been proposed as an alternative for work in complex solutions like wastewater. Key words--wastewater, copper, duckweed, Lemna, bioavailability, bioconcentration, complexation, toxicity, metals

INTRODUCTION Metals in natural waters are distributed among a number of physico-chemical forms or species by reactions with dissolved organic and inorganic ligands, colloids and particulate matter (Florence and Batley, 1980). These forms together make up the total concentration for a given element. Organic ligands in model wastewater (Morgan and Sibley, 1975) and samples of treated wastewater (Stiff, 1971) may account for most of the complexation of the Cu. However, inorganic ligands may provide a kinetically labile source of Cu that contradicts their smaller representation of the total Cu concentration. It is estimated that 40-50% of the total organic carbon of treated sewage is represented by humic and fulvic acids (Manka et al., 1974) in the ratio of 1:10, respectively (Lapin and Krasyukov, 1986). The remaining groups included eight fatty acids, proteins, carbohydrates, anionic detergents and tannins and lignins. Compounds identified in wastewater used in this study include humic substances, seven fatty acids, probably originating from hydrolysis of vegetable and animal fats, a number of *Present address: Metro Environmental Laboratory, 322 West Ewing Street, Seattle, WA 98119-1507, U.S.A.

low molecular weight acids from metabolism, including the Krebs-cycle acids and formic and lactic acids (Archer, 1975). In addition, trace amounts of serine, glutamic acid, glycine and alanine have been identified. Due to the widespread occurrence of domestic and industrial wastewater discharges and their often complex nature, the effect they have on the forms and bioavailability of metals is of special interest. In addition to speciation in wastewater, some of the processes that determine metal bioavailability are the reactivity of the form with biomembrane ligands, the presence of other metals or major cations that act antagonistically or synergistically and the rate of diffusion, carrier-mediated transport and active transport of metals across biomembranes. This work utilizes the response of Lemna minor (duckweed) as a measure of the bioavailability of Cu in wastewater. It is a free-floating aquatic macrophyte that reproduces rapidly and is found in ponds and lakes throughout the world. Duckweed has been used to measure the toxicity of a number of pollutants including industrial (Wang and Williams, 1988) and wastewater treatment plant effluent (Taraldsen and Norberg-King, 1990) and metals in defined media (Benda and Kouba, 1991; Huebert et al., 1993) and in fly ash settling ponds (Clark et al., 1981).

2457

JAMES A. BUCKLEY

2458

T h e p u r p o s e o f these e x p e r i m e n t s w a s to d e t e r m i n e the e x t e n t t h a t c o m p l e x a t i o n o f C u b y ligands in s e w a g e t r e a t m e n t p l a n t effluent influences the bioavailability o f Cu. This w a s c a r r i e d o u t by e x a m ination of relationships between growth and Cu u p t a k e by d u c k w e e d a n d t h e C u 2+ a n d total C u c o n c e n t r a t i o n a n d C u c o m p l e x i n g c a p a c i t y o f the wastewater. METHODS

Experimental procedures Duckweed (Lemna minor) was cultured at 25°C in Hoagland's medium (ASTM, 1988) modified by the omission of sucrose, yeast extract and bactotryptone. Cultures were maintained axenic by treatment with HOCI (Hillman, 1961). For each experiment, approximately 23 1 of final effluent was collected by 24-h composite or by grab sampling from a plant treating primarily domestic sewage by the activatedsludge process. The sample was dechlorinated, filtered through a 0.45 lam membrane and stored in polyethylene containers at 3.8-6.5°C for up to 14 days. The macronutrient content and other selected water quality factors in the collected samples is shown in Table 1. Wastewater was prepared for static-renewal experiments by adjusting the pH to 6.8 with HCI, aliquoting 100 ml into 250 ml beakers and adding CuSO4 to each aliquot to achieve an array of concentrations of Cu with two or four replicates each. Following incubation for 1 h at 25°C during which the pH typically increased to 7.0, each solution was inoculated with 6 plants. Four replicates o f the inoculum were measured for wet and dry weight and Cu content. The beakers were placed in a clear acrylic chamber using a random number assignment, to which CO2 was added to the headspace to maintain a pH of approximately 7.0 in the experimental solutions. The beakers were then incubated at 25_+0.5°C with a 24-h light period of 3100-37001x. Every 24h all plants were transferred to freshly-made experimental solutions. At the end of the 7-day experiments, the duckweed was rinsed in deionized water (DW), 10-2M EDTA and D W and then blotted dry. For measurement of Cu content and/or dry weight, the duckweed was placed in tared 50 ml beakers, weighed and dried for 20-24 h at 56-60°C and then reweighed and stored at - 2 5 ° C until analyzed.

Table 1. The content of selected nutrients, metals and the conductivity and total hardness of secondary-treated wastewater used in 3 experiments (mg/I or as noted) Experiment Measurement'

1

NO3-N (NH3 + NH4)-N NH3-N b Total PO4 T.R. PO~ Cd Cr Cu Ni Pb Zn

2

3

0.03 0.09 0.25 25.8 27.8 21.9 0.14 0.16 0.14 2.8 4.3 3.7 -3.0 2.3 < 0.002 < 0.002 < 0.002 <0.005 <0.005 <0.005 0.024 0.053 0.029 <0.01 <0.01 <0.01 < 0.03 < 0.03 < 0.03 0.050 0.048 0.053 T. Hardness~ 72 97 67 Conductivity ¢ 530 567 635 "Values (except NH3-N) from treatment plant records. bUnionized ammonia based on pH 7.0.

¢Total Reactive Phosphorus. dmg/I as CaCO3. C/~mhos/cm.

Preparation of samples for total copper analyses Duckweed was digested at 95°C in concentrated HNO 3 followed by 30% H202. The digestates were diluted to 25 ml with DW. Samples o f wastewater were prepared for determination o f total Cu according to standard EPA methodology (Anon, 1983). Three digestion blanks were included with each group of samples.

Measurement of copper by ICP Copper was measured by Inductively Coupled Argon Plasma-Atomic Emission Spectrometry (ICP) on a JobinYvon JY-50P instrument with a detection limit of 10 #g/l. Instrument drift (g = 0.006 mg/l, s = 0.012, n = 48) between standards positioned every 10 samples was corrected to zero drift. One replicate and one spike-recovery standard every 10 samples provided, respectively, a mean Relative Percent Deviation of 2.1 (s =2.5, n = 19) and mean Percentage Recovery of Cu of 101.5 (s = 3.8, n = 14).

Measurement of copper by 1SE The Cu 2+ concentration of experimental solutions was measured with an ion-selective electrode (ISE), Orion Model 94-29, together with a double junction reference electrode, Model 90-02. Inner and outer filling solutions were as recommended by the electrode manufacturer. The ionic strength of standards and wastewater were adjusted with 0.01 M NaNO3. Calibration standards (slope: g = 30.46, range = 29.46-31.80, ideal = 29.58) of CuSO4 in DW were plotted as millivolts (mV) vs Cu concentration on semi-log paper and then were fitted to a regression line. A 3-point standard (0.005, 0.05 and 0.5 mg/l) was analyzed immediately before and after measurements in wastewater. The good correspondence between these values insured that ISE performance was not compromised by coating during use in wastewater. The response was consistently Nernstian to 0.005 mg/1, a value which defined the practical limit of detection. Duplicate measurements showed a Relative Percent Deviation of 1.8% (s = 1.2%, n = 9). Measurements were made in 100 ml samples in polypropylene containers placed in a circulating water bath at 25 __+0.5°C. The electrode response was generally stable in between 1 and 10 min. The electrodes were shielded from sunlight. Mixing was by a d.c.-operated stirrer attached to the electrode arm. The electrodes were rinsed between measurements with 1 m M EDTA and DW to minimize cross-contamination of the samples and increase sensitivity o f the electrode.

Electrochemically measured copper complexing capacity Values were obtained by additions of Cu from 0 to 1 mg/1, with intervals of 0.1 mg/1, to 100 ml aliquots of a wastewater sample, followed by equilibration at 25°C for l h before measurement of the Cu 2+ concentration with an ISE. The titration data thus obtained were applied by three graphical methods to calculate complexing capacity (CC). Method 1 is a plot of Cu 2÷ vs added Cu and represents the basic titration data. The CuCC is defined as the x-axis intercept from the regression equation for the linear portion of a plot of instrument response (Y) vs added or total Cu (X). Methods 2 and 3 are transformations of the data from Method 1. In Method 2, the CC and K* are obtained from the slope and Y-axis intercept, respectively, o f a linear equation describing the relationship between free and complexed Cu (Hart, 1981). [Cu 2+ ]/[CuLl = (l/K*[CuCC]) + (1/[CuCCl)([Cu 2+ ]) (1) where: Cu 2+ = Cut = CuL = L =

cupric ion total Cu --- Cu 2+ + CuL bound Cu = Cu t - Cu 2+ free ligand = L t - CuL

2459

Bioavailability o f copper in wastewater

Water quality measurements

Lt = total ligand = L + C u L = C u C C (when Cu is in excess) C u C C = Cu complexing capacity K * = conditional stability constant

Total residual chlorine, measured amperometrically by the back titration method, and other water quality factors except total organic carbon (TOC) (Anon, 1983) were measured according to A P H A (1985).

In Method 3, the C C and K* can be derived from binding parameters from the Scatchard plot (Rosenthal, 1967). [b ]/[u] = K*n [M ] - K*[b ]

Statistical analysis

(2)

where [b ] = [u] = K* = n =

concentration of bound Cu concentration of free Cu (Cu 2÷) intrinsic association constant n u m b e r o f equivalent, noninteracting binding sites per molecule [M ] = concentration of the ligand n [ M ] = complexing capacity (CC) n [M ] - [b ] = concentration of free binding sites.

Biologically measured copper complexing capacity Biological measurements of CC define the concentration o f Ca-reactive ligands in a water based on an organism's response. Most methods for biological CC have been based on the sigmoid growth curve model in which a response such as growth is plotted as a function of concentration of metal. In this work, the sigmoid-shaped curve has been linearized by transformation of the response variable. The concentration of metal that causes a 50% response (EC50), the effective concentration at the 5% level o f response (EC5) and the N o Observed Effect Concentration (NOEC) were used to define the CC. The EC5 and N O E C define concentrations o f Cu at which low, and perhaps insignificant, levels o f growth inhibition has occurred. Cu in excess of this defined CC would be bioavailable and expected to significantly inhibit growth.

Bioconcentration factor The Bioconcentration Factor (BF) is defined as the ratio of the concentration of Cu in duckweed (Ca) to that in the wastewater solutions (Cww) when the concentration in duckweed is at steady-state

Cd/Cww = BF.

(3)

Dry weight (for NOEC), Cu bioconcentration and Cu BF were analyzed by A N O V A and D u n n e t t ' s or Tukey multiple comparisons. Data sets that did not pass tests for normality or for homogeneity o f variances prior to A N O V A were transformed to logt0 or were analyzed by the Kruskal-Wallis nonparametric A N O V A . The 7-day EC50 and 7-day EC5 and associated 95% confidence limits were calculated using simple linear regression followed by inverse prediction. The biological and electrochemical measures of C u C C were tested for significant difference by two-tailed Student's t-test which was preceded by variance ratio F-test for equality of variances.

RESULTS

The bioavailability o f copper in wastewater T h e r e l a t i o n s h i p b e t w e e n t o t a l C u a n d C u 2+ c o n c e n t r a t i o n s in w a s t e w a t e r a n d C u b i o c o n c e n t r a t i o n a n d g r o w t h in d u c k w e e d w a s e x a m i n e d in 3 e x p e r i m e n t s ( T a b l e 2). T h e d a t a s h o w t h a t as t h e level o f total Cu increased, there was a proportional increase in C u u p t a k e a n d i n h i b i t i o n o f g r o w t h . C u u p t a k e and depuration reached steady-state within 5 days. When total Cu reached a mean of 0.185mg/l, or m o r e , g r o w t h as m e a s u r e d b y d r y w e i g h t w a s signific a n t l y ( P < 0.05) d i m i n i s h e d relative to p l a n t s g r o w n in w a s t e w a t e r w i t h o u t a d d e d C u ( c o n t r o l s ) . A t a m e a n t o t a l C u c o n c e n t r a t i o n o f 0.397 mg/1 (0.4 m g / l a d d e d ) , o r m o r e , t h e r e were m e a s u r a b l e a m o u n t s ( > 0 . 0 0 5 mg/1) o f C u 2÷ p r e s e n t w h i c h m a y a c c o u n t

Table 2. The mean (n = 3)total Cu and mean (n = 2) Cu 2+ content (mg/l) of solutions of secondary-treated wastewater in which duckweed was grown for 7 days in three replicate experiments

Experiment No.

Test solutions Cu 2+ (mg/I)

Added Cu (nominal) (mg/l)

Total Ca (mg/l)

0h

0 0.05 0. I 0.2 0.4 0.8 0 0.05 0.1 0.2 0.4 0.8 0 0.05 0. I 0.2 0.4 0.8

0.024 0.073 0.113 0.206 0.378 0.741 0.022 0.072 0.119 0.222 0.405 0.796 0.029 0.079 0.128 0.225 0.409 0.809

ND c ND ND ND 0.010 0.046 -ND ND ND 0.010 0.045 -ND ND ND 0.015 0.058

12 h

-ND ND ND 0.009 0.045

Duckweed 24 h

Dry WP (mg)

Cu ab (#g/g)

---ND 0.007 0.038 --ND ND ND 0.016

74.8 75.4 73.7 56.0* 12.6" 6.3* 66.3 64.1 62.3 38.0* 12.0" 5.8* 65.5 65.8 57.8* 55.5* 12.1" 3.7*

53 84* 136* 303* 579* 1614" 62 133" 235* 473* 616" 1959" 29 81 * 159" 327* 532* 2972*

aMean of n = 2 (Experiment 1) and n = 4 (Experiments 2 and 3). bBased on dry weight. CNot detected ( < 0.005 mg/l). *Significant (P < 0.05) vs control (0 mg/I added Cu), Dunnett's ANOVA.

JAMES A. BUCKLEY

2460

for the accompanying marked decrease in dry weight. Based on the increase in Cu bioconcentration when total Cu averaged only 0.075 mg/l (0.05 mg/l added), forms of Cu in addition to Cu 2+ may have been bioavailable. In combination, these proposed forms were bioconcentrated to as much as 136, 235 and 81/~g Cu/g (dry wt) without significant (P > 0.05) impairment of growth. As the number of plants increased daily, there was a pattern of decreased alkalinity (from 136 to 72 mg/1 as CaCO 3 on day 7) and increased dissolved oxygen (9.7 to 12.3 mg/l on day 7) in the wastewater that reflected the pattern of duckweed growth and therefore was most pronounced at the lower levels of Cu (~<0.225 mg/1). The Cu content of the duckweed was significantly (P = 0.000) related to the total Cu content of the test solutions (Fig. 1). The coefficient of determination adjusted for number of independent variables (R 2. adjusted) was 93.5% indicating that 93.5% of the variation in Cu bioconcentration was "explained" by total Cu in the wastewater test solutions. As Fig. 1 shows, bioconcentration of Cu (/~g/g) is correlated with the concentration of total Cu over the range of 0.025--0.782 mg/l. However, the amount of Cu (/~g) taken up by duckweed is not a simple function of the concentration in the wastewater over the same range (Fig. 2). The Cu content increases to the point that growth is inhibited about 27% and then decreases with diminished growth. When growth is at a minimum the Cu content increases again slightly. Despite the uptake by duckweed, the content of total Cu in solution remained fairly constant over the 24-h period between renewals. Analysis of the solutions revealed an average 93% (s = 7, n = 7) of the initial measured concentration remained after 24 h. The mean (n = 10) BF varied from 90 to 259 and from 1379 to 2889 when based on wet and dry weight, respectively, for duckweed grown in waste4 -Log Y= 3.32 + 1.11 l o g X

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Log Cu (mg/l)

Fig. 1. The relationship between the log of Cu content of duckweed and the log of total Cu concentration in the secondary-treated wastewater in which the plants were grown for 7 days.

~

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"e (0,075)(0.120)(0.218)(0,397) (0.782) .

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Fig. 2. The relationship between the Cu content (open circles) and mean (n = 10) dry weight (solid circles) of duckweed grown for 7 days in the indicated concenations of Cu added to wastewater. Values in brackets are decoded from log X. water with mean (n = 9 ) total Cu contents of 0.025-0.782 mg/l (Table 3). The BFs were statistically similar (P > 0.05) for all but the highest level of Cu treatment (0.782 mg/1) when based on dry weight and all but the two highest levels (0.397 and 0.782 mg/1) when based on the wet weight. Use of dry weight to express Cu bioconcentration by duckweed markedly increases the BF over that based on wet weight. Figure 3 shows plots of the ISE-measured Cu 2÷ concentrations resulting from addition of Cu to wastewater used in the three experiments shown in Table 2. The CuCCs computed by Method 1 are given in Table 4. The electrode response shown in the plots a, b and c in Fig. 3 was transformed according to equation (1) and plotted as solid dots in Fig. 4 in order to calculate CuCC by Method 2. The plot was divided into lines 1 and 2 and from the respective regression equations the CuCC and log K*, defined in equation 1 were calculated and recorded under Method 2

Weight basis (0.025)

30

°

--

t

O (0.025)

-

0 0

Table 3. The mean (n = 10) bioconcentration factors for duckweed exposed for 7 days to the indicated concentrations of Cu in wastewater

O

45 E

Total Cu (mg/I) 0.025 0.075 0.120 0.218 0.397 0.782

Wet 127 ( _ 4 8 ) ~ 90(_+23) 98(-+22) 113(+25) 164(-+26)* 259 (+41)*

Dry 1972 ( _+862) 1380(_+416) 1528(-+399) 1727(-+351) 1436(_+140) 2889 ( -+842)*

~One standard deviation. *Significant (P < 0.05) vs control.

Bioavailability of copper in wastewater

2461

Table 4. The Cu complexing capacity (CuCC, mg/I) and log stability constant (K*) calculated by 3 methods for secondary-treated wastewater used in the three experiments shown in Table 2 Method I

Experiment No.

CuCC

1 2 3

0.29 0.27 0.26

2

Line 1 CuCC Log K* 0.49 0.40 0.47

7.26 7.59 7.56

3

Line 2 CuCC Log K* 1.23 1.35 1.52

CuCC

Log K*

NC" NC NC

NC NC NC

6.21 6.17 5.99

~Not calculated, see text.

in Table 4. Owing to the nonlinearity of the plotted points (dotted lines), it is apparent that attempted division of the curve into two linear portions representing, respectively, stronger and weaker ligand classes, is arbitrary. This raises doubts regarding the

validity of this method for the wastewater used in this work. The ISE measurements shown in Fig. 3 were again transformed, in this case by means of equation (2)

0.07 - (a)

0.08 - -

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Added Cu (mg/l) Fig. 3. Measurement of the Cu Complexing Capacity (CuCC) of secondary-treated wastewater measured as the x-axis intercept of the regression line for the 6 data points in the linear portion of each plot (r = 0.999 for all plots). WR

28/12--O

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7 I

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0.5

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Cu 2+ (BM) Fig. 4. Plots for the determination of Cu Complexing Capacity (CuCC) and stability constant (K*) for Cu added to secondary-treated wastewater as determined from the slope and intercept of the solid lines according to equation (l).

2462

JAMESA. BUCKLEY

and then tested by Scatchard Plot analysis according to Method 3 (Fig. 5). Nonlinearity of the plots indicates multiple binding sites with regions of strong (a) I

150

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[ [

100 [b] [u] 50

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primary binding, intermediate strength binding and weaker binding represented by the upper, middle and lower portions of the curves, respectively. The data in the lower part of the curve do not intercept the x-axis. Furthermore, there are no distinctly linear regions in the upper or lower portions of the curve to permit extrapolations to the x-axis that would permit graphical estimation of CuCC. Preparation of solutions that would represent data points beyond those in the lower portion of the curves requires addition of >/1.0 mg Cu/1 of wastewater. This typically results in nonlinearity beyond the regression lines shown as the extreme right data points in Figs 3 and 4 and has been interpreted as exceeding the solubility of Cu under these conditions. These results indicate that data obtained with the wastewater used in this work were not amenable to analysis for CuCC by the Scatchard Plot method as is indicated under Method 3 in Table 4. Table 5 shows CuCC based on the biological response of growth and on electrochemical technique using the ISE with measurements calculated by Method 1. When based on total Cu in the external solution, the values for NOEC (£ =0.104) and EC5 (~ = 0.102) are not significantly (P > 0.05) different. The same is true for the ECS0 (£ = 0.280) and ISE (£ = 0.273). However, the values for NOEC and EC5 are significantly (P < 0.05) different from those for EC50 and ISE. The statistical analysis indicates that under the conditions of these experiments the ISE was an acceptable analytical surrogate for the biologically-measured ECS0, but not for measurements by NOEC or EC5. When based on the content of Cu in duckweed, the values for NOEC were too variable to be useful. However, the variability in values for EC5 and EC50 based on Cu in solution [coefficient of variation (CV) = 24 and 8%, respectively] and internal Cu (CV = 18 and 7%, respectively) was similar. The water quality measurements of pH, dissolved organic carbon (DOC) and alkalinity (Table 5) are important in determining Cu speciation due to the ligand sources they represent which markedly affect the CuCC of a water. Similarities in pH, DOC and alkalinity in all 3 experiments is primarily due to constancy of these factors among the wastewater samples. The uniformity of water quality measurements and maintenance of pH near 7.0 (range 6.84-7.29) through the 7-day experiments was considered fundamental in providing uniform Cu speciation and resulting biological responses. DISCUSSION

I

0 0

5

I

I

I0

15

b (BM) Fig. 5. Scatchard/Rosenthal plot for the binding of Cu 2+ by ligands in secondary-treated wastewater according to equation (2). The data points are fitted by observation.

Bioeoncentration of copper Passage of Cu from solution into cells can be accounted for by diffusion and active transport mechanisms that have been identified in plants in general and, in some cases, in duckweed specifically. The free

Bioavailability of copper in wastewater

2463

Table 5. The Cu complexingcapacity of secondary-treated domestic wastewater based on biologicaland electrochemical measurements in the three experiments shown in Table 2 BiologicaP Cu in solutionb Cu contentc Experiment No. NOECd EC5c EC50r NOEC EC5 EC50 I 0.113 0.125 0.294 136 181 415 (0.2294).069) (0.514-0.168) (235-126) (466-363) 2 0.119 0.077 0.253 235 163 471 (0.123--0.049) (0.393-0.163) (292-34) (595-348) 3 0.079 0.104 0.293 81 125 416 (0.218-0.050) (0.593--0.145) (335--0) (620-211) Electrochemical Water quality 1SEg pHh DOCi AIk) 1 0.29 6.8915 135 7.09 2 0.27 6.9315 138 7.10 3 0.26 6.9913 138 7.12 aGrowth of duckweed for 7 days in concentrations of Cu added to wastewater, measured as dry weight. bExpressed as Cu in wastewater, mg/1. CExpressed as Cu in duckweed, #g/g. dNo observed effectconcentration. CEffectiveconcentration for 5% growth inhibition (upper-lower 95% C.I.). reflectiveconcentration for 50*/0growth inhibition (upper-lower 95% C.I.). Slon-SelectiveElectrode, mg Cu/l; data from Method 1. hRange. iDissolved Organic Carbon, mg/l. JTotal alkalinity, mg/I a s C a C O 3.

aquo ion, Cu(H20)~ + , referred to for simplicity as Cu 2+ , may enter by active transport (Pitman and Luttge, 1983). Cu 2÷ was measurable (>t 0.005 mg/1) in solution when the total Cu values were t>0.378 mg/I (Table 2). At lower concentrations, availability of Cu 2÷ for active transport was subject to the kinetics of transfer between ligands in wastewater and membrane carrier molecules. In both cases, uptake was limited by rates of transfer to, within and from duckweed cells. From the data in Fig. 2, when the total Cu in wastewater was ~<0.218mg/1, uptake appeared to be a function of the external concentration indicating that adsorption and diffusion processes were important. Between 0.218 and 0.397 mg/i growth was markedly diminished and Cu uptake became inversely related to the concentration of Cu in wastewater and directly related to growth. The total amount taken up was regulated by factors other than the ambient concentration and was not diffusion controlled. Active transport may have controlled Cu flux. From 0.397 to 0.782 mg/I uptake again appeared to be controlled by adsorption and diffusion processes. Cu, administered as CuSO4, was taken up by duckweed in a dose-dependent manner that was controlled by the medium-to-cell concentration gradient, whereas Cu coordinated in the neutral b/s (glycinate)--Cu complex was taken up by direct diffusion followed by intraeellular degradation of the toxic complex that controlled the tissue burden of Cu (Benda and Kouba, 1991). In this work, Cu bioconcentration occurred in a dose-dependent manner (Fig. 1), similar to that reported for exposure to CuSO4 and was not "self-limiting" as in the case of bis (glycinate)-Cu complex. However, owing to the

variety of ligands in wastewater, and their different sites of transport into duckweed, complexed Cu may also have contributed to the total burden of Cu reported here. In the present work, an undefined portion of the tissue burden of Cu may be due to uptake of Cuamino acid complexes. This reasoning is supported by the occurrence of amino acids in wastewater, including that from the source used in this work, the stable complexes they form with Cu 2+ (Stiff, 1971), and by their transport systems identified in duckweed (Datko and Mudd, 1985; Jung et al., 1982). The dose-dependent uptake of Cu recorded in this work may also be accounted for, in part, by active transport of ligands, other than the amino acids, that are known complexors of Cu 2+ . For example, Cu 2+ binds to N of guanine and forms chelates between guanine and phosphate (Sissoeff et al., 1976). This purine comprises a portion of humic substances (Lapin and Krasyukov, 1986) and as such may occur in wastewater where it is available for uptake by duckweed via a discrete transport system (Datko and Mudd, 1985). The discrete transport system reported for aldohexoses (Datko and Mudd, 1985) provides for speculation on a role for this class of compounds as carrier ligands during Cu uptake. It has been reported that Cu had a more marked effect on inhibition of frond multiplication when sucrose was present in the medium (Nasu and Kugimoto, 1981), implying that sucrose facilitated the toxicity of Cu. Hydrolysis of sucrose yields glucose and fructose which are aldohexose and ketohexose sugars, respectively. The aldohexoses were reported to be weak complexors of Cu 2+

2464

JAMESA. BUCKLEY

(Haas, 1986) and no information of a similar nature was available for ketohexoses, indicating that evidence for a ligand function for these sugars is not strong. It is worth noting that glucose and another aldohexose, mannose, have been identified in wastewater from the source used in this work (Archer, 1975) however it is not known if they had a similar effect on the Cu toxicity. The possible uptake of Cu that was complexed to organic compounds is also supported by the heterotrophic nature of duckweed (Hillman, 1961) and by the good relationship between Cu bioconcentrated and total Cu of the medium (Fig. 1). Cu 2÷ (Table 2) and CuCC (Table 4) measurements indicate that Cu is highly complexed in the wastewater, as has been reported previously (Stiff, 1971; Morgan and Sibley, 1975), and that total Cu values in that relationship primarily represent complexed forms of Cu. This suggests that in wastewater, total Cu can be a predictor of the bioavailability of Cu. However, from the preceding discussion as well as from literature sources (Morrison, 1989) Cu bioavailability is apparently due to a unique speciation and not simply to the total amount of Cu that is present. Due to heterotrophy, duckweed is apparently susceptible to forms of Cu not available to organisms that are solely autotrophic. After exposure for 41-46 days to 0.032 and 0.32 mg Cu/l in filtered ditch water, duckweed contained 150 and 300 pg Cu/g (dry wt), respectively (van der Werff and Pruyt, 1982). Duckweed inhabiting a fly-ash settling basin with a Cu concentration of 0.10 mg/l contained 250#g/g (dry wt) (Clark et al., 1981). These values are within the concentration range of 48-576 pg/g (dry wt) of tissue resulting from growth in water at 0.025-0.397 mg Cu/l, which were found in this study. This relationship occurred despite the differences in water sources, which supports the notion of bioavailability of a range of Cu species such as must have existed in these three studies. A similar conclusion regarding bioavailability of Cu species was reached when data from another study using ditch water was compared with results cited above. In that work, exposure to 1.3, 3.2, and 6.4mg Cu/l resulted in uptake of 1280, 3840 and 8960 pg Cu/g (dry wt) despite a pH of 8.0 +__0.2 in the water which presumably resulted in inorganic complexation of much of the added Cu (Ernst and van der Werff, 1978). Duckweed collected from 23 ponds and lakes in Southern Ontario with a mean concentration of 0.035 mg Cu/1 contained a mean 33 ttg Cu/g (dry wt) (Hutchinson and Czyska, 1975) which is slightly below the ranges for Cu bioconcentration despite a mean water concentration that is within the range. This may be accounted for by the selection of only healthy-looking fronds for Cu analysis in the work by Hutchinson and Czyska. The loss of alkalinity from wastewater commensurate with growth of duckweed is an outcome of the preference for fixing aqueous inorganic carbon, prob-

ably HCO~-, over CO2 from air (Filbin and Hough, 1985). This preference was demonstrated under an atmosphere enriched with CO2 to control pH. As expected, the dissolved oxygen values show a pattern inverse to that of alkalinity indicating photosynthetically produced 02. It is apparent that 02 is released to the aqueous phase as all vessels were open under the chamber and yet there were differences in dissolved oxygen corresponding to growth. Release of 02 to the aqueous phase has important implications for maintaining desirable levels of dissolved oxygen under conditions of abundant growth of duckweed in natural- and waste-waters. Electrochemical measurement o f copper complexing capacity

Copper complexing capacity, also known as Cu binding capacity or Cu uptake capacity, is a measure of these Cu-reactive ligands expressed as the Cuequivalent. Complexing capacity (CC) measurements have been stated to be: information on the ability of water to accept and detoxify metal ions (Florence and Batley, 1980), part of the total CC due to complexing centers already occupied by the strongly bound metals AI3+ and Fe 3+ (Linnik and Nabivanets, 1980) and the total concentration of potentially available metal binding ligands (Hart, 1981). There have been a considerable number of CuCC studies of diverse types of waters, however, only a small number of studies have involved work in treated sewage. The CuCC value of 0.55 mg/l obtained by ASV polarography (Blutstein and Extort, 1983) using Method 1 is higher than values of 0.256-0.320 mg/l (Buckley, 1983) and 0.26--0.29 mg/1 (Table 4) obtained by ISE. This probably results from differences in wastewater because the ASV measurements of CC are generally acknowledged to be lower due to a variable amount of complex destruction at the electrode surface (Florence and Batley, 1980). No references were available showing curved plots following transformations by Method 2 as occurred in this work (Fig. 4). Plots with one linear region indicating a single ligand class with uniform binding strength give values of 82 #g/g for sewage sludgeamended soil (Karapanagiotis et al., 1991) and 0.127mg/1 for activated sludge solids (Sterritt and Lester, 1985). Addition of Cu to fulvic acid isolated from sewage sludge resulted in definition of two ligand classes with CuCC values of 0.16 and 0.31 mg/1 for stronger binding ligands and values of 1.40 and 1.40mg/1 for weaker binding ligands (Sterritt and Lester, 1984). These values are comparable to those found in this work, however, owing to the arbitrary fit of the straight (solid) lines denoting ligand classes Lj, and L2, values for Method 2 (Table 4) provide a comparison only and were not considered definitive values. A detailed analysis of the Scatchard plot (Method 3) using the relationship of [b]/[u] vs [b] for CuCC was precluded by lack of linear regions (Fig. 5).

Bioavailability of copper in wastewater Curved Scatchard plots also have been reported for the binding of Cu on the surface of algal cells and nonlinearity was interpreted as a decrease in binding tendency with an increase in surface coverage as Cu 2÷ was first coordinated to the highest affinity surface ligands and increased surface charge induced some electrostatic repulsion (Xue et al., 1988). Biologically-measured copper complexing capacity The biologically-measured CuCC, reported as EC50s, for wastewaters used in this study varied from 0.253 to 0.294mg/1 (Cu in Solution, Table 5). Although no comparable studies with wastewater were available, the EC50 values may be compared with those obtained in defined medium using duckweed and in natural waters for which the values are reported as biologically-measured CuCC using bacteria and algae (Table 6). The values for duckweed in defined media and wastewater are higher than those for algae in natural waters which may be accounted for by: (1) the presence in defined media of Fe chelators such as EDTA (PAA, Gorham medium) and citric acid (B-D medium) which, if in excess of Fe, will also chelate Cu; (2) the presence in wastewater of chelators of natural and anthropogenic origin, although, as noted above, numerous organic compounds, including chelators, are available to duckweed (Datko and Mudd, 1985; Jung et al., 1982) through heterotrophy; (3) more of the toxic species of Cu were present in the natural waters used in the referenced work; and/or (4) duckweed is less sensitive to Cu than the other species listed. Duckweed may be less sensitive to Cu than the other species listed in Table 6 because of the magnitude of the difference in EC50s between duckweed and the other species and the very low concentrations of total Cu at the EC50 for the other species. The reasons for the presumed difference in sensitivity are difficult to identify, because no single mechanism accounts for Cu toxicity, and plants have multiple mechanisms to accommodate such toxicity. However, differences in P availability and status may

2465

have been a factor in the difference in sensitivity of the organisms. All duckweed are in a defined medium or wastewater (Table 6) which is rich in nutrients and luxury uptake of P can lead to formation of polyphosphate bodies which are important to sequestering otherwise toxic levels of metals (Morrson, 1989). Conversely, all the algae and bacteria listed in Table 6 were tested in freshwater or seawater which was presumably lower in nutrients and some algae have been reported to be more sensitive to Cu toxicity under conditions of P limitation (Hall et al., 1989). The ISE-measured CuCC, when analyzed by Method 1, was a good estimator (P > 0.50) of the biologically-measured CuCC when EC50 was the endpoint (Table 5). In order to provide comparable water quality and Cu speciation conditions for the comparison, both the biological and electrochemical measurements were made under conditions of similar pH, temperature and equilibration time. However, when Thalassiosira was grown in Cu-spiked, filtered and ultrafiltered water the CuCC based on the change in EC50 was correlated with, but was approximately twice, the CuCC measured by ASV (Srna et al., 1980). This finding was attributed to the complexation of Cu by high MW compounds that were not ASVlabile. The ISE measurements show that for all three experiments, Cu 2+ was below the detection limit of 0.005 mg/1 when the concentration of total Cu was at the EC50. In addition to 50% growth inhibition, there was significant (P < 0.05) bioconcentration of Cu which suggests that forms of Cu in addition to the free aquo ion (Cu 2+ ) may have been bioavailable. The NOEC and EC5 measures of CuCC based on Cu in solution were not significantly (P > 0.50) different from each other, but were both significantly (P < 0.001) lower than the ISE CuCC measurements (Table 5). They represent statistical estimates of a concentration of Cu that elicits no observed effect at a specified level of ct and that effects a 5% reduction in growth, respectively. Although the EC5 specifies

Table 6. The concentration of Cu effecting 50% response (EC50) or yielding a median response at the equivalence point (EP) on a dose-response plot, in freshwater, seawater, or prepared medium for the indicated organisms Ca

Measure EC50 EP EC50 EC50 EP EC50 EP EC50 EC50 EC50 EC50

(ttg/I)

Medium

Organism

References

8-13" 17~ 1-13 ~ 0-144 a 3~,0 ~ 8" 8.5-304 1100 100 130 253-294a

Freshwater Seawater Seawater Seawater Seawater Seawater Brackish PAA b~ B-D ~ Gorham b Wastewater

Selenastrum Thalassiosira Thalassiosira Thalassiosira

Allen et aL (1983) Davey et al. (1973) Imada et aL (1985) Srna et al. (1980) Gillespie and Vaccaro (1978) Anderson (1983) Newell and Sanders (1986) Wang (1986) Nasu et aL (1984) Jenner and Janssen-Mommen (1989) This work

Bacterium Bacterium Thalassiosira

Duckweed Duckweed Duckweed Duckweed

~Reported by authors as biologically-measured CuCC. bDefined medium. CProvisional algal assay. dBonner-Devirian.

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JAMESA. BUCKLEr

an amount of adverse effect while the NOEC appears to define a no-effect concentration, the selection of an • level also specifies a level of adverse effect in the case of the NOEC. In addition, the NOEC is sensitive to the concentration interval chosen for testing and therefore varies in precision inversely with the width of that interval. The EC5 and NOEC proved to be the best (relative to the EC50) measure of the mitigation of Cu toxicity to duckweed by wastewater and of duckweed to accommodate a certain burden of Cu. The mean of 0.103 mg/l for the two measures corresponds to uptake of about 170/~g Cu/g (dry wt) which caused only minimal reduction in growth. In comparison, at EC50 concentration there is sufficient bioavailable Cu to reduce growth by 50%. Consistent with the values for EC5 and NOEC (Table 5), is the conclusion that biological measures of CuCC generally give smaller values than those by ISE, Ion Exchange or ASV. The concentration of total Cu at which growth was inhibited was one-half to one-third of the value at which Cu 2÷ was detected by the ISE. That is, the biological method was 2-3 times more sensitive than the electrochemical method at detecting available Cu. That may be explained if the biological "probes" were influenced only by Cu 2÷ plus weakly bound forms of Cu (Hart, 1981) that are not measurable by electrochemical means. Or it could be that Cu that is complexed is still bioavailable by nature of the availability (uptake) of the ligand as appears to be possible for duckweed. The measurement of EC50 based on the internal concentration of Cu has been suggested as a means to measure Cu toxicity that bypasses relationships with Cu speciation in the external solution (Huebert et al., 1993) and thereby avoids the difficulty in relating toxicity (or bioconcentration) to solutions that are undefined with respect to the species of metal present and their individual bioavailabilities. This approach combines elements of toxicity and bioconcentration tests and should be especially useful for bioavailability studies with aquatic plants where uptake is from the bulk solution as opposed to studies with fish that also consider a variable dietary uptake. In this work, EC50's based on internal content of L e m n a minor averaged 434/zg Cu/g (dry wt) and were less than the 1300-1600pg Cu/g (dry wt) reported for L e m n a trisulca exposed to Cu-EDTA solutions (Huebert et al., 1993). However, more data are needed to determine if, as proposed, this means of expressing toxicity is related to organism tolerance (or dose) and is an improvement over traditional methods using the ambient concentration. The expression of toxicity based on the total internal concentration of Cu can be subject to some of the same well-known difficulties as is that expression based on total external concentration. For example, toxicity of Cu can vary with nutritional status of the organism, specifically the P status (Hall et al., 1989)

and formation of Cu-chelating polyphosphate bodies (Morrison, 1989) and with the amount of metalsequestering phytochelatins present (Reddy and Prasad, 1990). Definition of metal toxicity based on internal concentrations would be aided by knowledge of the internal partitioning between sensitive sites and storage forms. REFERENCES

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